Environmental Biotechnology and Engineering – 2014

Transcripción

Environmental Biotechnology
and Engineering - 2014
Volume 2
ISBN - 978-607-9023-29-4
Poggi-Varaldo, H.M.; Bretón-Deval, L.M.;
Camacho-Pérez, B.; Escamilla-Alvarado, C.;
Escobedo-Acuña, G.; Hernández-Flores, G.;
Muñoz-Páez, K.M.; Romero-Cedillo, L.;
Sastre-Conde, I.; Macarie, H.; Solorza-Feria, O.;
Ríos-Leal, E.; Esparza-García, F.
Poggi-Varaldo, H.M.; Bretón-Deval, L.M.;
Camacho-Pérez, B.; Escamilla-Alvarado, C.; EscobedoAcuña, G.;
Hernández-Flores, G.; Muñoz-Páez, K.M.; RomeroCedillo, L.;
Sastre-Conde, I.; Macarie, H.; Solorza-Feria, O.; RíosLeal, E.; Esparza-García, F.
“Environmental Biotechnology and
Engineering – 2014”
Volume 2
ISBN - 978-607-9023-29-4
México D.F., México, 2014
Are property and responsibility of Authors.
All or any part of this publication may be reproduced or transmitted, by any
means, electronic or mechanical (Including photocopying, recording or any
recovery system and storage), and must be included with the corresponding
citation of this compendious and their authors.
Environmental Biotechnology and Engineering – 2014
Editors
Héctor Poggi Varaldo, Beni Camacho Pérez, and others.
D.R. © This Edition Centro de Investigación y de Estudios Avanzados del I.P.N.
Cinvestav
2014
Publisher
Bonumedia
Amores 1166-4
Col. Del Valle
Del. Benito Juarez
CP 03100
CD version
400 copies
ISBN Vol. 2: 978-607-9023-29-4
ISBN Complete: 978-607-9023-27-0
Printed in Mexico
November 28th 2014
Environmental Biotechnology and Engineering - 2014
Content
Page
How to cite an article/chapter of this book
Preface
Section 1. Renewable and Alternative Energies and Biorefineries
Section 2. Sustainability and Environmental System Analysis
Section 3. Risk Assessment and Environmental Impact
Section 4. Air Pollution and Climate Change
Section 5. Aquifer Remediation
Section 6. Soil and Sediment Remediation
Section 7. Wastewater Treatment
Section 8. Solid Waste Management and Treatment
Section 9. Hazardous Waste Management and Treatment
Section 10. Environmental Toxicology
Section 11. Microbial Ecology
Section 12. Molecular Biology Applications to Environmental Problems
Section 13. Control and Modelling of Environmental Processes
Section 14. Environmental Chemistry
Section 15. Environmental Health
Section 16. Environmental Nanotechnology
Section 17. Miscellaneous
iii
iv
1
254
374
434
475
539
700
914
994
1060
1092
1249
1320
1410
1440
1458
1491
i
Content of Volume 2
Page
How to cite an article/chapter of this book
Preface
Section 6. Soil and Sediment Remediation
Section 7. Wastewater Treatment
Section 8. Solid Waste Management and Treatment
Section 9. Hazardous Waste Management and Treatment
iii
iv
539
700
914
994
ii
How to cite an article of this book
For example, the chapter by Oscar H. Ortiz-Méndez; Leopoldo J. Ríos-González;
José A. Rodríguez-de la Garza; German Aroca-Arcaya HQWLWOHG ³CHAPTER 1.1.
ETHANOL PRODUCTION FROM ENZYMATIC HYDROLYSATES OF Agave
lechuguilla PRETREATED BY AUTOHYDROLYSIS´published in the pages 5 to 13 of
this book, should be cited as follows:
Ortiz-Mendez, O.H.; Rios-Gonzalez, L.J.; Rodríguez-de la Garza, J.A.; Aroca-Arcaya, G.
(2014). Chapter 1.1. Ethanol production from enzymatic hydrolysates of Agave
lechuguilla pretreated by autohydrolysis. In: Poggi-Varaldo, H.M.; Bretón-Deval, L.M.;
Camacho-Pérez, B.; Escamilla-Alvarado, C.; Escobedo-Acuña, G.; Hernández-Flores,
G.; Muñoz-Páez, K.M.; Romero-Cedillo, L.; Sastre-Conde, I.; Macarie, H.; Solorza-Feria,
O.; Ríos-Leal, E.; Esparza-García, F. (Editors): Environmental Biotechnology and
Engineering ± 2014, Volume 1, pages 5-13. Ed. Cinvestav, Mexico D.F., Mexico.
iii
Preface
Environmental Biotechnology and Environmental Engineering are two faces of a
modern, valuable, and indispensable scientific and technical coin. The growing
significance and awareness of environmental problems, caused especially by use of
fossil resources in connection with industrial pathways of production, depletion of finite
natural resources, mismanagement of renewable resources, etc., have led to the
development of both disciplines. They have their own historical roots, i.e., one has
blossomed from Biotechnology and the other has grown from the old Civil and Sanitary
Engineering. Yet, they have developed in full fledged branches of knowledge and
specialization, and at the same time they complement each other.
Regarding Environmental Biotechnology, its contributions span from
environmentally-IULHQGO\ DQG FRVW HIIHFWLYH ³HQG-of-the-SLSH´ VROXWLRQV WR HQYLURQPHQWDO
pollution and problems (bioremediation of soils and aquifers, biological waste treatment),
to the development of sustainable alternatives for their prevention and alleviation, such
as the replacement of fossil fuels by biohydrogen and methane from wastes and
IXWXULVWLF³ELRUHILQHULHV´%LRWHFKQRORJ\KDVWKHSRWHQWLDORIDUHGXFWLRQRIRSHUDWLRQDODQG
investment costs for the design and operation of more sustainable processes based on
microbes and other living organisms as agents. Yet, so far the sustainability of technical
processes is more the exception than the rule. In this regard, Environmental
Biotechnology is a serious candidate to provide substantial advances in the near future
On the other hand, Environmental Engineering has developed several significant
fields of research and applications (everything matters in Environmental Engineering;
natural sciences and social sciences are as significant to its practice as classical
engineering skills); some of them partially overlap with Environmental Biotechnology (for
instance, biological waste treatment), whereas other subjects are original and cover
issues that Environmental Biotechnology can not, and have proved to be of use to other
branches of knowledge. With respect to this, we would like to highlight a significant
contribution of Environmental Engineering that has trascended to other fields of
Engineering and Technology: sound Environmental Engineering has designed the
imprescindible framework of System Engineering Analysis applied to environmental
issues, also known as Life Cycle Analysis (LCA) and other denominations. The
contemporary history of industry and technology has sadly taught us that new
technological solutions and new processes derived from Environmental Biotechnology
(and from other fields of knowledge) should be examined under the light of LCA and
environmental impact analysis before attempting their implementation. Very often, a
precipitated and immature application of a new product or process has led to adverse
impacts on health and the environment that have become technical, ethical and
economic burdens to modern societies.
The synergistic interaction of Environmental Biotechnology and Environmental
Engineering has a tremendous potential for making outstanding contributions to the
sustainable development and sustainable management of resources in modern
societies. To a great extent, we expect that these contributions will also positively impact
RQ VRFLHWLHV¶ RUJDQL]DWLRQ DQG LPSURYH SHRSOH¶V FRQVFLHQFH HGXFDWLRQ DQG KDELWV
iv
Sustainable development should become the basis for the life of future generations as
opposed to over-exploitation of non-renewable energy and material resources.
In 2003, a group of pioneering biotechnologists in Mexico led by Dr. Hector M.
Poggi-Varaldo, Dr. Fernando Esparza-García and Professor Elvira Ríos-Leal,
accompanied by a constellation of international scientists such as Dr. Isabel SastreConde from Spain, Dr. Hervé Macarie from France, Dr. Franco Cecchi and Dr. Paolo
Pavan from Italy, Dr. E. Foresti from Brazil, Dr. Irene Watson-Craik from Scotland, Dr.
Jose Luis Sanz from Spain, and others, identified a gap in the dissemination of both
Environmental Biotechnology and Environmental Engineering. This was particularly true
for developing countries, although the situation in developed countries was not much
better.
On the one hand, there were several international and regional events dealing with
Biotechnology but no international event was devoted to Environmental Biotechnology.
At most, Environmental Biotechnology has one or two sessions in a Biotechnology
Congress. On the other hand, most regional Environmental Engineering events showed
a strong commercial component that negatively competed with the exchange of
advanced knowledge and the formation of research networks. Moreover, Environmental
Biotechnology and Environmental Engineering are two dynamic drives with a strong
interaction, and the scientific community could obtain several advantages from their joint
diffusion. In short, there was a need for an international event dedicated to both
disciplines, with a strong vocation for serious dissemination of scientific and
technological knowledge, as well as research networking.
The synthesis to this diagnostic was to launch a new event focussed on both
disciplines. In this way, the First International Meeting on Environmental Biotechnology
and Engineering was born and held in 2004 in Mexico City. This first event was coorganized by the Dept. of Biotechnology and Bioengineering of CINVESTAV del IPN in
Mexico, the IRD of France, the IMIA from Spain, the Mexican Polytechnic Institute (IPN)
from Mexico, the National University of Mexico (UNAM, México), the University of
Hidalgo (UAEH, México), among others. The event was backed-up by a diverse
International Scientific Committee that had the contributions of outstanding scientists
and professionals from Brazil, Italy, Spain, Scotland, France, and Mexico.
After the Second International Meeting on Environmental Biotechnology and
Engineering also held in Mexico City, Mexico, in 2006, we had the satisfaction to see
that the 3rd International Meeting on Environmental Engineering held in Palma de
Mallorca had exponentially grown and matured. Its outreach was multiplied by a factor of
10 compared to that of the 1st IMEBE. The Organizing Committee led by Dr. Isabel
Sastre-Conde and Dr. Hervé Macarie should be congratulated for the success and
resonance of the third version of this event. This fact is a confirmation of the original
diagnostic: the scientific community was avid of an international event with the
characteristics of the IMEBE now ISEBE.
Indeed, the name of the event has been changed from Meeting to Symposium, in
order to reflect the increases on both quantity and quality. So, in 2014, the name of the
event is the Fourth International Symposium on Environmental Biotechnology and
Engineering.
This book entitled Environmental Biotechnology and Engineering-2014 in three
volumes, contains the edited articles of the contributions presented in the 4ISEBE and it
v
is both a reference and a reminder. It is a reference of fine research and works on
Environmental Biotechnology and Environmental Engineering, for personal and Library
consultation, since several copies of the books will also be distributed among the main
Universities of the countries that have participated in the event. Furthermore, the book is
a reminder of the efforts that we should still make in order to improve our environment
and quality of life, as well as the commitment in further continuing the dissemination and
exchange of these efforts in the upcoming 5th ISEBE.
We want to acknowledge all authors of the works presented in the 4ISEBE. Also,
we express our gratitude to the support to 4ISEBE from our alma mater the CINVESTAV
del IPN and its Department of Biotechnology and Bioengineering, CONACYT (Mexican
Council of Science and Technology of Mexico), the Institute de Recherche et
Developpement and IMBE from France, the American Chemical Society from the USA,
la Fundacion Semilla from the Baleares Islands, Spain, the Mexican Society of
Biotechnology and Bioengineering (SMBB), the Mexican Association of Solar Energy
(ANES), the Mexican Society for Hydrogen (SMH), and a constellation of Mexican
private companies and Mexican higher education institutions, among others. Without
their varied contributions and support, the 4ISEBE would have not happened. We are
also very grateful to Ms Ana Lucía Castro-Ríos for her excellent work in the production
of the CD-ROM books of 4ISEBE. Finally, we are very grateful to the members of the
Scientific Committee who have evaluated the articles published in this book.
We look forward to meeting all of you and as well as a stream of new participants
in the next 5th ISEBE in 2016.
Professor Dr. Héctor M. Poggi-Varaldo
vi
Section 6.
Soil and Sediment Remediation
539
Chapter 6.1. Response of barley plants in an arsenic-polluted soil treated
with nanoscale zero valent iron (nZVI)
S. Diez-Pascual; M.C. Lobo; A. González; J. Alonso; E. Rodríguez-Valdés;
M. Gil-Díaz
542
Chapter 6.2. Microorganismos fijadores de nitrogeno de vida libre con
capacidad de degradación de clorpirifos
David A. Moreno-Medina; Julián Delgadillo-Martínez;
David Espinosa-Victoria; Rosa M. López-Romero;
Ma. Remedios Mendoza-López; Oscar García-Barradas
560
Chapter 6.3. Isolation and characterization of arsenic-resistant endophytic
bacteria associated with plant growing on mine tailings
Brenda Román-Ponce; Salvador Embarcadero-Jiménez;
María S. Vásquez-Murrieta; En Tao Wang
574
Chapter 6.4. Acumulación de Hg y as por Brassica rappa L. inducida por
agentes quelantes y tratamientos nutricionales cultivada en residuos mineros
Juan A. Flores-de la Torre; Rubí E. Luna-Sandoval;
Guillermo Rodríguez-Hernández; Eduardo Manzanares-Acuña
590
Chapter 6.5. Effect of hydrocarbons on germination and phenolic profile of
in vitro cultures of Cyperus laxus
Morales Navarro Sarahí Itzel; Rivera Casado Noemí Araceli;
Mónica S. Salazar-Plata; Octavio Gómez-Guzmán;
Josefina Pérez-Vargas; Graciano Calva-Calva
601
Chapter 6.6 Biolixiviacion de metales de sedimentos anaerobicos del rio
reconquista (Argentina) como estrategia potencial de remediación
Natalia Porzionato; Roberto Candal; Gustavo Curutchet
610
Chapter 6.7. Comparación entre bioestimulación y bioaumentación para la
biorremediación de suelos contaminados con hidrocarburos
Diana Cabañas-Vargas; Alma I. Corona-Cruz; Rafael Rojas-Herrera
623
Chapter 6.8. Biodegradación de atrazina en suelos agrícolas del estado de
Morelos México
Ana G. Ortega-Hernández; Luis G. Treviño-Quintanilla;
Carlos Díaz-Gómez; Rosa A. Guillén-Garcés
Chapter 6.9. Efecto del 1,2,3-triclorobenceno en la biodegradación de un
suelo contaminado con hexaclorociclohexano en un reactor de suelos
activados en ambiente desnitrificante
Gonzalo M. Escobedo-Acuña; B. Camacho-Pérez; E. Ríos-Leal;
M. T. Ponce-Noyola; J. Barrera-Cortés; Héctor M. Poggi-Varaldo;
H. B. C. Poggi-Rinderknecht; N. Rinderknecht-Seijas
632
643
Chapter 6.10. Comparison of methods for fitting kinetic data of inhibition of the
biodegradation of a main pollutant for remediation assessment
540
Gonzalo M. Escobedo-Acuña; B. Camacho-Pérez; E. Ríos-Leal;
M. T. Ponce-Noyola; J. Barrera-Cortés; Héctor M. Poggi-Varaldo;
H. B. C. Poggi-Rinderknecht; N. Rinderknecht-Seijas
653
Chapter 6.11. Bioremediation of a soil contaminated with lindane in an
electrobiochemical slurry reactor: role of surfactant and characterization
of microbial community
Beni Camacho-Pérez; Jaime García-Mena; Omar Solorza-Feria;
Elvira Ríos-Leal; Selvasankar Murugesan; Alberto Piña Escobedo;
Héctor M. Poggi-Varaldo
666
Chapter 6.12. Potential crude oil degradation by a microbial consortium isolated
from the Gulf of Mexico
Ulises García-Cruz; Rosa Canul-Puc; Juan Sanchez-Avila; Gerardo Gold-Bouchot;
Leopoldina Aguirre-Macedo
684
Chapter 6.13. Evaluation of hexadecane degradation using different marine
sediments from rivers and lagoons of the Gulf of Mexico
Ulises García-Cruz; Juan Sanchez-Avila; David Valdés-Lozano;
Gerardo Gold-Bouchot; Leopoldina Aguirre-Macedo
693
541
CHAPTER 6.1. RESPONSE OF BARLEY PLANTS IN AN ARSENIC-POLLUTED
SOIL TREATED WITH NANOSCALE ZERO VALENT IRON (nZVI)
S. Diez-Pascual (1); M.C. Lobo (1); A. González (1); J. Alonso (1);
E. Rodríguez-Valdés (2); M. Gil-Díaz*(1)
,0,'5$)LQFD³(O(QFtQ´$OFDOiGH+HQDUHV0DGULG6SDLQ
(2) Universidad de Oviedo, Mieres 33600 (Asturias), Spain
ABSTRACT
The use of nanoscale zero valent iron (nZVI) as a new tool to degrade and/or
immobilize a vast range of environmental pollutants in soil and water, such as
chlorinated organic compounds, metals, metalloids and inorganic anions has
resulted effective. The present study tests the effectiveness of using nZVI for in situ
immobilization of As in highly polluted soil from a brownfield site. The doses of 1%
and 10% of the commercial nanoparticle suspension significantly decreased the
availability of As in the soil, according to the results of the sequential extraction
procedure and the TCLP test. The As immobilization was more effective at 10%
than at 1% of nZVI commercial suspension. In view of these results, an experiment
in growth chamber in controlled conditions was performed to determine the stability
of the As immobilization with nZVI and the impact on a barley crop. Physicochemical soil parameters were also evaluated to determine the impact of this
nanoremediation strategy on soil properties. The lower availability of As in soils
treated at 10% led to lower uptake of As by barley plants which showed the highest
growth rate. Iron concentration in 10%-treated soils increased in CB, OX, OM and
RS fraction; Fe leachability initially increased moderately (< 2 mg/L), but a release
of Fe was not observed over the time studied. In turn, the application of nZVI on
the brownfield soil did not induce a higher uptake of Fe in barley plants. No
negative effects on soil physico-chemical properties have been observed at the
used nZVI doses (1% and 10% of the commercial suspension).The As
immobilization presented some instability after the barley crop. Thus, the proposed
nanotechnology could be proposed as an alternative for in situ remediation of Aspolluted soils and could be combined with remediation process where barley plants
are involved, although long-term monitoring studies are necessary to evaluate the
stability of the As immobilization with nZVI.
Key words: As immobilization, barley, brownfield, iron nanoparticles, nanoremediation.
Introduction
542
Soil is a living system, and a nonrenewable resource due to the slow speed of the
formation process. It plays an important role due to its capacity to store, filter, and
transform many substances, including water, nutrients, and carbon (Lobo et al.,
2009). Soil serve as a habitat for humans, animals, plants and soil organisms. It is
a medium for decomposition, balance, and restoration as a result of its filtering,
buffering, and substance-converting properties, especially for groundwater
protection (Miehlich, 2001). Soil is subject to a series of degradation processes or
threats. These include erosion, loss of organic matter, contamination (local and
diffuse), sealing, compaction, decline in biodiversity, salinization, floods, and
landslides (Lobo et al., 2009). Contaminants may have an anthropogenic or natural
origin. Since the 1970s, the increase in industrial processes and in waste treatment
systems has led to the appearance of different types of substances, including
heavy metals and organic compounds, that constitute contamination sources in
the soil and prevent the soil from performing its functions. Nowadays, soil
contamination is a global problem. European governments have considered that
soil degradation has a direct impact on water and air quality, biodiversity, and
climate change, and therefore consequences on human health.
Arsenic is a toxic metalloid which can have natural origins or anthropogenic
sources such as mining, burning of coal, smelting of non-ferrous metals, and from
pesticides used in agriculture. Arsenic is one of the 10 more toxic substances to
human health (Environment Agency, 2009). Plants respond to As-toxicity by
showing some physiological changes in their shoots and roots (Shaibur et al.,
2013). Toxicity, bioavailability and mobility of As highly depend on its chemical
speciation (Ruiz-Chancho et al., 2007). Inorganic As is more toxic and mobile than
organic arsenic (Smith et al., 1998; Larios et al., 2012). The availability and mobility
of As in soils depends on soil characteristics such as pH, Fe-oxides, phosphorus,
co-precipitation in salts, organic matter, clays, etc. (Moreno-Jiménez et al., 2012).
The available As in soils is the most ecologically relevant because it is readily
mobile and bioavailable within the environment and can be potentially taken up by
plant roots (Moreno-Jiménez et al., 2012; Mench et al., 2009).
Thus, the main objective of many in situ remediation strategies is to reduce the
mobile fraction of metals and metalloids in the soil which could reach the
groundwater or be taken by soil organisms (Kumpiene et al., 2006). In this sense,
iron compounds have been commonly used for As stabilization. Iron oxides and
their precursors (zero-valent iron and iron sulphates) have been extensively
studied for immobilization of As in polluted soils due to their important sorption
properties (Kumpiene et al., 2006; Kim et al., 2003; Hartley and Lepp 2008;
Kumpiene et al., 2008; Komárek et al., 2013). Adsorption and co-precipitation are
the stabilization mechanisms of As with iron oxides (Komárek et al., 2013). In
recent years, the reactivity of zero-valent iron has been increased by the
development of nanoscale zero-valent iron (nZVI), with smaller particle size and
larger surface area. The use of nZVI to remediate contaminated soil and
groundwater has received increasing amounts of attention within the last decade,
primarily due to its potential for broader application, higher reactivity, and costeffectiveness compared to conventional zero-valent iron applications and other in
543
situ methods (Karn et al., 2009; Grieger et al., 2010). For As remediation, several
studies have shown effective the use of ZVI or iron oxides nanoparticles in polluted
groundwater (Jegadeesan et al., 2005; Kanel et al., 2006; De et al., 2009) and soils
(Zhang et al., 2010; Shipley et al., 2011; An and Zhao 2012).
Nanoremediation with nZVI is an established treatment in the USA, whereas
only a few full-scale applications have been performed so far in Europe (Karn et
al., 2009; Mueller et al., 2012). The currently available ecotoxicology data of this
new remediation technology are limited. Thus, it is necessary to know its benefits
and risks and determine if that such technology restores soil functionality (Wang et
al., 2014). Recently, more attention has been paid in the evaluation of ecological
impact of nZVI on soil organisms. Result comparison is difficult because ZVI
nanoparticles are produced from different methods, consequently, show different
composition, size and reactivities (Ma et al., 2010).
Plants strongly depend on soil composition, and alterations in their growth may
reflect the presence of toxic substances (Gyuricza et al., 2010). El-Temsah and
Joner (2012) studied the effects of different concentrations of nZVI on seed
germination of flax, barley and ryegrass, and they found no significant effect at
concentration up to 250 mg/L nZVI. Gil-Díaz et al. (2014) concluded that the nZVI
suspension at 1% (0.14% Fe0) and 10% (1.4% Fe0) did not induce phytotoxicity in
the germination of barley and common vetch seeds. Wang et al. (2014) studied the
toxic effects of soil treated with nZVI (1.5 g Fe0 per kg of soil) on two crops, rape
and Chinese cabbage, and they found that both plant species were significantly
affected after 72 hours of nZVI application, but in a month the phytotoxicity tests
showed an improvement in cultivation for both plants. In a hydroponically
experiment performed with two plant species (cattail and hybrid poplars) for a
month, Ma et al. (2013) also observed toxic effects at nZVI concentrations > 200
mg/L. Here, we present a study about the immobilization of As in brownfield soil,
using nZVI, and its impact on barley plants development. The specific objectives of
the present study were to (1) investigate the effectiveness of nZVI to immobilize As
in brownfield soil highly polluted; (2) determine the impact of this nanoremediation
technology on a barley crop, as well as, the bioaccumulation of As and Fe in
plants; (3) evaluate the stability of the As immobilization after the growing period;
(4) determine the effect of the nZVI treatment on soil properties.
Materials and methods
Soil. Soil from a brownfield located in Asturias (north of Spain) was selected for
this study. In this site, arsenic ore processing and metallurgy were carried out
intermittently for decades until the end of the 1970s. This activity promoted dust
and steam emissions and the dumping of mining and smelting waste. Currently,
the distribution of the pollutants throughout the site is caused mainly by the
mechanical dispersion of the above mentioned waste, together with the oxidation
and lixiviation of As rich materials, and also the processes of complexation and
immobilization related to soil particles (Sierra et al., 2011).
544
Soil samples were collected from the surface layer (0±30 cm depth), air-dried
and sieved (<2mm) before analyses. Soil properties were determined according to
Spanish official methodology for soil analysis (MAPA, 1994), and are shown in
Table 1. Briefly, electrical conductivity (EC) and pH were measured in 1:2.5 soil-towater ratio; organic matter and total nitrogen content were determined using the
Walkley-Black and Kjeldahl methods, respectively; the percentage of carbonates
was measured using a calcimeter; available phosphorus was evaluated using
sodium bicarbonate at pH 8.5 according to the method proposed by Watanabe and
Olsen (1965); available nutrients (K+, Na+, Ca2+, Mg2+ y Fe2+) were extracted with
NH4Ac 0.1 N at pH 7, and assessed using flame atomic absorption spectrometry
(FAAS) (AA240FS, Varian, Victoria, Australia). The soil texture was analyzed using
a Bouyoucos densimeter; water holding capacity was determined using Richard
Plates at 0.33 and 1515 kPa; soil porosity was measured from bulk and real
density according the method proposed by Monnier et al., 1973. Iron, aluminum
and manganese oxides were extracted by the sodium citrate-bicarbonate-dithionite
method (Jackson et al., 1986). Iron and Mn in the extracts were analyzed by FAAS
(AA240FS, Varian), and Al was measured by Inducible Conductivity PlasmaAtomic Emission Spectrometry (Varian LibertyAX, Victoria, Australia).
Total concentration of As, Cd, Cr, Zn, Pb, Ni, Cu, Mn, and Fe in soil samples
were determined after acid digestion with a mixture of 6 mL of nitric acid (69%
purity) and 2 mL of chlorhydric acid (37% purity), in a microwave reaction system
(Multiwave 3000, Anton Paar GmbH, Graz, Austria). Arsenic was analyzed with a
Zeeman Atomic Absorption Spectrometer (AA240Z, Varian) equipped with a
graphite tube atomizer and a programmable sample dispenser, and the other
elements were quantified by FAAS (AA240FS, Varian). The As limit of quantitation
was 0.016 mg/L; for the rest of the element the limit of quantitation was in the
range of 0.01-0.10 mg/L. Standard As solution of As2O5 in water (1,000 mg/L) and,
for the other metals, nitrate solutions in nitric acid supplied by Panreac (Barcelona,
Spain) were used for calibration purposes. All analytical determinations were
performed in duplicate.
Zero-valent iron nanoparticles. A commercial stabilized water dispersion of zero
valent iron nanoparticles NANOFER 25S (NANO IRON Rajhrad, Czech Republic)
was used in this experiment. The Fe (0) percentage in the commercial product was
between 14-18%, according to the commercial specifications. Klimkova et al.
(2011) presented a complete characterization of the used commercial
nanoparticles.
Batch experiments. To test the nZVI effectiveness for the immobilization of As, 50
g of polluted soil were mixed with 0.5 g of the commercial nanoparticle suspension
and 20 mL of milli-Q water in a plastic vial of 100 mL (1% of the commercial nZVI).
The same procedure was applied at a higher dose of nanoparticles (10% of the
commercial nZVI), mixing 50 g of soil, 5 g of commercial nZVI and 15 mL of milli-Q
water. Three independent vials were used per treatment. The mixtures were
shaken for 72 hours at 100 rpm with a shaker Reax 2 (Heidolph Instruments GmbH
545
& Co. KG, Schwabach, Germany). Control tests were carried out in parallel with
the same amount of soil (50 g), 20 mL of deionized water without nZVI addition.
Plant experiment. An experiment in growth chamber in controlled conditions was
carried out to determine the stability of the As immobilization and the influence of
nZVI treatment on barley crop. Plastic pots (0.3 L) were filled with 130 g of soil
according to the following treatment: 0% nZVI (control); 1% nZVI; 10% nZVI;
unpolluted soil (positive control). Four pots were used per treatment. Seeds of
barley (Hordeum vulgare L. cv. Pedrezuela) were germinated on filter papers in a
Petri dish moistened with distilled water. The Petri dishes were firmly closed with
parafilm and then incubated at 25ºC in the dark. After 72 h, uniform seedlings were
transplanted into each pot (two seedlings per pot). The pots were placed in a
growth chamber at a 16-h photoperiod, 22:9ºC day: night temperatures and relative
humidity of 80:30% (day: night). The pots were watered with tap water during the
experiment. Plant height was measured periodically. Plants were harvested 33
days after transplanting. Roots were rinsed, firstly, with tap water thoroughly, and
secondly, with distilled water. Roots and shoots were then dried at 65 ºC for 72 h to
constant weight.
As, Fe and macronutrients determination in barley plants. Total As, Fe, Ca,
Mg, Na, K in roots and shoots were determined according to the method proposed
by Zhao et al. (1994). Dried plant sample (30 mg) was weighted in a 4 mL glass
vial, and 1 mL HNO3 65% and 1 mL HClO4 70% were added. After a day, vials
were placed in a Techne Dri-Block DB-3D (Camlab, Cambridge, UK) heater at 130
ºC for 2.5 h. After the vials were cooled, the solutions were filtered (Whatman 541)
and diluted to 10 mL with Milli-Q water. The As concentration was measured by
graphite furnace atomic absorption spectrometry and Fe, Ca, Mg, Na, K were
measured by FAAS.
Sequential extraction procedure. The availability of As in soils was analyzed
after applying the sequential extraction procedure proposed by Tessier et al.
(1979) with some modifications. Extractions with solutions of increasing strengths
were sequentially added to the soil sample. The relatively availability is:
exchangeable (EX) > carbonate-bound (CB) > Fe/Mn oxides-bound (OX)> organic
matter-bound (OM) > residual (RS). The As concentration was measured by
graphite furnace atomic absorption spectrometry in all the extracts obtained in the
sequential extraction procedure described below. Iron concentration was also
measured by FAAS to evaluate the impact of the nZVI treatment on the Fe
availability in the soil.
EX: 2.5 g of dried soil were weighed and transferred to 50 mL centrifuge tubes
where 25 mL of MgCl2 (1M, pH 7) were added. The tubes were vigorously shaken
at room temperature for 1 hour and then centrifuged at 13000 rpm for 30 min with a
Beckman (Fullerton, CA) centrifuge (J2-21). The supernatant was filtered through a
Whatman filter paper (nº 542) and made up to 25 mL. The extracts were stored at
4ºC until analysis.
546
CB: the residue from exchangeable fraction was mixed with 25 mL of buffer
CH3COONa/CH3COOH (1M, pH 5); the tubes were shaken at room temperature
for 5 hours and then centrifuged and treated at the same conditions described
above.
OX: the residue from carbonates fraction was mixed with 25 mL of NH2OH.HCl
(0.04M in acetic acid 25%); the tubes were shaken at 96ºC in a water bath for six
hours, and then centrifuged and treated at the same conditions described above.
OM: the residue from Fe/Mn oxides-bound fraction was mixed with 5 mL H2O2 30%
and 3 mL HNO3 0.01M; the tubes were shaken at 85ºC in a water bath for five
hours, followed by 2 mL H2O2 30%, one hour at 85ºC in water bath, then 15 mL of
NH4NO3 1M, then 10 minutes shaking at room temperature. Then the tubes were
centrifuged and treated at the same conditions described above.
RS: the residue from organic matter-bound fraction was air dried and ground with
an agate mortar. The ground residue (0.5 g) was digested in mixture of 6 mL of
nitric acid (69% purity) and 2 mL of chlorhydric acid (37% purity) in a microwave
reaction system (Multiwave 3000).
Toxicity characteristics leaching procedure (TCLP). TCLP test was carried out
to quantify the potential As leachability of soils according to the USEPA Method
1311 (1992). One gram of the air dried soil sample (untreated and nZVI treated)
was mixture with 20 mL of TCLP extraction solution (0.1 M glacial acetic and
0.0643 M NaOH, pH 4.94±0.03) in 50 mL centrifuge tubes. The mixture was
shaken for 18 h at room temperature (22±2 ºC) and then centrifuged at 13000 rpm
for 15 min. The supernatant was filtered with Whatman filter paper (nº 542),
acidified with HNO3 and brought to 25 mL with the TCLP solution. The TCLP
extracts were stored at 4ºC until analysis.
Statistical analysis. One-way ANOVA analysis at significant level of p<0.05 was
SHUIRUPHG ZLWK WKH 6366¶V VWDWLVWLFDO SDFNDJH IRU :LQGRZV UHOHDVH (SPSS Inc., IBM Company). The differences between means were compared using
Tukey test.
Results and discussion
Table 1 shows physicochemical parameters of the brownfield soil. This soil has
neutral pH, high percentage of organic matter and high concentration of As.
Arsenic concentration was in the range of 5.000-7.000 mg/kg, exceeding the
maximum levels allowed by Spanish legislation and other Europeans regulations
(BOPA 2014; BOCM 2006; DEFRA 2002).
547
TABLE 1. Physico-chemical properties of the soil studied.
pH
EC (dS/m)
CaCO3 (%)
N (%)
MO (%)
C/N
P (mg/kg)
Ca (mg/kg)
Mg (mg/kg)
Na (mg/kg)
K (mg/kg)
Cd (mg/kg)
Cr (mg/kg)
Zn (mg/kg)
Pb (mg/kg)
Ni (mg/kg)
Cu (mg/kg)
Mn (mg/kg)
As (mg/kg)
Fe (g/kg)
Fe oxides (g/kg FeOOH)
Mn oxides (g/kg MnO2)
Al oxides (g/kg Al(OH)3)
Porosity (%)
Field capacity (%)
Wilting point (%)
Sand (%)
Silt (%)
Clay (%)
7.14 ± 0.50
0.93 ± 0.17
2.8 ± 0.2
0.38 ± 0.02
6.9 ± 0.01
10.5 ± 0.0
219 ± 11
4,302 ± 969
403 ± 29
59 ± 15
150 ± 40
not detected
57.6 ± 5.2
117 ± 22
15.6 ± 4.2
36.6 ± 9.1
28.0 ± 6.3
522 ± 45
5,000-7,000
33.0-37.0
36.6 ± 0.55
0.64 ± 0.02
3.55 ± 0.20
52 ± 5
28.6 ± 0.34
22.2 ± 0.10
64.1±10.3
25.0±7.4
10.9±4.3
As and Fe fractionation. Tables 2 shows the As-distribution in soil fractions from
interaction assay. Quantitatively, the most important fraction was the RS fraction.
The application of nZVI decreased the As in EX and CB fractions, whereas an
increase of As in RS fraction was observed. The best immobilization results were
found at 10% of nZVI. These results evidence that nZVI can effectively immobilize
As in brownfield soil. Kumpiene et al. (2006) also observed a reduction of As
availability in soil after the application of iron grit.
To evaluate the stability of the As-immobilization, sequential extraction
procedures of soil samples collected after plant harvest were performed (Table 3).
In general, similar results of As-distribution to those found in the previous assay
were obtained, and the treatment at 10% showed the most effective immobilization.
We can highlight a slight increase of As in CB fraction for nZVI-treated and
untreated soils. Arsenic in CB fraction is potentially available for plants and other
soil organisms, so, long-term studies based on the stability of the As immobilization
in soil with nZVI are necessary.
TABLE 2. Mean concentration of As in soil fractions from batch experiment
(mg/kg).
548
Treatment
EX
CB
OX
OM
RS
0%
114 ± 5.29 a
377 ± 45.1 a
990 ± 36.4 a
136 ± 45.3 a
3608 ± 34.8 a
1%
62.2 ± 7.39 b
215 ± 15.2 b
1015 ± 49.5 a
76.8 ± 20.3 ab
3900 ± 121 b
10%
5.26 ± 0.57 c
46.2 ± 10.5 c
624 ± 58.2 b
28.2 ± 3.82 b
4372 ± 53.0 c
Values within a column followed by the same letter are not significantly different (p<0.05).
TABLE 3. Mean concentration of As in soil fractions from plant experiment
(mg/kg).
Treatment
EX
CB
OX
OM
RS
0%
83.2 ± 6.55 a
440 ± 22.5 a
926 ± 103 a
127 ± 39.6 a
4139 ± 99.2 a
1%
59.6 ± 4.94 b
410 ± 24.4 a
941 ± 162 a
137 ± 83.9 a
4469 ± 130 b
10%
12.1 ± 1.78 c
214 ± 58.1 b
624 ± 54.5 b
24.2 ± 10.1 b
5321 ± 80.3 c
Iron concentration was also determined in the different soil fractions to evaluate
the quantitative impact of this added element in the soil. Tables 4 and 5 show
mean Fe concentrations in each soil fraction in soil samples from the batch and pot
experiments, respectively. Iron was predominantly associated to RS fraction, and,
as expected, the highest total concentrations of Fe were detected in soils treated at
10%. According to the commercial specifications of the nZVI suspension previously
explained, the dose of 1% supposed an increase of Fe between 1,400 and 1,800
mg per kilogram of soil; and the dose of 10%, the increment was 14,000-18,000
mg per kg of soil. Iron was mainly in the less available fractions, in decreasing
order, RS>OX>OM>CB>EX. An and Zhao (2012), in a experiment with Fe-Mn
oxide nanoparticles for the immobilization of As, concluded that once delivered, the
nanoparticles remain nearly immobile and serve as a reactive sink for
immobilization of As in soil. The dose of 10% of nZVI showed a significant
increase of Fe in CB, OX, OM and RS fractions compared with soils treated at 1%
and untreated ones. However, the application of nZVI to soil at 1% and 10% did
not induce an important release of Fe to the EX fraction, the most mobile and
available soil fraction. In CB fraction, which also is available for plant and other soil
organisms, an increase of Fe was detected, which could suppose a stimulus for
plant development due to a better assimilation of Fe. After the plant experiment
(Table 5), a decrease of Fe bound to CB fraction was detected. This result is in
agreement with that found by Gil-Díaz et al. (2014), who observed a reduction of
Fe in CB fraction after three months of the application of nZVI on soil.
TABLE 4. Mean concentration of Fe in soil fractions from batch experiment
(mg/kg).
549
Treatment
EX
CB
OX
OM
RS
0%
5.56 ± 1.17 a
24.4 ± 0.34 a
2419 ± 296 a
1135 ± 186 a
33053 ± 960 a
1%
6.99 ± 1.39 a
43.4 ± 2.59 a
2301 ± 92.0 a
1136 ± 128 a
35567 ± 813 b
10%
6.92 ± 1.31 a
350 ± 44.8 b
7726 ± 361 b
1770 ± 38.1 b
43200 ± 350 c
TABLE 5. Mean concentration of Fe in soil fractions from plant experiment
(mg/kg).
Treatment
EX
CB
OX
OM
RS
0%
2.50 ± 0.35 a
5.08 ± 1.54 a
2564 ± 68.7 a
991 ± 88.7 a
27970 ± 25.3 a
1%
1.79 ± 0.39 a
4.70 ± 0.18 a
3103 ± 119 b
1006 ± 177 a
32458 ± 226 b
10%
3.82 ± 0.42 b
15.8 ± 2.62 b
7089 ± 314 c
1355 ± 258 a
44305 ± 147 c
As and Fe leachability. Effects of ZVI nanoparticles on immobilization of As in
brownfield soil was evaluated comparing As leachability in untreated and nZVItreated soils (Figure 1). Iron leachability was also measured to evaluate the
potential leaching of Fe after soil was treated with these nanoparticles. The TCLP
test was performed for soil samples from batch and plant experiments (Figure 1).
The application of nZVI at 1% and 10% to this As-polluted soil significantly
(p<0.05) reduced the amount of leachable As from 13.8 mg/L to 4.2 (69%
reduction) and 0.54 mg/L (96% reduction), respectively, after 72 hours of contact
time (Figure 1A). In soil samples collected after plant harvest, nZVI-treated soils
showed higher As leachability than in the previous assay. Soil treated at 10%
showed the lowest concentration of As, and no significant differences were
detected between untreated and 1%-treated soils. Immobilization at 10%-nZVI was
more effective than at 1% (Figure 1B). These results are in agreement with those
found by the sequential extraction procedures. In addition, Gil-Díaz et al. (2014)
also found a increase of the As leachability after three months of the application of
nZVI. Thus, further studies about the stability of As immobilization with iron
nanoparticles are necessary.
Regarding the Fe leachability, no substantial increase of Fe was observed
between untreated and treated soils, being the Fe mean concentration in the TCLP
extracts below 2 mg/L for soil samples from batch experiment. Statistically,
untreated soils and soils treated at 1% of nZVI leached similar content of Fe (below
0.5 mg/L), while the dose of 10% showed concentrations significantly higher (1.3
mg/L) (Figure 1A). In soil samples collected after pot experiment, the Fe
leachability was lower than in those from the batch experiment; soils from pot
experiment showed similar results, below 0.2 mg/L (Figure 1B). Thus, we can
550
conclude that the use of nZVI at the experimental conditions did not induce a
release of Fe over time.
a
B
16
1,8
14
1,6
1,4
12
As (mg/l)
2,0
1,2
10
1,0
8
0,8
A
b
6
4
0,6
0,4
A
2
Fe (mg/l)
18
A
0,2
c
0
0,0
0%
1%
10%
0%
As
1%
10%
Fe
16
2,0
a
1,8
a
1,6
14
1,4
As (mg/l)
12
1,2
10
1,0
8
0,8
b
6
Fe (mg/l)
18
B
0,6
4
A
2
A
A
0
0,4
0,2
0,0
0%
1%
10%
0%
As
1%
10%
Fe
FIGURE 1. As and Fe leachability (mg/L) from batch experiments (A) and
plant experiment (B). Bars with the same letter are not significantly different
(p<0.05). Lowercase for As; uppercase for Fe.
Plant experiment. A pot experiment in controlled conditions was performed to
determine the influence of this nanoremediation technology on the development of
barley plants; the accumulation of As and Fe in plants was also determined. In a
previous experiment performed with this soil, Gil-Díaz et al. (2014) observed that
the application of nZVI reduced the soil phytotoxicity for barley and vetch
germination. The stability of the As immobilization with nZVI was also evaluated.
Figure 2 and 3 illustrate the plant height monitoring, and the dried weight data of
the barley plants, respectively. Barley plants showed differences in their growth
551
depending on the treatment. Plants from soils treated at 10% of nZVI reached a
higher height and biomass than plants from 1%-treated and untreated soils. Plants
from unpolluted soils and from 10%-treated soils presented similar height values
although the biomass was higher in plants from unpolluted soils than in those from
10%-treated soils. Comparing with the positive control, the reached height by
barley plants grown in soils treated at 10% was around 94%, being by 31% and
19% for plants from 1% and 0% treatments, respectively. These results can be
explained by the reduction of the soil phytotoxicity since As availability was
reduced after the treatment with nZVI at 10%. The application of ZVI nanoparticles
to the As-polluted soil improved the development of barley plant.
35
30
height (cm)
25
20
0%
15
1%
10
10%
C+
5
0
4
7
11
14
18
21
25
28
32
Days FIGURE 2. Growth rate of the barley plants.
0,20
c
0,18
0,16
Dry weight (g)
0,14
b
0,12
0,10
0,08
0,06
0,04
a
a
0,02
0,00
0%
1%
10%
C+
FIGURE 3. Biomass of the barley plants. Bars with the same letter are not
significantly different (p<0.05).
552
As and Fe in barley plants. Arsenic concentration was measured in plant shoots
and roots (Figure 4). In all cases, the roots showed higher As uptake than shoots; it
is the so-called root barrier (Adriano 2001; Basta et al., 2005; Madejón and Lepp,
2007), which suggests the existence of a defense mechanism in the plant which
limits the transfer of heavy metal(loid)s accumulated in roots to the food chain
(Soriano-Disla et al., 2014). Plants from untreated soils concentrated the highest
amount of As. It can be explained by the higher availability of As in untreated soils
than in nZVI-treated soils. Plants grown in soils treated at 10% showed the lowest
accumulation of As, especially in shoot; a reduction of 97% of As concentration in
root was observed.
As in the case of Arsenic, Fe was mainly accumulated in roots (Figure 4). The
use of nZVI did not induce a higher accumulation of Fe, in fact, the highest
concentration of Fe was found in roots from untreated soils. Similar Fe
concentrations were detected in the roots from soils treated at 1% and 10% of
nZVI. All treatments showed similar amount of Fe in shoot. These results are in
agreement with those found by the sequential extraction procedure and the TCLP
which concluded that the nZVI application on soil did not induce a significant
increase of available Fe.
553
1400
5000
A
a
4500
As (mg/kg)
1200
4000
1000
3500
800
3000
AB
600
2500
2000
B
b
As (mg/kg)
A
1500
400
1000
200
500
c
0
0
0%
1%
10%
Shoot
0%
1%
10%
Root
250
Fe (mg/kg)
200
a
14000
A
a
12000
a
10000
150
8000
100
6000
B
50
B
Fe (mg/kg)
B
4000
2000
0
0
0%
1%
10%
Shoot
0%
1%
10%
Root
FIGURE 4. As (A) and Fe (B) concentration in shoot and root from barley
plants. Bars with the same letter are not significantly different (p<0.05).
Lowercase for shoot; uppercase for root.
Impact of the application of nZVI on soil properties. Different physico-chemical
soil parameters were analyzed in soil samples collected after plant harvest to
evaluate the impact of this nanoremediation technology on soils properties (Table
6). Similar values were obtained for pH, EC and C/N relation between nZVI-treated
and untreated soils. Organic matter and N were slightly higher in untreated soils
than nZVI-treated soils, probably due to a lower development of the plants. The
most important difference was found in available P, which, in decreasing order,
was 10%>1%>0%. This is due to two facts: firstly, P and As are in the same group
on the periodic table and have similar atomic configuration; phosphate and
arsenate are analogues of each other and therefore compete for the same sorption
554
sites in root apoplast and for the same uptake system in the root plasmalemma
(Asher and Reay, 1979; Rahman et al., 2008). The decrease of As availability in
10%-treated soils reduces the competition between phosphate and arsenate, and
more phosphate can be uptaken by plant roots. Secondly, ZVI nanoparticles could
immobilize phosphate in soil. In this sense, different studies have shown that these
nanoparticles have a high efficiency of removal of phosphate in water solutions
(Wu et al., 2013; Liu et al., 2013). These results showed that the application of
nZVI for As immobilization in brownfield soil at the assayed conditions did not
induce negative effects on soil physico-chemical properties.
TABLE 6. Soil properties after plant experiment.
Treatment
pH
CE (dS/m)
OM (%)
N (%)
C/N
P (mg/kg)
0%
7.27 ± 0.07 a
0.97 ± 0.09 a
8.01 ± 0.33 a
0.43 ± 0.02 a
10.9 ± 0.1 a
273 ± 35.5 a
1%
7.38 ± 0.04 ab
1.07 ± 0.08 a
6.76 ± 0.37 b
0.37 ± 0.02 b
10.8 ± 0.3 a
214 ± 3.56 b
10%
7.41 ± 0.07 b
1.15 ± 0.14 a
6.88 ± 0.36 b
0.37 ± 0.01 b
10.7 ± 0.5 a
165 ± 6.03 c
Conclusion
The present study shows the remediation of a highly As-polluted soil using nZVI.
The application of nZVI at doses of 1% and 10% significantly decreased the
availability of As in the soil, according to the results of the sequential extraction
procedure and the TCLP test. The As immobilization was more effective at 10%
than at 1% of nZVI commercial suspension. The lower availability of As in soils
treated at 10% led to lower uptake of As by barley plants which showed the highest
growth rate. Arsenic was mainly accumulated in root.
In relation to the impact of the nZVI application on Fe availability in soil, no
important differences were observed between untreated and treated at 1% soils,
while in soils treated at 10%, an increase of Fe in CB, OX, OM and RS fractions
were detected; Fe leachability initially increased moderately (< 2 mg/L), but a
release of Fe was not observed over the time studied. The use of nZVI did not
induce a higher uptake of Fe in barley plants. The As immobilization presented
some instability after the barley crop, since an increase of As, mainly, in CB
fraction was detected. Regarding the impact on the soil physico-chemical
properties, no negative effects were observed at the assayed conditions.
Thus, these results suggest that the ZVI nanoparticles are effective for As
immobilization in brownfield soils at the assayed conditions and could be combined
with remediation process where barley plants are involved. Long-term monitoring
studies are necessary to evaluate the stability of the As immobilization with nZVI.
Acknowledgments
555
This work has been supported by Project CTM 2010-20617-C02-02 and EIADES
PROGRAM S2009/AMB-1478.
References
An, B.; Zhao, D. (2012). Immobilization of As(III) in soil and groundwater using a new class of
polysaccharide stabilized Fe±Mn oxide nanoparticles. Journal of Hazardous Materials 211-212,
332-341.
BOCM, Boletín Oficial de la Comunidad Autónoma de Madrid 2006. Orden 2770/2006, August 11,
2006, Comunity of Madrid. Generic reference levels of heavy metals and other trace elements
in
contaminated
soils
from
the
Comunity
of
Madrid.
Available
at:
http://www.madrid.org/wleg/servlet/Servidor?opcion=VerHtml&nmnorma=4097&cdestado=P
BOPA, Boletín Oficial del Principado de Asturias, 91, April 21, 2014. Generic reference levels for
heavy metals in soils from Principality of Asturias, Spain. Available at:
https://sede.asturias.es/bopa/2014/04/21/2014-06617.pdf
De, D.; Mandal, S.M.; Bhattacharya, J.; Ram, S.; Roy, S.K. (2009). Iron oxide nanoparticle-assisted
arsenic removal from aqueous system. Journal of Environmental Science and Health Par A.
44: 155-162.
DEFRA, Department for Environment, Food & Rural Affairs, 2002. Soil Guideline Values for Arsenic
Contamination. Available from the R&D Dissemination Centre. WRc plc, Swindon, Wilts, UK.
El-Temsah, Y.S.; Joner, E.J.; (2012). Ecotoxicological effects on earthworms of fresh and aged
nano-sized zero-valent- iron (nZVI) in soil. Chemosphere 89: 76-82.
Environment Agency-Soil Guideline Values for inorganic arsenic in soil. Science Report SC050021/
arsenic SGV, 2009. Available at:
Gil-Díaz, M.; Alonso, J.; Rodríguez-Valdés, E.; Pinilla, P.; Lobo, M.C. (2014). Reducing the mobility
of arsenic in brownfield soil using stabilised zero-valent iron nanoparticles. Journal of
Environmental Science and Health, Part A, DOI: 10.1080/10934529.2014.928248.
Grieger K.D., Fjordbøge, A., Hartmann, N.B., Eriksson, E., Bjerg, P.L., Baun, A. (2010).
(QYLURQPHQWDO EHQH¿WV DQG ULVNV RI ]HUR-valent iron nanoparticles (nZVI) for in situ
remediation: Risk mitigation or trade-off?. Journal of Contaminant Hydrology 118: 165-183.
Gyuricza, V.; Fodor, F.; Szigeti, Z. (2010). Phytotoxic Effects of heavy metal contaminated soil
reveal limitations of extract-based ecotoxicological tests. Water, Air & Soil Pollution 210: 113122.
Hartley, W., Lepp, N.W., (2008). Remediation of arsenic contaminated soils by iron-oxide
application, evaluated in terms of plant productivity, arsenic and phytotoxic metal uptake.
Science of the Total Environment 390: 35-44.
http://www.epa.gov/waste/hazard/testmethods/sw846/online/index.htm#table
https://www.gov.uk/government/uploads/system/uploads/attachment_data/file/313869/scho0409bpv
y-e-e.pdf
Jackson, M.L.; Lim, C.H.; Zelazny, L.W. (1986). Methods of Soil Analysis. Part 1. A. Klute, Ed., Soil
Science Society of America, Madison, WI, 101-150.
-HJDGHHVDQ*0RQGDO./DOYDQL6%$UVHQDWH5HPHGLDWLRQ8VLQJ1DQRVL]HG0RGL¿HG
Zerovalent Iron Particles. Environmental Progress 24: 289-296.
Kanel, S.R.; Grenèche, J.M.; Choi, H. (2006). Arsenic(V) Removal from Groundwater Using Nano
Scale Zero-Valent Iron as a Colloidal Reactive Barrier Material. Environmental Science &
Technology 40: 2045-2050.
Karn, B., Kuiken, T., Otto, M. (2009). Nanotechnology and in Situ Remediation: A Review of the
Benefits and Potential Risks. Environmental Health Perspectives 117: 1823-1831.
Kim, J.Y.; Davis, A.P.; Kim, K. W. (2003). Stabilization of available arsenic in highly contaminated
mine tailings using iron. Environmental Science & Technology 37: 189-195.
Klimkova, S.; Cernik, M.; Lacinova, L.; Filip, J.; Jancik, D.; Zboril, R. (2011). Zero-valent iron
nanoparticles in treatment of acid mine water from in situ uranium leaching. Chemosphere 82:
1178-1184.
556
.RPiUHN 0 9DQČN $ (WWOHU 9 &KHPLFDO VWDELOL]DWLRQ RI PHWDOV DQG DUVHQLF LQ
contaminated soils using oxides. A review. Environmental Pollution 172: 9-22.
Kumpiene, J.; Lagerkvist, A.; Maurice, C. (2008). Stabilization of As, Cr, Cu, Pb and Zn in soil using
amendments- A review. Waste Management 28: 215-225.
Kumpiene, J.; Ore, S.; Renella, G.; Mench, M.; Lagerkvist, A.; Maurice, C. (2006). Assessment of
zerovalent iron for stabilization of chromium, copper, and arsenic in soil. Environmental
Pollution 144: 62-69.
Larios, R.; Fernández-Martínez, R.; LeHecho, I.; Rucandio, I. (2012). A methodological approach to
evaluate arsenic speciation and bioaccumulation in different plant species from two highly
polluted mining areas. Science of the Total Environment 414: 600-607.
Liu, H., Chen, T., Zou, X., Xie, Q., Qing, C., Chen, D., Frost, R.L. (2013). Removal of phosphorus
using NZVI derived from reducing natural goethite. Chemical Engineering Journal 234: 80-87.
Lobo, M.C, Pérez-Sanz A, Martínez-Iñigo, M.J., Plaza A. (2009). Influence of coupled
electrokinetic-phytoremediation on soil remediation In: Electrochemical Remediation
Technologies for Polluted Soils, Sediments and Groundwater Chapter 20. pp: 417-437 John
Wiley & Sons.
Ma, X.; Geisler-Lee, J.; Deng, Y.; Kolmakov, A. (2010). Interactions between engineered
nanoparticles (ENPs) and plants: Phytotoxicity, uptake and accumulation. Science of the Total
Environment 408: 3053-3061.
Ma, X.; Gurung, A.; Deng, Y. (2013). Phytotoxicity and uptake of nanoscale zero-valent iron (nZVI)
by two plant species. Science of the Total Environment 443: 844-849.
Madejón, P. and Lepp, N.W. (2007). Arsenic in soils and plants of Woodland regenerated on an
arsenic-contaminated substrate: A sustainable natural remediation? Science of the Total
Environment 379: 256-262.
MAPA (1994). Métodos Oficiales de Análisis, vol. III, Secretaría General Técnica Ministerio de
Agricultura, Pesca y Alimentación, Vol III pp 205-324. Spain.
Mench, M.; Schwitzguébel; J.P.; Schroeder; P.; Bert; V.; Gawronski; S.; Gupta; S. (2009).
Assessment of successful experiments and limitations of phytotechnologies: contaminant
uptake, detoxification, and sequestration, and consequences to food safety. Environmental
Science and Pollution 16: 876-900.
Miehlich G. (2001). Do contaminated soils have to be decontaminated? In Treatment of
Contaminated Soil. Fundamentals, Analysis, Applications (eds. R Stegmann, Grunner,
WCalamano, GMatz). Berlin: Springer-Verlag.
Monnier, G.; Stengel, P.; Fies, J.C. (1973). Une Méthode de Mesure de la Densité Apparente de
Petits Agglomérats Terreux: Application a l'analyse des Systèmes de Porosité du Sol. Ann.
Agron 24: 533-545.
Moreno-Jiménez, E.; Esteban, E.; Peñalosa, J.M. (2012). The fate of arsenic in soil-plant systems.
Reviews Environmental Contamination and Toxicology 215: 1-37.
0XHOOHU 1& %UDXQ - %UXQV - ýHUQtN 0 5LVVLQJ 3 5LFNHUE\ ' 1RZDFN % Application of nanoscale zero valent iron (nZVI) for groundwater remediation in Europe.
Environmental Science and Pollution 19: 550-558
Ruiz-Chancho, M.J.; López-Sánchez, J.F.; Rubio, R. (2007). Analytical speciation as a tool to
assess arsenic behaviour in soils polluted by mining. Analytical and Bioanalytical Chemistry
387: 627-635.
Shaibur, M.R., Adjadeh, T.A., Kawais, S. (2013). Effect of phosphorus on the concentrations of
arsenic, iron and some other elements in barley grown hydroponically. Journal of Soil Science
and Plant Nutrition 13: 87-98.
Shipley, H.J.; Engates, K.E.; Guettner, A.M. (2011). Study of iron oxide nanoparticles in soil for
remediation of arsenic. Journal of Nanoparticle Research 13: 2387-2397.
Sierra, C.; Menéndez-Aguado, J. M.; Afif, E.; Carrero, M.; Gallego, J.R. (2011). Feasibility study on
the use of soil washing to remediate the As-Hg contamination at an ancient mining and
metallurgy area. Journal of Hazardous Materials 196: 93-100.
Smith, E.; Naidu, R.; Alston, A.M. (1998). Arsenic in the soil environment: A review. Advances in
Agronomy 64: 149-195.
557
Soriano-Disla, J.M., Gómez, I., Navarro-Pedreño, J., Jordán, M.M. (2014). The transfer of heavy
metals to barley plants from soils amended with sewage sludge with different heavy metal
burdens. Journal of Soils and Sediments 14: 687-696.
Tessier, A.; Campbell, P.G.C.; Bisson, M. (1979). Sequential extraction procedure for the speciation
of particulate trace metals. Analytical Chemistry 51: 844-850.
US-EPA (1992). Environmental Protection Agency, 1992. Method 1311: Toxicity Characteristic
Leaching Procedure. Available at:
Wang, Y.; Fang, Z.; Kang, Y.; Tsang, E. (2014). Immobilization and phytotoxicity of chromium in
contaminated soil remediated by CMC-stabilized nZVI. Journal of Hazardous Materials 275:
230-237.
Watanabe, F.S. y Olsen, S.R. (1965). Test of an Ascorbic Acid method for determining phosphorus
in water and NaHCO3 extracts from soil. Soil Science Society of America 29: 677-678.
Wu, D., Shen, Y., Ding, A., Qui, M., Yang, Q., Zheng, S. (2013). Phosphate removal from aqueous
solutions by nanoscale zero-valent iron. Environmental Technology 34: 2663-2669.
Zhang, M.Y.; Wang, Y.; Zhao, D.Y.; Pan, G. (2010). Immobilization of arsenic in soils by stabilized
nanoscale zero-valent iron, iron sulfide (FeS), and magnetite (Fe3O4) particles. Chinese
Science Bulletin 55, 365-372.
Zhao, F., McGrath, S., Crosland, A. (1994). Comparison of three wet digestion methods for the
determination of plant sulphur by Inductively-Coupled Plasma-Atomic Emission-Spectroscopy
(ICP-AES). Communications in Soil Science and Plant Analysis 25: 407-418.
558
LIST OF COAUTHORS:
S. Diez-Pascual, M. Sc. presenting author.
,0,'5$ 'HSW RI $JURHQYLURQPHQWDO 5HVHDUFK )LQFD ³(O (QFtQ´ $OFDOi GH
Henares, 28800 (Madrid) Spain
Tel.: 00 34 918879473
E-mail: [email protected]
Dr. M.C. Lobo, Researcher.
Tel.: 00 34 918879472
Dr. A. González, Researcher.
Tel.: 00 34 918879453
J. Alonso, M Sc.
Tel.: 00 34 918879467
E. Rodríguez-Valdés, Researcher.
Universidad de Oviedo, Dept. of Technology, Biotechnology and Environmental
Geochemistry, Mieres 33600 (Asturias), Spain
Tlf: 00 34 985103000 Ext. 5856
Dr. M. Gil-Díaz, Researcher. Author to whom all correspondence should be
addresed
Tel.: 00 34 918879475
559
CHAPTER 6.2. MICROORGANISMOS FIJADORES DE NITROGENO DE VIDA
LIBRE CON CAPACIDAD DE DEGRADACIÓN DE CLORPIRIFOS
David Antonio Moreno-Medina (1); Julián Delgadillo-Martínez*1(1);
David Espinosa-Victoria (1); Rosa María López-Romero (1);
Ma. Remedios Mendoza-López (2); Oscar García-Barradas (2).
(1) CP. Campus Montecillo, Edo. De México, México;
(2) SARA-UV, Veracruz, México.
ABSTRACT
Chlorpyrifos [O, O-diethyl O-(3, 5, 6-trichloro-2-pyridinil) -fosforotioate] is a nonsystemic organophosphate pesticide which affects the nervous system target
insects by inhibiting the acetylcholinesterase activity by phosphorylation. The
accumulation in soil and water as a result of indiscriminate use cause potential
problems to the environment and public health. The main objective of this work
was the isolation and characterization of microorganisms capable of degrading
these compounds to have tools to restore contaminated environments and treat the
waste for disposal. Microorganisms were isolated from chrysanthemum crop soils
with a long historical of application of chlorpyrifos. First nine strains capable of
growth at concentrations up to 350 mg L-1 of pesticide in a culture medium rich in
nutrients were isolated. Subsequently underwent grow in minimal media culture,
where only four strains grown in these conditions and in the presence of the
pesticide as sole carbon source were selected. These four strains also had the
ability to fix nitrogen, the results with a significance level of p <0.05 showed
acetylene reduction rates of 15.5, 23.9, 28.3 and 30.1 nmol h -1 mL-1 for the strains
CP30, CP37, CP46 and CP47, respectively. It is noteworthy that for the CP30
strain growth was higher in the presence of the pesticide (50 mg L -1) compared to
growth in medium without pesticide on measurements during 96 hours. Maximum
growth was presented at 72 (9.2x109 CFU mL-1) and 48 (6.7x108 CFU mL-1) hours,
for the strain grown with and without pesticide, respectively. Additionally, and
considering that the microorganisms capable of degrading these compounds do
with its enzymes, a phylogenetic analysis of genes coding for enzymes that
degrade organophosphorus compounds, and which may reside within the isolated
strains was performed. The evaluating Chlorpyrifos removal is ongoing.
Palabras clave: biorremediación, plaguicida, degradación
*1 Autor de correspondencia: [email protected]
560
Introducción
La contaminación ambiental representa un problema complejo de resolver, debido
a los múltiples aspectos que involucra. Es provocada por una gran diversidad de
contaminantes que ingresan al ambiente, dentro de los cuales destacan los
plaguicidas, debido a la variedad de usos que han recibido en la agricultura, la
salud, la industria, la ganadería, la silvicultura, entre otros. El impacto de los
plaguicidas se puede abordar desde dos puntos de vista: el ambiental, que se da
cuando estos se introducen en las cadenas tróficas y el de salud pública, originado
por la exposición directa o indirecta con los compuestos químicos.
Los plaguicidas son sustancias o mezclas de sustancias que se destinan a
prevenir, destruir o controlar cualquier plaga. Se pueden clasificar con base en
diferentes criterios, el grupo de los organofosforados (OF) es el más importante
debido a la cantidad de uso (34%) y a su nivel de toxicidad. Los OF son derivados
del ácido fosfórico, tienen efectos a corto plazo, son altamente tóxicos y
potentemente inhibidores de la acetilcolinesterasa (CICOPLAFEST, 2004). Para
mitigar el problema de contaminación por OF existen tratamientos en los que
intervienen procesos físicos, químicos y biológicos orientados a detoxificar o
degradar al plaguicida.
El Clorpirifos es un plaguicida OF de amplio espectro para el control de
plagas, de baja solubilidad en agua (0.002 gL-1 a 25 °C) y un coeficiente de
partición carbono orgánico-agua de log Koc 3.78 lo cual le confiere una fuerte
absorción por las partículas del suelo (Tomlin, 1994). Se hidroliza fácilmente en su
metabolito primario, 3,5,6-tricolopyridinol, el cual es moderadamente móvil y
persistente en el suelo. La vida media del Clorpirifos depende del tipo de suelo y
las condiciones del ambiente, sin embargo, para suelos arcillosos el rango es
entre 120 y 450 días. Es por ello que es necesario contar con herramientas
biotecnológicas para disminuir su presencia en el ambiente, y para ello están los
diferentes tratamientos (Fang et al., 2009).
El tratamiento biológico, utiliza sistemas vivos o sus partes (células completas
o enzimas aisladas), los cuales catalizan reacciones químicas, sobre estos
xenobióticos hasta llevarlos a compuestos más sencillos y menos tóxicos,
inclusive se puede llegar a la mineralización completa del contaminante. El
aislamiento y caracterización de microorganismos capaces de degradar este tipo
de compuestos posibilita contar con herramientas para restaurar ambientes
contaminados (biorremediación) y tratar los residuos para su disposición final
(Imfeld y Vuilleumier, 2012). El objetivo del presente trabajo fue caracterizar e
identificar cepas bacterianas capaces de crecer en presencia de Clorpirifos.
Materiales y métodos
Aislamiento y caracterización. El muestreo se realizó en dos diferentes suelos
en la zona de Texcoco, estado de México. El primer suelo de invernadero tiene
con un historial amplio de aplicación de plaguicidas, principalmente
organofosforados, para el control de plagas y enfermedades relacionadas con la
floricultura; aplicación de bromuro de metilo diferente con el otro suelo. Posterior al
muestreo, se realizaron pruebas de crecimiento de los microorganismos en
561
medios sólidos, Agar nutritivo, Medio modificado de Rennie (1981) y Medio de
sales minerales (MSM, Moreno-Medina, 2014); con y sin extracto de levadura
como fuente alterna de carbono y se incrementó progresivamente la concentración
del
plaguicida
Clorpirifos
[O,O-diethyl
O-(e,5,6-trichloro-2-pyridinyl)-1
phosphorothioate] grado técnico hasta 350 mg L . Para incrementar la superficie
de contacto se procedió al crecimiento de cepas seleccionadas, pero en medio
líquido, conteniendo 100 mg L-1 de Clorpirifos y 0.5 g L-1 de extracto de levadura.
La incubación fue por 48 h., 28 °C, agitación orbital, inóculo de 8.0x108 UFC mL-1
(de acuerdo a escala de McFarland). Una vez aisladas e identificadas se
determinó su crecimiento durante 96 horas en MSM con Clorpirifos (50 mg L -1) e
inoculando 9.0x109 UFC mL-1 y tomando muestras para cuantificar la degradación
del plaguicida.
Identificación molecular y análisis filogenético. Una vez aisladas la cepas, se
realizó la extracción de su ADN (Kid de extracción ZR Fungal/Bacterial DNA
MiniPrepTM de Zymo Research), amplificación y purificación del gen 16S RNAr
para así secuenciar este fragmento de su genoma y compararlo filogenéticamente.
El programa para la amplificación por PCR fue el siguiente:
Paso
Primera desnaturalización
Desnaturalización
Hibridación
Extensión
Post-ciclo (extensión)
Temperatura (°C)
95
95
50
72
72
Tiempo
1 min
45 seg
60 seg
30 seg
5 min
Ciclos
1
30
1
Además, se hizo un análisis filogenético de los posibles genes involucrados en
la degradación del Clorpirifos por las cepas aisladas e identificadas. El análisis
filogenético se realizó con el software MEGA 6.0 (Tamura et al., 2013).
Fijación de nitrógeno. Se realizó la prueba de reducción de acetileno por acción
de la enzima nitrogenasa, para verificar la capacidad de fijar nitrógeno de cada
cepa, además de crecer en presencia de Clorpirifos. La determinación del etileno y
acetileno (inyección de 0.5 mL) se realizó en un Cromatógrafo de gases Hewlett
Packard 5890 series II, columna poraplot Q (25 m, 0.32 mm), con un detector FID;
las temperaturas del detector, el inyector y del horno fueron 150, 100 y 50 °C,
respectivamente.
Los análisis estadísticos se realizaron en SAS 9.0, en base a un diseño
completamente al azar con tres repeticiones y comparación de medias por Tukey
con un nivel de significancia de 0.05. Para el caso de los valores de fijación de
nitrógeno se realizó una comparación de medias por LSD de Fisher.
Resultados y discusión
Aislamiento de cepas. En general los plaguicidas afectan el equilibrio y la
biodiversidad de los suelos. Algunos organofosforados estimulan el crecimiento de
562
bacterias endémicas del suelo, pero otros disminuyen la concentración de
bacterias nitrificantes afectando directamente la fertilidad del suelo (Kalia y Gosal,
2011; Lo, 2010). Por su parte, el Clorpirifos (CLP) es un plaguicida de amplio
espectro usado en la agricultura, y de acuerdo con la RAP-AL (2012) hay
suficiente evidencia de que CLP es persistente, tiene el potencial para transporte
ambiental de largo alcance y causa efectos adversos tanto al ambiente como a la
salud humana.
Debido a ello la importancia de contar con la exploración de bacterias
degradadoras de este compuesto. Primeramente con los medios sólidos, fuente
alterna de carbono y CLP (10 mg L-1) se seleccionaron 12 cepas del suelo 1, y 12
del suelo 2. Posteriormente con 350 mg L-1 de CPL, solo se seleccionaron dos
cepas del suelo 1 (CP5 y CP6), y siete del suelo 2 por su crecimiento en estas
condiciones (Figuras 1, 2 y 3).
FIGURA 1. Crecimiento de cepas
bacterianas en caldo nutritivo y
presencia de Clorpirifos a 200 mg
L-1.
FIGURA 2. Crecimiento de cepas
bacterianas en medio de Rennie
modificado
y
presencia
de
Clorpirifos a 100 mg L-1.
1.2E+10
UFC mL-1
1E+10
8E+09
6E+09
4E+09
2E+09
d
d
abc cd bcd cd
ab
ab
a
5
6
30
38
46
47
0
33
34 37
Cepa
FIGURA 3. Crecimiento de cepas bacterianas en medio de sales minerales
(MSM) y presencia de Clorpirifos a 100 mg L-1.
En la figura 1 se muestra el crecimiento de las cepas en un medio rico en
nutrientes, se observa que CP30, CP34, CP37, CP46 y CP47 muestran los
mayores valores a pesar de están sometidas a estrés por presencia de CLP. Por
otro lado, el medio de Rennie (1981) es un medio específico para organismos
fijadores de nitrógeno, en este medio, las cepas CP37, CP46 y CP47 crecieron en
563
más que las otras seis cepas (figura 2). Esto es posible ya que estas cepas fijan
nitrógeno y no sean simbióticas. También, en la Figura 3 se muestra el crecimiento
que tuvieron las cepas en un medio mínimo (MSM), y agregando CLP y una fuente
alterna de carbono como co-sustrato. En este medio las CP30, CP38, CP36 y
CP47 mostraron el mayor crecimiento.
De esta manera, se seleccionaron las cepas CP30, CP37, CP46 y CP47 por
presentar características deseables para utilizar al CLP como fuente de carbono y
fijar nitrógeno. Comparado con Li et al. (2008), ellos reportaron siete cepas
bacterianas degradadoras de hasta 100 mg L-1 de CLP aisladas en MSM, el
género Pseudomonas fue el que mayor estuvo presente, cabe resaltar que el
muestreo lo realizaron de suelos contaminados con historial de aplicación de CLP
durante dos años. También, Ning et al. (2010) realizaron aislamiento de seis
cepas degradadoras de diclorvos en un ambiente modificado con 400 mg L -1 de
plaguicida, reportaron por primera vez que los géneros Acidovorax
y
Chryseobacterium pueden degradar organofosforados.
Sin embargo, después de la selección de cepas se confirmó el crecimiento de
estas cuatro cepas incrementando gradualmente la fuente de nitrógeno (0, 0.04,
0.06 y 0.08 % de KNO3) y la fuente alterna de carbono (0, 0.2, 0.4, 0.6, 0.8, 1.0,
1.2 y 1.4 % de C6H12O6).
A (630 nm)
0.8
0.6
0.4
0.2
0
0%
0.02% 0.04% 0.06%
0%
MCC
0.02% 0.04% 0.06%
0%
MSM
0.02% 0.04% 0.06%
MCC+EL
0.0%
0.2%
0.4%
0.6%
0.8%
1.0%
1.2%
1.4%
KNO3
A (630 nm)
FIGURA 4. Crecimiento de cepa CP30 con incremento en fuente de nitrógeno y
carbono.
0.8
0.0%
0.6
0.4%
0.4
0.6%
0.2
0.8%
1.0%
0
0%
0.04% 0.06% 0.08%
MCC
0%
0.04% 0.06% 0.08%
MSM
KNO3
0%
0.04% 0.06% 0.08%
MCC+EL
1.2%
1.4%
1.6%
carbono.
564
A (630 nm)
0.0%
0.8
0.4%
0.6
0.6%
0.4
0.8%
0.2
1.0%
0
0%
0.04% 0.06% 0.08%
0%
MCC
0.04% 0.06% 0.08%
0%
MSM
0.04% 0.06% 0.08%
MCC+EL
1.2%
1.4%
1.6%
KNO3
A (630 nm)
carbono.
0.8
0.0%
0.6
0.4%
0.6%
0.4
0.8%
0.2
1.0%
0
0%
0.04% 0.06% 0.08%
MCC
0%
0.04% 0.06% 0.08%
MSM
KNO3
0%
0.04% 0.06% 0.08%
MCC+EL
1.2%
1.4%
1.6%
carbono.
En la figura 4 se observa el crecimiento de la cepa CP30 resaltando que en
MSM no afecta la fuente de nitrógeno alterna ya que contiene per se, no obstante,
en el MCC es indispensable ya que según Rennie (1981) la disponibilidad de
nitrógeno determina un factor de selección para ciertos microorganismos. Así, se
observa que el extracto de levadura (MCC+EL) posee compuestos nitrogenados
que hacen que no afecte severamente el crecimiento. Por su parte, en incremento
en la fuente de carbono no es significativo, a excepción del crecimiento en MCC,
ya que a partir de 0.04% favorece el crecimiento.
La cepa CP37 (figura 5) se comportó diferente a la CP30, se observa una
preferencia para crecer en MCC, posiblemente sea un género específico para fijar
nitrógeno (ya que crece no importando si hay o no fuente de nitrógeno en el
medio) en la que la fuente de carbono no es importante. A su vez, las cepas CP46
y CP47 (figuras 6 y 7) presentan un comportamiento parecido en su
aprovechamiento de carbono y nitrógeno en estos medios. Cabe resaltar que el
incremento en la fuente de nitrógeno para la cepa CP47 aparentemente favorece
su crecimiento. De esta manera, se puede inferir que las cepas fijan nitrógeno y
además son capaces de crecer en presencia del plaguicida.
565
FIGURA 8. Cinética de crecimiento
cepa CP30 en MSM y 50 mg L-1 de
Clorpirifos.
FIGURA 9. Cinética de crecimiento
cepa CP46 en MSM y 50 mg L-1 de
Clorpirifos.
566
Environmental Biotechnology and Engineering-2014
Las figuras 8 y 9 ilustran ejemplos de la cinética de crecimiento. Se observa que
la cepa CP30, con un comportamiento parecido a la CP47, mostraron el mayor
crecimiento que las otras cepas (datos no mostrados). El crecimiento en presencia
del CLP fue notorio y significativo, posiblemente debido a que el plaguicida fue
aprovechado como fuente de carbono o co-metabolizado. Para esta cepa, el
crecimiento máximo se presentó a las 72 (9.2x109 UFC mL-1) y 48 (6.7x108 UFC
mL-1) horas, para la cepa cultivada con y sin plaguicida, respectivamente. Ya se
han se han realizado diversos ensayos considerando al plaguicidas como fuente
principal de carbono y como co-metabolismo.
El co-metabolismo se refiere a la trasformación biológica de un co-sustrato
(compuesto que se co-metaboliza por la adición de un sustrato al medio) mientras
que el sustrato (compuesto adicionado al medio que permite la inducción de
enzimas requeridas para la degradación) sirve como energía primaria como fuente
de carbono (García-Rivero y Peralta-Pérez, 2008). En los ensayos realizados en
este trabajo se utilizó extracto de levadura como un co-sustrato, aunque hay otros
estudios que mencionan otros co-sustratos como la glucosa, el succinato, el
salicilato, la peptona, inclusive algunos hidrocarburos policíclicos aromáticos de
bajo peso molecular (Zhong et al., 2010).
Por ejemplo, especies del genero Bacillus y Flavobacterium utilizan este
compuesto como co-metabolismo, pero Bacillus cereus, Bacillus licheniformis,
Klebsiella sp., Paracoccus sp., Pseudomonas nitroreducens, Pseudomonas
aeruginosa, Pseudomonas stutzeri, Serratia sp. y Sphingomonas sp. lo hacen
como parte de su catabolismo. Incluso, algunas otras especies como Alcaligenes
faecalis y Enterobacter sp. realizan catabolismo de P además del C (Chishti et al.,
2013). La diferencia es la concentración del plaguicida y la matriz en donde se
desarrolle el ensayo de aislamiento y degradación. Así también, se ha encontrado
degradación de hasta 98% de CLP (30 mg L -1) en 10 días en suelo (Li et al.,
2007).
Identificación molecular y análisis filogenético. Los datos preliminares de la
secuenciación del gen 16S RNAr muestran que las cepas CP30, CP37 y CP46
pertenecen al género Pseudomonas mientras que CP47 al género Serratia.
Algunas especies de estos géneros han sido evidenciados con capacidad de
degradar este compuesto. Por ejemplo, según Lakshmi et al. (2008) Pseudomonas
aeruginosa degrada hasta el 87 % de CLP en su metabolito principal 3,5,6trichloro-2-pyridinol (TCP) tan solo en 20 dias en medio líquido. Por su parte,
Yadav et al. (2014) reportaron que Pseudomonas sp. puede degradar hasta el
91% de CLP (300 mg L-1 d-1) en un biorreactor, no obstante, los metabolitos
afectan la eficiencia del equipo.
También, Awad et al. (2011) obtuvieron una cepa de Pseudomonas stutzeri (BCP5) que tolera al CLP (300 mg L-1) hasta por siete dias. Entonces, este género
es de relevancia para la transformación de este plaguicida degradandolo incluso
totalmente como lo alcanzado con la cepa Pseudomonas nitroreducens (PS-2)
inoculada en rizosfera (Korade y Fulekar, 2009). Otro ejemplo es lo encontrado
por Sasikala et al. (2012), en donde sus aislamientos de Pseudomonas putida (NII
1117), Klebsiella sp., (NII 1118), Pseudomonas stutzeri (NII 1119) y Pseudomonas
aeruginosa (NII 1120) resultaron eficientes para degradar hasta 500 mg L-1 de
567
CLP, cabe resaltar que los análisis de metabolitos demostraron la presencia de los
metabolitos Clorpirifosoxon y Dietylfosforotioato a diferencia del TCP reportado
mayormente.
Por otro lado, Serratia sp. de igual manera se ha considerado para degradar
CLP. Cycon et al. (2013) hacen mencion de un 58.9% de degradación durante 14
dias en MSM (y CLP como única fuente de carbono), en suelo arenos un 45.3%
(en 42 dias), en suelo franco arenoso 61.4% y 68.9% en suelo limoso. Indicando
que la atenuacion de CLP con Serratia marcescens es lenta y esta dada en
funcion del tipo de suelo y otros factores, pero sí alcanza efectividad considerable
si se combina con otras bacterias u hongos nativos del suelo. Ejemplo de ello es lo
reportado por Xu et al. (2007), quienes consiguieron una mineralizacion completa
de CLP (50 mg L-1) en co-cultivo con Serratia sp. y Trichosporon sp.en solo 18
horas.
En la figura 10 se observa que los genes con mayor énfasis en las
investigaciones recientes son el opd presente en especies de géneros como
Pseudomonas, Flavobacterium, Sphingomonas
y el gen mph en
Stenotrophomonas. Considerando los parrafos anteriores y el analisis filogenetico
es importante mencionar que el género Pseudomonas presenta en su genoma
genes como el opd, mpd y pepA por lo cual puede considerarse muy versátil para
su uso en biorremediación. Los géneros Burkholderia y Flavobacterium toman
importancia por la presencia de genes de degradación de OF al igual que algunas
cepas de E. coli.
Para el caso del Clorpirifos, Yang et al. (2006) aislaron una cepa de
Stenotrophomonas con capacidad de degradar CLP e incrementaron su capacidad
de metabolizarlo al insertar el gen mpd de Plesiomonas sp. (M6). Li et al. (2007)
tambien aislaron una Sphingomonas sp. (Dsp-2) e indicaron que el gen mpd es el
responsable de la alta degradación de CLP (100 mg L-1). Recientemente, Barman
et al. (2014) reportan que el gen ophB, que codifica para una proteina involucrada
en la hidrólisis de CLP, se encuentra en una bacteria endofitica Pseudomonas sp.
(BF1-3), tambien clonaron y expresaron este gen en E. coli '+Į TXH GHJUDGy
hasta el 97% de CLP (100 mg L-1) en nueve dias.
568
Brevundimonas diminuta strain MG plasmid pCMS1 (opd)
Flavobacterium sp. MTCC 2495 organophosphorus hydrolase (opd) gene
Flavobacterium sp. ATCC 27551 gene (opd) gene
Sphingomonas fuliginis ATCC 27551 plasmid pPDL2 tnp gi408386312 (opd)
Flavobacterium sp. parathion hydrolase gene (opd)
Pseudomonas putida organophosphate hydrolase (opd) gene
Flavobacterium balustinum plasmid pBC9 (opd) gene for parathion hydrolase
Sphingomonas sp. JK1 organophosphorus hydrolase (opd) gene
PSEPTE Plasmid pCMS1 (from P.diminuta) phosphodiesterase (opd) gene
Flavobacterium sp. (opd) protein gene
FVBOPDA Flavobacterium sp. (opd) protein gene
Agrobacterium tumefaciens phosphotriesterase (opdA) gene
Pseudomonas putida organophosphate hydrolase (opd) gene
Synthetic construct organophosphorus hydrolase (OPD) gene
BCU44852 Burkholderia caryophylli phosphonate monoester hydrolase (pehA) gene
Pseudomonas putida phosphotriesterase gene (opd)
Cupriavidus taiwanensis strain X1 organophosphorus hydrolase (opdB) gene
Pseudomonas sp. Dsp-1 (mpd) gene
Stenotrophomonas sp. Dsp-4 (mpd) gene
Agrobacterium tumefaciens methyl parathion hydrolase (mpd) gene
Ochrobactrum sp. Yw28 methyl parathion hydrolase (mpd) gene
Arthrobacter sp. 68b putative phthalate ester hydrolase (pehA) gene
ASU29240 Alteromonas sp. organophosphorus acid anhydrolase-2 (opaA) gene
Enterobacter sp. CCF01Et-UAEM OpdE (opdE) gene
Enterobacter sp. Ceibso-001 Mpc (mpc) gene (opdE)
Burkholderia sp. FDS-1 organophosphorus insecticide hydrolase (opdB) gene
Burkholderia sp. FDS-1 (opdB mpd1)
Cupriavidus sp. DT-1 Mpd (mpd) gene
Stenotrophomonas acidaminiphila organphorous hydrolase (mpd) gene
Pseudomonas sp. WBC-3 methyl parathion hydrolase (mpd)
Pseudomonas sp. YF-5 methyl parathion hydrolase (mpd) gene
Pseudomonas stutzeri strain ZK-5 methyl parathion hydrolase (mpd) gene
Ochrobactrum sp. mp-4 organophosphate pesticide hydrolase (mpd) gene
Sphingopyxis sp. D-8 (Mpd) gene
Plesiomonas sp. M6 methyl parathion hydrolase (mpd) gene
Pseudaminobacter sp. mp-1 organophosphate pesticide hydrolase (mpd) gene
Pseudomonas stutzeri strain HS-D36 Mpd (mpd) gene
Achromobacter sp. mp-2 organophosphate pesticide hydrolase (mpd) gene
Brucella sp. mp-7 organophosphate pesticide hydrolase (mpd) gene
Sphingomonas sp. Dsp-2 (mpd) gene
Pseudomonas putida phosphotriesterase gene (hocA)
Stenotrophomonas maltophilia strain MHF ENV 20 organophosphorous hydrolase (mpd) gene
Stenotrophomonas sp. OP-1 methyl parathion hydrolase (mpd) gene
Stenotrophomonas sp. YC-1 methyl parathion hydrolase (mpd) gene
FIGURA 10. Análisis de taxa inferido mediante Máxima Parsimonia (MP). El
porcentaje de árboles replica en los cuales se asociaron los taxa en el test
bootstrap (500 réplicas) se muestra en las ramas. El árbol de MP fue
obtenido usando el algoritmo Tree-Bisection-Regrafting (TBR). El análisis
involucro 49 secuencias de nucleótidos, en total 8876 posiciones.
569
Fijación biológica de Nitrógeno. Las bacterias diazotróficas asimbioticas son
aquellas que pueden fijar nitrogeno atmosferico sin necesidad de formar simbiosis
con plantas, ya que estas poseen diferentes estrategias para proteger el complejo
de la enzima nitrogenasa. La tabla 1 ilustra indirectamente la actividad de la
enzima nitrogenasa presente en cada cepa aislada. Se puede denotar que la cepa
CP47 es la que mayor concentración de etileno logró reducir a partir del acetileno
inyectado, lo cual indica que puede fijar nitrógeno a taza de hasta 30 nmol h -1mL-1.
En la filósfera (parte interior y exterior de las hojas), se ha reportado que las
bacterias fijadoras de nitrógeno de vida libre son las pigmentadas, como por
ejemplo, Methylobacterium mesophilicum y Pseudomonas syringae, a las cuales
se les ha atribuido una mejor adaptación a los rayos solares (Hirano y Upper,
2000). Dentro de las relaciones favorables en la rizósfera se encuentra la
asociación con especies de Azospirillum, Enterobacter, Klebsiella, Pseudomonas y
Burkholderia (Estrada et al., 2001). También se ha encontrado que las bacterias
fijadoras de nitrógeno incrementan la capacidad radical de absorción de nitrato,
indirectamente como una consecuencia de la estimulación del desarrollo radical y
directamente por estimulación del sistema transportador del compuesto.
Este hecho no se puede pasar por alto cuando los microorganismos están
presentes en el suelo con algún contaminante orgánico, de modo que idealmente
se podría inferir que las cepas aisladas pueden fijar nitrógeno, y a su vez expresar
genes específicos para poder metabolizar al Clorpirifos.
TABLA 1. Actividad de la enzima nitrogenasa de las cepas aisladas
Medio de Rennie
Semisólido
Líquido
Cepa
CP30
CP37
CP46
CP47
CP30
CP37
CP46
CP47
Concentración de C2H4
(nmol h-1 mL-1)
15.500 a
23.915 abc
28.343 bc
30.115 c
15.057 a
16.829 a
18.157 ab
24.800 abc
Conclusión
Considerando que la biorremediación es una estrategia eficaz en donde el
potencial metabólico microbiano se puede aprovechar para su aplicación in-situ y
ex-situ, las cepas aisladas presentaron características deseables para ser usadas
como estrategias de detoxificación de suelos contaminados con Clorpirifos. Hay
muchos estudios que muestran la degradación completa de Clorpirifos en cultivos
en líquido y en suelo, dentro de ellos se incluyen algunas especies de los géneros
encontrados con la identificación molecular.
Es posible que las especies encontradas contengan genes que codifican para
proteínas degradadoras, lo cual es una opción viable para lograr la mineralización
completa de Clorpirifos. De manera que la inducción de la expresión de genes
570
deseados en microorganismos endémicos también puede aumentar el uso
potencial de biorremediación en desintoxicar ambientes contaminados.
Finalmente, es de gran importancia encontrar géneros bacterianos con
múltiples funciones dentro del micro ecosistema que es el suelo. Más aún si estos
favorecen su fertilidad fijando nitrógeno a tazas considerables, además de realizar
labores de degradación de compuestos tóxicos.
Agradecimientos
Se agradece el apoyo obtenido del PNPC para la realización de esta fase del
proyecto de investigación.
Referencias
Awad, N.S.; Sabit, H.H.; Abo-Aba, S.E.M.; Bayoumi, R.A. (2011). Isolation, characterization and
fingerprinting of some chlorpyrifos-degrading bacterial strains isolated from Egyptian
pesticides-polluted soils. Afr. J. Microbiol. Res. 5: 2855-2862.
Barman, D.N.; Haque, M.A.; AsrafulIslam, S.M.; Yun, H.D.; Kim, M.K. (2014). Cloning and
expression of ophB gene encoding organophosphorus hydrolase from endophytic
Pseudomonas sp. BF1-3 degrades organophosphorus pesticide chlorpyrifos. Ecotoxicology
and Environmental Safety 108: 135-141
Chishti, Z.; Hussain, S.; Arshad, K.R.; Khalid, A.; Arshad, M. (2013). Microbial degradation of
chlorpyrifos in liquid media and soil. Journal of Environmental Management 114: 372-380.
Cycon, M.; Zmijowska, A.; Wójcik, M.; Piotrowska-Seget, Z. (2013). Biodegradation and
bioremediation potential of diazinon-degrading Serratia marcescens to remove other
organophosphorus pesticides from soils. Journal of Environmental Management 117: 7-16.
Comisión intersecretarial para el control del proceso y uso de plaguicidas, fertilizantes y sustancias
toxicas (CICOPLAFEST). 2004. Catálogo oficial de plaguicidas. México 483 pp.
Estrada, P.; Bustillos, R.; Caballero, J. (2001). Burkholderia, a genus rich in plant-associated
nitrogen fixers with wide environmental and geographic distribution. Appl. Environ. Microbiol.
67 (6): 2790-2798.
Fang, H.; Yu, X.Q.; Wang, X.G.; Yang, X.; Yu, J.Q. (2009). Degradation of, chlorpyrifos in laboratory
soil and its impact on soil microbial functional diversity. Journal of Environmental Sciences. 21:
380-386.
García-Rivero M. and Peralta-Pérez M. R. 2008. Cometabolism in the biodegradation of
hydrocarbons. Revista mexicana de ingeniería química 7 (1): 1-12.
Hirano, S.S.; Upper, C.D. (2000). Bacteria in the leaf ecosystem with emphasis on Pseudomonas
syringae a pathogen ice Nucleus, and epiphyte. Microbiol. Mo. Biol. Rev. 64 (3): 624-653.
Imfeld, G.; Vuilleumier, S. (2012). Measuring the effects of pesticides on bacterial communities in
soil: A critical Review. European Journal of Soil Biology 49: 22-30.
Kalia, A.; Gosal, S.K. (2011). Effect of pesticide application on soil microorganisms, Archives of
Agronomy and Soil Science, 57 (6): 569-596.
Korade, D.L.; Fulekar, M.H. (2009). Rhizosphere remediation of chlorpyrifos in mycorrhizospheric
soil using ryegrass. J. Hazard. Mater. 172: 1344-1350.
Lakshmi, C.V.; Kumar, M.; Khanna, S. (2008). Biotransformation of chlorpyrifos and bioremediation
of contaminated soil. International Biodeterioration and Biodegradation 62: 204-209.
Li, X.; He, J.; Li, S. (2007). Isolation of a chlorpyrifos-degrading bacterium, Sphingomonas sp. strain
Dsp-2, and cloning of the mpd gene. Res. Microbiol. 158: 143-149.
Li, X.; Jiang, J.; Gu, L.; Ali, S.W.; He, J.; Li, S. (2008). Diversity of chlorpyrifos-degrading bacteria
isolated from chlorpyrifos-contaminated samples. International Biodeterioration and
Biodegradation 62: 331±335.
Lo, C-C. (2010). Effect of pesticides on soil microbial community. Journal of Environmental Science
and Health, Part B: Pesticides, Food Contaminants, and Agricultural Wastes. 45 (5): 348-359.
571
Moreno-Medina, D.A.; Sánchez-Salinas. E; Ortíz-Hernández, M.L. (2014). Removal of methyl
parathion and coumaphos pesticides by a bacterial consortium immobilized in Luffa cylindrical.
Revista Internacional de Contaminación Ambiental 30: 51-63.
Ning, J.; Bai, Z.; Gang, G.; Jiang, D.; Hu, Q.; He, J.; Zhang, H.; Zhuang, G. (2010). Functional
assembly of bacterial communities with activity for the biodegradation of an organophosphorus
pesticide in the rape phyllosphere. FEMS Microbiology Letters 306: 135-143
Red de Acción en Plaguicidas y sus Alternativas-América Latina (RAP-AL). (2012). Clorpirifos: un
posible
COP
a
nivel
mundial.
Disponible
en
línea:
http://www.rapal.org/articulos_files/Clorpirifos_Rev_.pdf. Consultado en 2014.
Rennie, R.J. (1981). A single medium for the isolation of acetylene-reducing (dinitrogen-fixing)
bacteria from soils. Can. J. Microbiol. 27:8-14.
Sasikala, C.; Jiwal, S.; Rout, P.; Ramya, M. (2012). Biodegradation of chlorpyrifos by bacterial
consortium isolated from agriculture soil. World J. Microbiol. Biotechnology. 28: 1301-1308.
Tamura, K.; Stecher, G.; Peterson, D.; Filipski, A.; Kumar, S. (2013). MEGA6: Molecular
Evolutionary Genetics Analysis version 6.0. Molecular Biology and Evolution 30: 2725-2729.
Tomlin, C. (1994). The Pesticide Manual. The British Crop Protection Council, Surrey, UK and the
Royal Society of Chemistry, Cambridge, United Kingdom.
Xu, G.M.; Li, Y.Y.; Zheng, W.; Peng, X.; Li, W.; Yan, Y.C. (2007). Mineralization of chlorpyrifos by
co-culture of Serratia and Trichosporon sp. Biotechnology Letters. 29: 1469-1473.
Yadav, M.; Srivastva, N.; Singh, R.S.; Upadhyay, S.N.; Dubey, S.K. (2014). Biodegradation of
chlorpyrifos by Pseudomonas sp. in a continuous packed bed bioreactor. Bioresource
Technology 165: 265-269.
Yang, C.; Liu, N.; Guo, X.M.; Qiao, C.L. (2006). Cloning of mpd gene from a chlorpyrifos degrading
bacterium and use of this strain in bioremediation of contaminated. FEMS Microbiology Letters
265: 118-125.
Zhong, Y.; Zou, S.; Lin, L.; Luan, T.G.; Qiu, R.; Tam, N.F.Y. (2010). Effects of pyrene and
fluornthene on the degradation characteristics of phenanthrene in the cometabolism process
by Sphingomonas sp. Strain PheB4 isolated from mangrove sediments. Marine Pollution
Bulletin 60: 2043-2049.
572
LISTA DE COAUTORES:
David Antonio Moreno-Medina, Candidato a Doctor en Ciencias, autor presencial
Misma dirección que Dr. Julián Delgadillo.
Dr. Julián Delgadillo-Martínez
Colegio de Postgraduados, Campus Montecillo. Posgrado en Edafología. 56230
Montecillo, Texcoco, Edo. de México, Carretera México - Texcoco Km 36.5.
México
Tel.: (595) 95 20 200 Ext. 1280
Dra. Rosa María López-Romero
México
Tel.: (595) 95 1 14 75 o 95 20 200 Ext. 1238 o 1237
Dr. David Espinosa-Victoria
México
Tel.: (595) 95 20 200 Ext. 1273
Dr. Oscar García-Barradas
Unidad de Servicios de Apoyo en Resolución Analítica (SARA), Universidad
Veracruzana.
Tel.: (228) 842-1700, ext. 13916
Dra. Ma. Remedios Mendoza-López
Unidad de Servicios de Apoyo en Resolución Analítica (SARA), Universidad
Veracruzana
Tel.: (228) 842-1700, ext. 13556
573
CHAPTER 6.3. ISOLATION AND CHARACTERIZATION OF ARSENICRESISTANT ENDOPHYTIC BACTERIA ASSOCIATED WITH PLANT GROWING
ON MINE TAILINGS
Brenda Román-Ponce* (1); Salvador Embarcadero-Jiménez (1);
María Soledad Vásquez-Murrieta (2); En Tao Wang (1)
(1)Escuela Nacional de Ciencias Biológicas del IPN, Dept. of Microbiology, Microbial Ecology
Laboratory, México D.F., México.
(2) Escuela Nacional de Ciencias Biológicas del IPN, Dept. of Microbiology, Industrial Microbiology
Laboratory, México D.F., México.
ABSTRACT
Mexico is one of the most important mining countries in the world; therefore huge
mine spoils can be found throughout the country. Some endemic plant species that
tolerate high concentration to heavy metals have been found in the mine spoils.
These plants may harbor microbial communities with efficient capacity to detoxify
metals due transforming them into insoluble salts or relatively oxidation states nontoxics. Until now, there are no reports of endophytic bacteria associated with plant
resistant to heavy metals in mine tailings from Villa de la Paz, San Luis Potosí, so
the aim of this work was isolation, characterization and evaluate the capacity to
arsenic transforming of these bacteria.
A total of 52 morphologically different colonies were isolated from the root
samples of Prosopis sp. and Spharealcea angustifolia. The most abundant colony
morphology of the strains was circular, convex and entire type for 80% the strains,
which had irregular colonial margins; the remaining 20% had irregular colonial
margins. Cellular shapes of 75% the strains were found to be Gram positive bacilli,
23% of them were Gram positive cocci and only 2% were Gram negative bacilli.
The endophytic bacteria showed a high degree of resistance to heavy metals,
especially to arsenic, zinc, and copper. More than 40 % isolates showed heavy
metals resistance in liquid MES buffered minimal medium (MBMM) containing
heavy metal ions at 1 mM for Cu2+, Zn2+ and Pb2+, 50 mM for As5+and 10 mM As3+,
respectively. Some of endophytic bacteria had promoted plant growht
characteristics. Arsenic transformation capacity of 30 isolates was tested on CDM
supplemented with 2 mM of As3+ and As5+. Fifty-seven percent of the tested isolates
were capable of transforming arsenate to arsenite based on the formation of redbrownish precipitate in the CDM medium extract. Our results demonstrated that the
endophytic bacteria isolated carried the innate capability of expressing multiple
heavy metal resistance. In addition, some of them could transform arsenic, which is
a feature with potential in the remediation of sites contaminated with this harmful
element.
Key words: endophytic bacteria, heavy metals, arsenic-resistant bacteria
transformation
*Author for correspondence: [email protected]
574
Introduction
Arsenic (As) is known to be one of the most hazardous substances in the
environment and the attention of which has increased in the past two decades due
to the mounting number of affected people exposed to it (Hughes et al., 2011). In
addition to naturally occurring As, it is estimated that over 80 % of all the As ever
produced by man has dissipated to the environment (Nriagu et al., 2007).
Antrophogenic sources including food additives, coal burning (Nordstrom, 2002),
mining, smelting, ore processing, and using As pesticides in agriculture and
chromated copper arsenate in wood preservation. Arsenate (As 5+) and arsenite
(As3+) are reported to be more predominantly present in polluted soils (Pacyna and
Pacyna, 2001).
Microorganisms have an important role in the cycling of As in the environment
as these can either oxidize As3+ to As5+ or reduce As5+ to As3+. The more
widespread chromosomal and plasmid Ars system takes up As5+ into the cell and
reduces it to As3+ and expel the As3+ through an ATP-dependent efflux pump.
Anaerobic periplasmic arsenate reductase also occurs in few bacteria, which
enables these bacteria to use arsenate as terminal electron donor (Silver et al.,
2002). These mechanisms affect speciation and mobility of As which can therefore
affect its bioavailability and toxicity in soils (Cai and Ma, 2003; Oremland et al.,
2005; Borch et al., 2010).
It becomes essential to mitigate the hazardous As from the contaminated soils.
As cannot be easily degraded as any other organic pollutants, and thus, requires
apposite methods for its removal (Rajkumar et al., 2009). Remediation
technologies such as physical and chemical techniques are not only costly but also
affect soil physical, biological, and chemical properties (Pulford and Watson, 2003).
An alternative technology that has advanced in recent years is phytoremediation, a
promising technology for As remediation, has attracted much attention. Compared
with traditional approaches, it has the advantages of being low cost and
ecologically benign (Chen et al., 2010). However, there are several limiting factors
affecting its effectiveness (Shin et al., 2012). To enhance phytoremediation, more
researchers have focused on the roles of plant-associated microbes, which impact
the efficiency of metal extraction by plants (Ma et al., 2009; Weyens et al., 2009).
Endophytic bacteria, which have great potential in enhancing phytoremediation
(Ryan et al., 2008; Weyens et al., 2009), can colonize internal parts of plants
without causing adverse effects on host plants (Khan and Doty, 2009; Long et al.,
2011). Some endophytes promote plant growth by improving mineral nutrient
uptake and increasing resistance to metals (Luo et al., 2011b). One reason why its
potential in phytoremediation has been explored is that endophytes have more
competitive advantages than rhizosphere bacteria. In addition to closer relationship
with host plants, endophytes suffer from fewer restrictions and enjoy preferable
protection (Ma et al., 2009). It has been shown that endophytes alleviate plants
from metal toxicity and enhance phytoremediation. For example, Cd-resistant
endophytes from Sedum alfredi decrease Cd toxicity by increasing Zn and Fe
uptake (Long et al., 2011). The Ni-resistant endophytes from Alyssum
serpyllifolium enhance plant biomass and Ni accumulation in plant tissues (Ma et
al., 2009). Similarly, Pb-resistant endophytes from Brassica napus promote plant
575
growth and plant Pb uptake (Sheng et al., 2008).
Until now, there are little information is available about the arsenic-resistant
endophytic bacteria associated with plant growing on mine tailings in Mexico, and
no any reports about these bacteria in Villa de la Paz, San Luis Potosí. Aiming of
this study were (a) Isolate and characterize arsenic-resistant endophytic bacteria,
(b) evaluate their ability in As tolerance and As transformation for the endophytic
bacteria from two plant species grown in one mine tailing of Villa de la Paz, San
Luis Potosí. Our findings substantiate the potential application of native bacterial
species in the detoxification of As in contaminated soil environments.
Sampling site. The sampling site is located in Villa de la Paz in the state of San
Luis Potosí, Mexico (23.7 Latitude Nort, 178.7 Longitude West). The altitude of the
sampling sites ranges from 1557 to 1830 m above the sea level with a mean
annual temperature of 18°C and average annual precipitation of 486 mm. In the
mine tailing of this site, Arsenic above 8420 mg kg-1, Pb above 754 mg kg-1, Cu
greater than 1154 mg kg-1 and Zn above 1386 mg kg-1 have been detected
(Franco-Hernández et al., 2010). Samples were collected in two areas: a
contaminated area in mine tailings and an uncontaminated place out of the mines
where settlement is antropogenic. Two endemic and common plant species in both
areas were sampled (Prosopis sp. and Spharealcea angustifolia. Once collected
the samples were stored in polyethylene bags at 4°C until further analysis.
Isolation and preservation of the endophytic bacteria. For isolation of
endophytic bacteria, the roots (1 g) were desinfested following the protocol
described by Márquez-Santacruz et al., (2010). Subsequently the samples were
ground with 9 mL of 0.85% sodium chloride under aseptic condition (Barzanti et al.,
2007). Aliquots (100 µL) of tissue extracts (10-1 a 10-3) were plated in triplicate on
TSA plates. To confirm that the sterilization process was successful, the aliquots of
the sterile distilled water used in the final rinse were set on TSA plates and
incubated under the same conditions.
The plates inoculated with root extracts were incubated at 28°C for one to 14
days (Márquez-Santacruz et al., 2010). Single colonies with different morphologies
were selected and purified. The conservation of the isolates was performed by
growing the bacteria on TSA plates and then in 5 mL of nutrient broth at 28 °C for
24-48 h. The bacterial cells were settled down in Eppendorf tubes by centrifugation,
re-suspended in 1 mL of the nutrient broth with 50% glycerol (w/v), and stored at 70°C.
Heavy metal resistance Determination. Toxicity of the select metals to the
bacterial isolates was determinated using 96-well microliter plate. Each well was
filled with 190-ȝl sterile MES buffered minimal medium (Rathnayake et al., 2013)
and supplemented with three concentrations of each heavy metals Cu2+, Zn2+ and
Pb (0.025-1 mM), As3+ (1-10 mM), and As5 + (5-50 mM). Strains were grown in 5
mL TSI medium without heavy metals for 24 h at 28 ±2 °C on a rotary shaker (150
rpm). Ten microliters of bacterial inoculum (1.0 OD at 600 nm) was placed in each
576
respective well. Medium without metal but the bacterial inoculum (bacterial growth
control), and medium with metal but without bacteria (abiotic control) were
included. Plates were incubated at 28 ± 2 °C. Bacterial growth after 4 days was
measured using an EZ Read 400 Microplate Reader (Biochrom) at 620 nm.
Genomic DNA extraction, PCR Amplification, and sequencing of bacterial
16S rDNA. Genomic DNA of 30 arsenic-resistant endophytic bacteria was
H[WUDFWHG IURP HDFK LVRODWH LQ P/ RI 76, PHGLXP DW Û& ZLWK DJLWDWLRQ K
using the protocols described by Zhou et al., (1995), and was used as template to
amplify 16S rRNA genes. The 16S rRNA gene was amplified by PCR with a
thermocycler (Maxygene Thermal Cycler Therm 1061 Axygen Scientific) using an
initial denaturing step of 5 min al 94°C followed by 30 cycles of 45 s at 94°C, 1 min
of annealing at 57°C, 90 s extension at 72°C and a final polymerization step for 8
min at 72°C. The PCR mixture (25 µL) containing 10-100 ng of isolated DNA, 1.5
mM MgCl2, 2.5 U Taq DNA polymerase (Invitrogen, USA), 1u PCR buffer, 100
pmol of primers fD1 (5'-AGA GTT TGA TCC TGG CTC AG-3') and rD1 (5'-AAG
GAG GTG ATC CAG CC-3') (Weisburg et al., 1990) and 200 ȝM of each dNTP.
The PCR products were visualized after electrophoresis in agarose gel 1% (w/v) in
the buffer of 0.5u TAE, by staining with an aqueous solution of ethidium bromide
(0.5 ȝg mL-1).
For sequencing reaction, the amplified 16S rDNA was purified with a
commercial kit PURE LINK (Invitrogen 310002) and sequenced under Big Dye TM
terminator cycling conditions with the same primers using Automatic Sequencer
3730XL in Macrogen (Korea). The sequences obtained were compared with those
in the GenBank database using the program BLAST (Altschul et al., 1997;
http://blast.ncbi.nlm.nih.gov/Blast.cgi). The sequences were aligned using
CLUSTAL X (2.0) software (Thompson et al., 1997). The presence of chimerical
sequences was checked with the RDP Chimera Check program. The acquired
sequences were manually edited with SEAVIEW software (Galtier et al., 1996).
The jMODELTEST 3.06 software (Posada, 2008) was used to select appropriate
models of sequence evolution by the AIC (Akaike Information Criterion).
Similarities among sequences were calculated using the MatGAT v.2.01 software
(Campanella et al., 2003). Taxonomic assignment was obtained by using the
Roselló-Mora prokaryotes criteria (Roselló-Mora and Amann, 2001).
Plant Growth Promoting Characters. Siderophore production was performed
following the protocol described by Schwyn & Neilands (1987) and Perez-Miranda
et al., (2007). Mineral phosphate solubilization activity was assayed according to
Kuklinsky-Sobral et al., (2004). Indoleacetic acid (IAA) production was analysed
using a modification of the qualitative method developed by Bric et al., (1991) and
the ability of nitrogen fixing was estimated according Rodriguez-Caceres, (1982).
After being treated with Salkowski reagent for 30 min, appearance of a pale pink
halo on the filter paper from LB medium was positive for the IAA assay. An orange
annulus appeared around the colony was defined as positive for siderophore
production. Growth in the nitrogen-free medium was taken as indicator of nitrogen
fixation. Presence of a clear halo around the colony in the medium supplemented
with inorganic phosphate was considered as positive for solubilisation of phosphate.
577
Arsenic-transforming ability. Bacterial strains growing at 10 mM and 50 mM of
As3+ and As5+, respectively, were used for arsenic transformation ability. Bacterial
strains were inoculated in vials containing 5 mL of chemically defined medium
(CDM) as defined by Weeger et al. (1999), amended with 2 mM of NaAsO2 or
NaH2AsO4· 7H2O and incubated at in the dark at room temperature for 4 to14
days. After this time, the cell suspension was centrifuged at 4,7500 g for 10 min.
Each well of a 96-well microliter plate was filled with 100 ȝL of the each culture
supernatant and 100 ȝL of the 0.1 M AgNO3 was added to each well-plate. The
arsenite-oxidizing reaction was indicated by a change of the medium to a bright
yellow color to brownish color, while the change of a brownish color to a bright
yellow color indicated an arsenate-reducing reaction (Lett et al., 2001). For As3+
oxidation test Pseudomonas aeruginosa ATCC 25619 was used and for As 5+
reduction test, Escherichia coli DH5D was used as positive strain.
Isolation of endophytic bacteria. Although special attention has been paid to
endophytic bacteria isolated from plants in special environments, such as heavy
metals contaminated soils (Idris et al., 2004; Barzanti et al., 2007; Sheng et al.,
2008; Guo et al., 2010; Zhang et al., 2011; Shing et al., 2012), our study on the
communities of endophytic bacterial of two heavy metals tolerant plants grown in
mine tailings offered some interesting information. In the present study, we
obtained fifty-two endophytic bacteria isolated, 36 from unpolluted site and 16 from
polluted area. (Table 1). Based upon the microscopic observation, the isolates
were divided into actinomycetes, bacilli and cocci, and the Gram-positive bacteria
were predominant (over 50 percent). These results were consistent with the
previous reports that more than 90% of the endophytic bacteria associated with
heavy metal-resistant plants were Gram-positive bacteria (Barzanti et al., 2007;
Guo et al., 2010; Sun et al., 2010).
Heavy metals resistance. This study represents an attempt to evaluate the multiheavy metals resistance to endophytic baceterial associated with the native plants
Prosopis sp and Spharealce angustifolia.The endophytic bacteria showed a high
degree of resistance to heavy metals assayed, especially to Cu, As+3and As+5. The
MIC of the 52 isolates in minimal medium buffered MES was 1 mM Cu 2+ (CuSO4),
Zn2+ (ZnSO4) and Pb2+ (PbNO3), 10 mM As3+ (NaAsO2) and 50 mM As5+
(NaH2AsO4) (Table 1). The order of the toxicity of the metals to the isolates was
Pb> Zn> Cu> As3+ > As5+. The Tukey test showed significant differences between
sampling sites, plants species and isolates (data not show). A direct comparison of
578
TABLA 1. Morphological and multi heavy metals resistance of the endophytic
bacteria isolate
Isolate
Site
Plant
Colony
pigmentation
Shape
Gram
reaction
Heavy metals resistance (mM)
Cu+2
Zn+2
Pb+2
As+3
As+5
NS1E1g
S1
S. angustifolia
beige
cluster cocci
Gram +
1
1
1
10
50
NS1E3
S1
S. angustifolia
light yellow
small road
Gram +
1
1
0.5
>10
50
NS1E7
S1
S. angustifolia
light yelow
small road
Gram +
1
0.5
0.5
5
50
NS2E1
S1
S. angustifolia
yellow
tetrads cocci
Gram +
1
1
10
50
NS2E2
S1
S. angustifolia
maize yellow
small road
Gram +
1
1
>10
>50
NS2E3
S1
S. angustifolia
bright yellow
small road
Gram +
1
1
>10
>50
NP1E2
S1
Prosopis sp
cream
road
Gram +
1
1
0.5
1
50
NP1E3
S1
Prosopis sp
cream
road
Gram +
1
1
1
10
50
NP1E4
S1
Prosopis sp
light orange
small road
Gram +
1
1
1
1
20
NP1E5
S1
Prosopis sp
cream
road
Gram +
1
1
1
10
50
NP2E1
S1
Prosopis sp
yellow
tetrads cocci
Gram +
1
1
0.5
10
50
NP2E2
S1
Prosopis sp
bright tun
road
Gram -
1
0.5
0.5
5
20
NP2E4
S1
Prosopis sp
bright run
Road
Gram -
1
1
0.5
5
50
NP2E5
S1
Prosopis sp
cream
big road
Gram +
1
1
1
10
50
NP2E6
S1
Prosopis sp
cream
sporulated
road
Gram +
1
1
1
10
50
NP2E7
S1
Prosopis sp
bright tun
road
Gram -
1
1
1
10
50
NP2E8
S1
Prosopis sp
cream
road
Gram +
1
1
1
10
20
NP2E9
S1
Prosopis sp
cream
road
Gram +
1
1
1
10
50
NP2E10
S1
Prosopis sp
ceige
cocci
Gram +
1
0.5
1
10
50
NP2E11
S1
Prosopis sp
cream
sporulated
road
Gram +
1
0.5
1
10
20
NP2E12
S1
Prosopis sp
cream
road
Gram +
1
1
1
5
50
NP2E13
S1
Prosopis sp
pale cream
road
Gram +
1
1
1
>10
50
NP2E14
S1
Prosopis sp
cream
road
Gram +
1
0.5
1
5
50
NP2E15
S1
Prosopis sp
cream
road
Gram +
1
1
1
1
50
NP2E16
S1
Prosopis sp
white
chain road
Gram +
1
1
1
>10
50
NP2E17
S1
Prosopis sp
cream
road
Gram +
1
1
1
>10
50
NP2E18
S1
Prosopis sp
cream
sporulated big
road
Gram +
1
1
1
10
>50
1
1
1
579
Tabla 1. Continuated
Isolate
Site
Plant
Colony
pigmentation
Shape
Gram
reaction
Heavy metals resistance (mM)
Cu+2
Zn+2
Pb+2
As+3
As+5
NP3E1
S1
Prosopis sp
pale cream
road
Gram +
1
0.5
0.5
5
50
NP3E2
S1
Prosopis sp
white
road-cocci
Gram +
1
0.25
0.5
>10
5
NP3E3
S1
Prosopis sp
peach
sporulated
road
Gram +
1
0.5
1
5
>50
NP3E7
S1
Prosopis sp
cream
road
Gram +
1
1
1
>10
50
NP3E9
S1
Prosopis sp
peach
sporulated
road
Gram +
1
1
0.5
1
50
NP3E10
S1
Prosopis sp
cream
sporulated
road
Gram +
1
0.5
1
5
50
NP3E11
S1
Prosopis sp
cream
sporulated
road
Gram +
1
0.5
0.05
10
50
CS1E1
S2
S. angustifolia
bright tun
road
Gram -
1
1
0.1
10
50
CS2E1
S2
S. angustifolia
white
road-cocci
Gram +
1
1
1
>10
>50
CS3E1
S2
S. angustifolia
beige
cluster cocci
Gram +
1
1
1
>10
>50
CS3E2
S2
S. angustifolia
peach
road
Gram +
1
1
1
>10
>50
CS3E3
S2
S. angustifolia
peach
road
Gram +
1
1
1
>10
>50
CS3E4
S2
S. angustifolia
beige
cluster cocci
Gram +
1
1
0.1
0.0125
50
CS3E5
S2
S. angustifolia
beige
cluster cocci
Gram +
1
1
0.1
0.5
5
CP1E1
S2
Prosopis sp
yellow
cocci
Gram +
1
1
0.5
10
20
CP1E2
S2
Prosopis sp
beige
chain road
Gram +
1
1
0.5
10
50
CP1E3
S2
Prosopis sp
beige
chain road
Gram +
1
1
1
10
>50
CP1E4
S2
Prosopis sp
white
chain road
Gram +
1
1
1
>10
>50
CP1E5
S2
Prosopis sp
light beigl
large chain
road
Gram +
1
0.5
1
10
50
CP1E7
S2
Prosopis sp
black pigmented
road
Gram +
1
1
1
10
20
CP1E8
S2
Prosopis sp
light yellow
small road
Gram +
1
0.5
0.5
10
50
CP2E1
S2
Prosopis sp
light tun
cluster cocci
Gram +
0.5
1
0.5
0.5
50
CP2E2
S2
Prosopis sp
peach
sporulated
road
Gram +
1
1
1
>10
20
CP2E3
S2
Prosopis sp
peach
sporulated
road
Gram +
5
5
1.5
>10
50
CP2E4
S2
Prosopis sp
beige
cluster cocci
Gram +
5
0.5
0.05
0.5
50
CP2E4c
S2
Prosopis sp
tun
cluster cocci
Gram +
5
0.5
0.05
0.5
5
580
our results with the other studies is difficult because (a) different plants species
were used, and (b) bioavailability of the metals to bacteria was different, resulting
in the difference in the MICs of the metals (Viti and Pace, 2003; Kamala-Kannan
Isolate
Porcentage
similarity (%)
Closest neighbour
Taxonomic asignation
Arsenic transforming test
and Krishnamorthy, 2006).
A metal hyperaccumulating plant could accumulate huge amounts of heavy
metals also they can accumulate more than one metal and can therefore provide a
specific niche for endphytic bacteria. Recently, considerable attention was focused
on the mechanisms that the endophytes may be developed for withstand high
heavy metal concentrations (with an effective metal reduction or oxidation,
biosorption, inmobilization, intracellular accumulation, intracellular sequestration
and extracellular precipitation) (Kuiper et al., 2004). For instance, endophytic
bacteria isolated from Sedum alfredii and Alyssum bertolonii were resistant to more
than one metals (Barzanti et al., 2007; Xixian et al., 2011). In the present study,
some endophytic bacteria isolates from Prosopis sp and Spharealcea angustifolia
also exhibited resistance to more than one metal (Table 1). Comparing these
studies mentioned above, we observed that the endophytic bacteria isolated from
Prosopis sp and S. angustifolia have a higher minimum inhibitory concentration in
minimal medium buffered MES. However, the endophytic bacteria have an MIC (1
mM) for Cu lower than those reported by Luo et al., (2010). When the minimum
inhibitory concentration for Zn for these isolates was compared, it was also noted
that this was lower (1 mM) than that (15 mM) reported by Barzanti et al. (2007).
The difference may be due to factors such as the sampling site, the plant hosts, the
pollutant concentration and type of metal. In the particulary case of arsenic, the
isolate NS2E2 and NS2E3 were resistant more than 50 mM in MMBM, which is a
high level never reported previously for the endophytic bacteria.
Identification of arsenic-resistant endophytic bacterial isolates. The 16S rRNA
sequences of all 30 isolates were subjected to nucleotide BLAST and the bacteria
were classified according to their similarity to sequences in the GenBank database.
Table 2 shows the percentage similarity with the closest neightbour of the arsenic
resistant endophytic bacteria and taxonomic asignation. Phylogenetic analysis
revelated that therty endophytic bacteria belonged to 8 genera coresponding to
Bacillus, Staphylococccus, Pseudomonas, Kocuria, Micrococcus, Microbacterium,
Leucobacter, Arthrobacter (Table 2). Bacillus sp being the most dominant bacteria
(57 %), Microbacterium was the second more dominant genera with 12 % of the
total of arsenic-resistant endophytic bacteria. Several genera, like Bacillus and
Microbacterium have been reported as a metals resistant endophytic (Barzanti et
al., 2007; Luo et al., 2011a). In addition, the arsenic- resistant endophytic bacteria
belonged to Firmicutes (60%), Actinobacteria (33 %) Proteobacteria (7 %). Clearly
Firmicutes occupied the largest proportion of the isolates. Luo et al. (2011a)
isolated 30 Cd-resistant endophytes from S. nigrum, with 43% belonging to
Actinobacteria and 23% Proteobacteria. Barzanti et al. (2007) obtained 83 Niresistant endophytes from A. bertolonii. Firmicutes again is the most dominant
group (60%), followed by Actinobacteria (28%). Similarly, 14 endophytes isolated
from S. alfredii belong to Proteobacteria (57%) and Firmicutes (43%) (Long et al.,
2011). These results suggested that most endophytes belong to Firmicutes,
Proteobacteria and Actinobacteria.
581
Arsenate
reducing
+
Arsenite
oxidizing
-
NS1E3
95.9
Microbacterium arborescens
NS2E1
97.9
Micrococcus luteus
Micrococcus luteus
+
-
NS2E2
97.7
Microbacterium schleiferi
Microbacterium schleiferi
+
-
NS2E3
98.4
Microbacterium oxydans
Microbacterium oxydans
+
-
NP1E3
97.9
Bacillus sp.
-
-
NP1E5
98.9
Bacillus sp.
-
-
NP2E1
98.9
Bacillus malacitensis
Bacillus axarquiensis
Bacillussubtilis
Micrococcus luteus
Micrococcus luteus
+
-
NP2E5
98.6
Bacillus sp.
+
-
NP2E6
97.6
Bacillus megaterium
Bacillus aryabhattai
Bacillus simplex
Bacillus simplex
ng
ng
NP2E7
99.9
Pseudomonas stutzeri
-
-
NP2E9
98.4
Bacillus sp.
-
-
NP2E10
97.7
Kocuria rhizophila
Kocuria rhizophila
+
-
NP2E13
98.7
Leucobacter aridicollis
Leucobacter aridicollis
-
-
NP2E16
99.4
Bacillus endophyticus
-
-
NP2E17
99.7
Bacillus sp.
-
-
NP2E18
99.5
Bacillus sp.
+
-
NP3E2
98.9
Bacillus megaterium
Bacillus aryabhattai
Arthrobacter scleromae
+
-
NP3E7
98.5
Bacillus sp.
-
-
NP3E11
97.5
Bacillus cereus
Bacillus cereus
+
-
CS1E1
99.7
-
-
CS2E1
98.1
ng
ng
CS3E1
99.9
Staphylococcus sp.
+
ng
CS3E2
98.2
Bacillus sp.
ng
ng
CS3E3
99
Staphylococcus warneri
Staphylococcus pasteri
Bacillus sp.
+
-
CP1E2
98.9
Bacillus vallismortis
+
-
CP1E3
99.4
+
-
CP1E4
99
+
-
CP1E5
97.3
Bacillus niacini
Bacillus niacini
-
ng
CP1E8
98.5
+
-
CP2E3
98.5
+
-
TABLA 2. Identification of arsenic-resistant endophytic bacteria and their arsenic
transformation
Note: ng: no growth
Arsenic-resistant endophytic bacteria has a limited potential to promote
plant growht. In stressed environments, another important factor to evaluate for
582
the arsenic-resistant endophytic bacteria was the plant growth promoting (PGP)
characteristics. The potential plant-promoting ability of the 30 arsenic-resistant
endophytic bacteria was qualitatively evaluated, including production of indole
acetic acid (IAA), mineral phosphate solubilization, nitrogen fixation and
siderophores secretion. Tabla 3 showed that in particular, strains NP1E3, NP2E5,
CS3E2, and CS3E3 showed positive for at least a half of evaluated traits.
Plant growth promoting by endophytic bacteria may be exerted by several
mechanisms, e.g. production of plant hormones, siderophores synthesis, nitrogen
fixation, solubilization of minerals such as phosphorous, or via enzymatic activities,
for suppression of ethylene by
1-aminocyclopropane-1-carboxylate (ACC)
deaminase, etc. (Chernin and Chet, 2002). Only three endophytic bacteria
allocated in the genus of Microbacterium and Bacillus were able to produce barely
detectable amounts of IAA (Table 3). The genera mentioned above had already
been reported as IAA producers in other studies on endophytic bacteria associated
with plants grown on mine tailings (Sheng et al., 2008; Zhang et al., 2011).
Siderophore helps plants to overcome adversity of metal stress and iron deficiency
(Qi and Zhao, 2013). However, only five arsenic-resistant endophytic bacteria had
little siderophore production capacity (Table 3), which was different from other
metal-resistant endophytes (Sheng et al., 2008; Luo et al., 2011a).
Some endophytic isolates also demonstrated to solubilize mineral phosphates
(Rodríguez and Fraga, 1999; Verma et al. 2001), suggesting that during the initial
colonization, endophytic bacteria could enhance phosphate availability to the host
plant. Results from Kuklinsky-Sobral et al. (2004) supported this suggestion,
showing that 52 % of the endophytic bacteria isolated from soybean could
solubilize mineral phosphate.
Althought the endophytic bacteria such as
Pseudomonas, Ralstonia, Klebsiella, Erwinia, etc. have been defined as P
solubilizer, there is a litter evidence to show the P solubilizing endophytic bacteria
could improve the P uptake of their host plants. In our study, only the Bacillus
isolates were able to solubilizing mineral phosphate in the medium tested (Table
3). Therefore, it is needed to elucidate the effects of phosphate solubilizing
endophytes on the plant growing in a heavy metals contaminated soil and the
uptake of heavy metals from this endophytes. In the past, most interests has
focused on the fixation of atmospheric nitrogen by free-living endophytic bacteria,
especially of diazotriphs (Döbereiner and Pedrosa, 1987; Hecht- Buchholz 1998;
Estrada et al., 2002; Hurek and Reinhold-Hurek, 2003). In the present study only
17 % of the isolates (Bacillus, Pseudomonas, Leucobacter and Microbacterium)
were able to grow in a nitrogen-free medium.
583
TABLA 3. Plant growht promoting (PGP) characteristics of arsenic-resistant endophytic
bacteria
Isolate
NS1E3 Microbacterium arborescens
NS2E1 Micrococcus luteus
NS2E2 Microbacterium schleiferi
NS2E3 Microbacterium oxydans
NP1E3 Bacillus sp.
NP1E5 Bacillus sp.
NP2E1 Micrococcus luteus
NP2E5 Bacillus sp.
NP2E6 Bacillus simplex
NP2E7 Pseudomonas stutzeri
NP2E9 Bacillus sp.
NP2E10 Kocuria rhizophila
NP2E13 Leucobacter aridicollis
NP2E16 Bacillus endophyticus
NP2E17 Bacillus sp.
NP2E18 Bacillus sp.
NP3E2 Arthrobacter scleromae
NP3E7 Bacillus sp.
NP3E11 Bacillus cereus
CS1E1 Pseudomonas stutzeri
CS2E1 Arthrobacter scleromae
CS3E1 Staphylococcus sp.
CS3E2 Bacillus sp.
CS3E3 Bacillus sp.
CP1E2 Bacillusvallismortis
CP1E3 Bacillus vallismortis
CP1E4 Bacillus endophyticus
CP1E5 Bacillus niacini
CP1E8 Microbacterium arborescens
CP2E3 Bacillus endophyticus2DT
plant growth promoting (PGP) characteristics
Siderophore
P solubilization IAA
N2-Fixing
ng
+
ng
ng
+
ng
+
+
+
ng
+
+
+
ng
ng
ng
ng
+
ng
ng
ng
ng
+
+
+
ng
+
ng
+
ng
ng
+
+
+
+
+
ng
ng
-
Notes:
Siderophore Production : -, no able to producing siderophore
solubilization: -, no able to phosphate solubilization
c IAA production: -, no able to produce IAA
d Nitrogen-fixing assay: - no able to growght in nitrogen-free médium
NG, no growht
a
b phosphate
584
As transformation by endophytes related to their As tolerance. To better
understand the potential role of endophytes in As transformation, 30 bacteria were
tested on CDM supplemented with As3+ or As5+. The As3+ and As5+ were added to
CDM at a final concentration of 2 mM. The cultures were incubated at 28 °C for 4
to 14 days, and transformation ability was evaluated by using the silver nitrate
screening method. The interaction of silver nitrate with arsenite generates a bright
yellow precipitate while a brownish precipitate forms with arsenate (Simeonova et
al. 2004). Fifty-seven percent of the tested isolates were capable of transforming
arsenate to arsenite based on the formation of bright yellow precipitate surrounding
the colonies (Table 1). However, none of the bacteria produced a redbrownish
precipitate, suggesting that no arsenate was generated. This indicated that 17
isolates had reducing activity while none of the colonies displayed arsenic-oxidizing
activity under the conditions of the assay. Unfortunately, this study could not
determine the proportion of arsenite and arsenate from the microplate analyses
because of the rapid color reaction and the lack of color scale standard. However,
our results indicated that the CDM medium assay done using silver nitrate was a
useful and effective tool for rapidly determining arsenic transformation of bacterial
isolates on a large scale. Moreover, application of this method was successful only
when it was tested using the CDM medium (Simeonova et al. 2004).
Conclusions
Our study characterized arsenic-resistant endophytic bacteria associated with
Prosopis sp. and S. angustifolia plants grown in the mine tailings contaminated
with multi heavy metals. The isolated endophytic bacteria had an innate capability
of multiple heavy metal resistance and some plant growth-promoting
characteristics. Bacillus sp was the most dominant species. More than fifty percent
of the isolates had to reduce As5+. More studies should be carried out to
characterize their potential in As tolerance and transformation. Characterization of
arsenic-resistant endophytic bacteria reported here is the first step to use them to
enhance bioremediation of As-contaminated sites. These strains will also be useful
in further research on As tolerance and transformation mechanisms involved in
arsenic-resistant endophytic bacteria.
Acknowledgements
Projects SIP-IPN 20130722 and 20130828 funded the research. B.R.P and S.E.J.
received scholarships support from the CONACyT and PIFI. M.S.V.M and E.T.W.
appreciate the scholarships of COFAA and EDI-IPN and SNI-CONACyT.
585
References
Altschul, SF, Madden TL, Shäffer AA, Zhang, J, Zhang Z. Miller W, Lipman, DJ (1997) Gapped
BLAST and PSI-BLAST: a new generation of protein database search programs. Nucleic Acids
Res 25: 3389±3402.
Barzanti R, Ozino F, Bazzicalupo M, Gabbrielli R, Galardi F, Gonnelli C, Mengon A (2007) Isolation
and Characterization of Endophytic Bacteria from the Nickel Hyperaccumulator Plant Alyssum
bertolonii. Microb Ecol 53: 306±316.
Borch T, Kretzschmar R, Kappler A, Cappellen PV, Ginder Vogel A, Voegelin A, Campbell K (2010)
Biogeochemical redox processes and their impact on contaminant dynamics. Environ Sci
Technol 44: 15±23.
Bric JM, Bostock RM, Silverstone S (1991). Rapid in situ assay for indoleacetic acid production by
bacteria immobilized on a nitrocellulose membrane. Appl Environ Microbiol 57: 535±538.
Cai Y, Ma LQ (2003) Metal tolerance, accumulation and detoxification in plants with emphasis on
arsenic in terrestrial plants. In: Cai Y, Braids O (eds) Biochemistry of environmentally important
trace elements. Oxford University Press, London, pp 95±114.
Campanella JJ, Bitincka L, Smalley J (2003) MatGAT: an application that generates
similarity/identity matrices using protein or DNA sequences. BMC Bioinformatics 4:29.
Chen L, Luo S, Xiao X, Guo H, Chen J, Wan Y, Li B, Xu T, Xi Q, Rao C, Liu C, Zeng G. (2010)
Application of plant growth-promoting endophytes (PGPE) isolated from Solanum nigrum L. for
phytoextraction of Cd-polluted soils. Appl. Soil Ecol. 46: 383±389.
DöbereinerJ, Pedrosa FO (1987) Nitrogen-fixing bacteria in non-leguminous crop plants. Science
Tech, Madison, WI.
Estrada P, Mavingui P, Cournoyer B, Fountaine F, Balandreau J, Caballero-Mellado J (2002) A N2fixing endophytic Burkholderia sp. associated with maize plants cultivated in Mexico. Can J
Microbiol 48: 285-294.
Franco-Hernández MO, Vásquez-Murrieta MS, Patiño-Siciliano A, Dendooven L (2010) Heavy
metals concentration in plants growing on mine tailings in Central Mexico. Biores Technol 101:
3864±3869.
Galtier N, Gouy M, Gautier C (1996) SEAVIEW and PHYLO_WIN: two graphic tools for sequence
alignment and molecular phylogeny. Comput Appl Biosci 12: 543±548.
Guo H, Luo S, Chen L, Xiao X, Xi Q, Wei W, Zeng, G, Liu C, Wan Y, Chen J, He Y (2010)
Bioremediation of heavy metals by growing hyperaccumulaor endophytic bacterium Bacillus sp.
L14. Biores Technol 101: 8599±8605.
Hecht-Buchholz C (1998) The apoplast-habitat of endophytic nitrogen-fixing bacteriaand their
significance for the nitrogen nutrition on nonleguminous plants. Z Pflanzenernähr Bodenkd
161: 509-520.
Hughes MF, Beck BD, Chen Y, Lewis AS, Thomas DJ (2011) Arsenic exposure and toxicology: a
historical perspective. Toxicol Sci 123(2):305±332.
Hurek T, Reinhold-Hurek B (2003) Azoarcussp. strain BH72 as a model for nitrogen-fixing grass
endophytes. J Biotechnol 106: 169-178.
Idris R, Trifonava R, Puschenreiter M, Wenzel W, Sessitsch A (2004) Bacterial communities
associated whit flowering plants of the Ni hyperaccumulator Thlaspi goesingense. Appl Environ
Microbiol.70: 2667±2672.
Kamala-Kannan S, Krishnamoorthy K (2006) Isolation of mercury resistant bacteria and influence of
abiotic factors on bioavailability of mercury ± a case study in Pulicat Lake north of Chennai,
south east India. Sci Total Environ 367: 341±353.
Kuiper I, Lagendijk EL, Bloemberg GV, Lugtenberg BJ. (2004) Rhizoremediation: a beneficial plantmicrobe interaction. Mol Plant Microbe Interact 17: 6±15.
Kuklinsky-Sobral J, Araujo WL, Mendes R, Geraldi IO, Pizzirani-Kleiner AA, Azevedo J L (2004)
Isolation and characterization of soybean-associated bacteria and their potential for plant
growth promotion. Environ Microbiol 6: 1244±1251.
Lett MC, Paknikar K, Lievremont D (2001) A simple and rapid method for arsenite and arsenate
speciation. In: Ciminelli, V. S. T. Jr. & García, O. (Eds) Biohydrometallurgy-fundamentals
technology and sustainable development, Part B. Elsevier Science, New York. 541-546.
586
Long X, Chen X, Chen Y, Woon-Chung WJ, Wei Z, Wu Q (2011) Isolation and characterization
endophytic bacteria from hyperaccumulator Sedum alfredii Hance and their potential to
promote phytoextraction of zinc polluted soil. World J. Microbiol. Biotechnol. 27: 1197±1207.
Luo S, Wan Y, Xiao X, Guo H, Chen L, Xi Q, Zeng G, Liu C, Chen J (2011a) Isolation and
characterization of endophytic bacterium LRE07 from cadmium hyperaccumulator Solanum
nigrum L. and its potential for remediation. Appl Microbiol Biotechnol 89: 1637±1644.
Luo S, Wan Y, Xiao X, Guo H, Chen L, Xi Q, Zeng G, Liu C, Chen J ( 2011b) Isolation and
characterization of endophytic bacterium LRE07 from cadmium hyperaccumulator Solanum
nigrum L. and its potential for remediation. Appl. Microbiol. Biotechnol. 89: 1637±1644.
Ma, Y., Rajkumar, M., Freitas, H., 2009. Isolation and characterization of Ni mobilizing PGPB from
serpentine soils and their potential in promoting plant growth and Ni accumulation by Brassica
spp. Chemosphere 75, 719±725.
Marquez-Santacruz HA, Hernández-León R, Orozco-Mosqueda MC, Velázquez-Sepulveda I,
Santoyo G (2010) Diversity of bacterial endophytes in roots of Mexican husk tomato plants
(Physalis ixocarpa) and their detection in the rhizosphere. Gen Mol Res 9: 2372±2380.
Nordstrom DK (2002) Public health ± worldwide occurrences of arsenic in ground water. Science
296: 2143±2145.
Nriagu JO, Bhattacharya P, Mukherjee AB, Bundschuh J, Zevenhoven R, Loppert RH (2007)
Arsenic in soil and groundwater: an overview. In: Bhattcharya P, Mukherjee AB, Bundschuh J,
Zevenhoven R, Loeppert RH (eds) Arsenic in soil and groundwater environment. Trace metals
and other contaminants in the environment. Elsevier Science Ltd, Oxford, pp 3±60
Oremland RS, Kulp TR, Blum JS, Hoeft, SE, Baesman S, Miller LG, Stolz JF (2005) A microbial
arsenic cycle in a salt-saturated, extreme environment. Science 308: 1305±1308.
Pacyna JM, Pacyna EG (2001) An assessment of global and regional emissions of trace metals in
the atmosphere from anthropogenic sources world. Environ Rev 9:269±298.
Pérez-Miranda S, Cabirol N, George-Téllez R, Zamudio-Rivera LS, Fernández FJ (2007) O-CAS, a
fast and universal method for siderophore detection. J Microb Methods 70: 127-131
Posada D (2008) jModelTest: phylogenetic model averaging. Mol Biol Evol 25: 1253-1256.
Qi W, Zhao L (2013) Study of the siderophore-producing Trichoderma asperellum Q1 on cucumber
growth promotion under salt stress. J. Basic Microbiol. 53: 355±364.
Rajkumar M, Prasad MNV, Freitas H, Ae N (2009) Biotechnological applications of serpentine soil
bacteria for phytoremediation of trace metals. Crit Rev Biotechnol 29:120±130
Rathnayake IVN, Megharaj M, Krishnamurti GSR, Bolan NS, Naidu R( 2013) Heavy metal toxicity to
bacteria ± Are the existing growth media accurate enoughto determine heavy metal toxicity?
Chemosphere 90: 1195-1200.
Rodriguez H, Fraga R (1999) Phosphate solubilizing bacteria and their role in plant growth
promotion. Biotechnol Adv 17: 319±339.
Rodríguez-Cáceres EA (1982) Improved medium for isolation of Azospirillum spp. Appl Environ
Microbiol 44: 990±991
Rosselló-Mora R, Amman R (2001) The species concept for prokaryotes. FEMS Microbiology
Reviews 25: 39±67.
Ryan RP, Germaine K, Franks A, Ryan DJ, Dowling DN (2008) Bacterial endophytes: recent
developments and applications. FEMS Microbiol. Lett. 278: 1±9.
Schwyn, B, Neilands JB (1987) Universal chemical assay for the detection and determination of
siderophores. Anal Biochem 160: 47±56.
Sheng XF, Xia JJ, Jiang CY, He LY, Quian M (2008) Characterization of heavy metal-resistant
endophytic bacteria from rape (Brassica napus) roots and their potential in promoting the
growth and lead accumulation. Environ Poll 156: 1164- 170.
Shing MN, Shim J, You Y, Myung M, Bang KS, Choa M, Kamala-Kannan S, Oh BT (2012)
Characterization of lead resistant endophytic Bacillus sp. MN3-4 and its potential for promoting
lead accumulation in metal hyperaccumulator Alnus firma. J Hazard Mater 199-200: 314±320.
Silver S, Phung LT (2005) Genes and enzymes involved in bacterial oxidation and reduction of
inorganic arsenic. Appl Environ Microbiol 71:599±608.
Simeonova DD, Lievremont D., Lagarde F, Muller DAE., Groudeva VI, Lett MC (2004). Microplate
screening assay for the detection of arsenite-oxidizing and arsenate-reducing bacteria. FEMS
Microbiol. Lett. 237: 249-253.
587
Sun LN, Zhang YF, He LY, Chen ZJ, Wang QY, Quian M, Shen, XF (2010) Genetic diversity and
characterization of heavy metal-resistant-endophytic bacteria from two copper-tolerant plant
species on copper mine wasteland. Biores Technol 101: 501±509.
Sun Y, Zhou Q, Wang L, Liu W (2009) Cadmiun tolerance and accumulation characteristics of
Bidens pilosa L. as a potential Cd-hyperacummulator. J Hazard Mater 161: 808-814.
Thompson JD, Gibson T.J, Plewniak F, Jeanmougin F, Higgins D (1997) The CLUSTAL_X windows
interface: flexible strategies for multiple sequence alignment aided by quality analysis tools.
Nucleic Acids Res 25: 4876±4882.
Verma SC, Ladha JK, Tripathi AK (2001) Evaluation of plant growth promoting and colonization
ability of endophytic diazotrophs from deep water rice. J Biotechnol 91:127±141.
Viti C, Pace A, Giovannetti L (2003) Characterization of Cr (VI)-resistant bacteria isolated from the
chromium-contaminated soil by tannery activity, Curr Microbiol 46 :1±5.
Weeger W, Lievremont D, Perret ML, Fagarde JC, Hubert M, Lett MC (1999). Oxidation of arsenite
to arsenate by bacteria isolation from an aquatic environment. Bio Metals. 12:141-149.
Weisburg WG, Barns SM, Pelletior DA, Lane DJ (1991) 16S ribosomal DNA amplification for
phylogenetic study. J Bacteriol 173:697±703.
Weyens N, van der Lelie D, Taghavi S, Vangronsveld J, (2009). Phytoremediation: plant-endophyte
partnerships take the challenge. Curr. Opin. Biotechnol. 20: 248±254.
Xinxian L, Xuemei C, Yagang C, Woon-Chung WJ, Zebin W, Qitang W (2011) Isolation and
characterization endophytic bacteria from hyperaccumulator Sedum alfredii Hance and their
potential to promote phytoextraction of zinc polluted soil. World J Microbiol Biotechnol 27:
1197±1207.
Zhang YF, He, LY, Che ZJ, Wang QY, Quian, Sheng, XF (2011) Characterization of ACC
deaminase-producing endophytic bacteria isolated from copper-tolerant plants and their
potential in promoting the growth and copper accumulation of Brassica napus. Chemosphere
87: 57±62.
Zhang YM, Li Y Jr, Chen WF, Wang, ET, Tian CF, Li QQ, Zhang YZ, Sui XH, Chen WX (2011)
Biodiversity and biogeography of rhizobia associated with soybean plants grown in the North
China Plain. Appl Environ Microbiol 77: 6331±6342.
Zhou J, Bruns MA, Tuedje J (1995) DNA recovery from soils of diverse composition. Appl Environ
Microbiol 62:316±322.
Notation
AIC
CDM
IAA
MBMM
PCR
PGP
TAE
TSA
TSI
Akaike information criterion
chemical definide medium
Indolacetic acid
MES buffered minimal medium
polymerase chain reaction
plant growth promoting
buffer solution containing Tris base, acetic acid and EDTA
tripticasein soy agar medium
tripticasein soy infusion
588
LIST OF COAUTHORS:
Brenda Román Ponce, Doctorate Candidate, Author to whom all
correspondence should be addressed
ENCB-IPN, Dept. of Microbiology, Plan de Ayala y Prolongación de Carpio s/n,
Col. Santo Tomás, México D.F., 04578, México.
Tel.: (5255) 57296300 ext. 62385
E-mail:[email protected]
Salvador Embarcadero Jiménez, Doctorate Candidate; presenting author.
Tel.: (5255) 57296300 ext. 62385
Dra. María Soledad Vásquez Murrieta, Full Professor
Tel.: (5255) 57296300 ext. 62382
Dr. En Tao Wang Hu,Full Professor.
Tel.: 57296300 ext. 62385
589
CHAPTER 6.4. ACUMULACIÓN DE Hg Y As POR Brassica rappa L. INDUCIDA
POR AGENTES QUELANTES Y TRATAMIENTOS NUTRICIONALES
CULTIVADA EN RESIDUOS MINEROS
Flores-de la Torre Juan Armando (1); Luna-Sandoval Rubí E. (1);
Rodríguez-Hernández Guillermo (2); Manzanares-Acuña Eduardo (3)
(1) Universidad Autónoma de Zacatecas, Unidad Académica de Ciencias Químicas, Carr. a
Guadalajara km 6, Ejido la Escondida, Zacatecas, Zacatecas, México, (492)9256690,
[email protected]
(2) Universidad Autónoma de Zacatecas, Unidad Académica de Ciencias Agronómicas, Zacatecas,
México.
(3) Universidad Autónoma de Zacatecas, Unidad Académica de Estudios Nucleares, Zacatecas,
México.
RESUMEN
La minería genera una gran cantidad de residuos ricos en metales pesados que
son depositados en la superficie del entorno minero, ocasionando que dichos
metales puedan incorporarse a la cadena trófica, además de provocar un impacto
ambiental. Entre las estrategias que se han empleado con éxito para inmovilizar o
extraer a los metales pesados en los ecosistemas se encuentra la fitoremediación,
donde además se puede ver favorecida por la acción de agentes quelantes que
puedan coordinar simultáneamente a un ión metálico y tratamientos calcáreos,
orgánicos y nutritivos en los suelos contaminados. Se obtuvieron jales mineros de
ODPLQD³(O%RWH´=DFDWHFDV1:HQXQPXHVWUHRGHWLSR
exploratorio basado en la Norma NMX-AA-132-SCFI-2006, logrando obtener
muestras representativas del área, las cuales se tamizaron, se les determinó el pH
y cuantificó mercurio y arsénico, por medio de espectroscopía de fluorescencia de
rayos X (EDXRF) para posteriormente realizar cultivos en dichas muestras, con 6
tratamientos de mezclas diferentes de soluciones nutritivas, así como con 4
agentes quelantes (Tiocianato de amonio, ácido cítrico, ácido Arscórbico y EDTA),
usados con el propósito de inducir la hiperacumulación de metales. Una vez
cosechadas las plantas se cuantificó arsénico (As) y mercurio (Hg) en las
diferentes secciones de la planta (hoja, tallo, raíz) y en las rizosferas donde fueron
cultivadas. Resultando concentraciones de As de hasta 2500 ppm y 430 ppm de
Hg en los suelos y valores de pH desde 7 hasta 7.83, lo que pudiese provocar un
comportimiento distinto en la bioquímica de la absorción de los metales por las
plantas. Referente a la aplicación de materia orgánica en combinación con
soluciones nutritivas incrementaron la biomasa y en combinación con agentes
quelantes como el EDTA y el ácido ascórbico, se triplicó la concentración de As en
la planta (100 ppm) y la concentración de Hg absorbida por la planta se
incrementó de 45 ppm a 70 ppm.
Palabras clave: Agentes quelantes, Brassica rappa L, fitoremediación, metales.
590
ABSTRACT
Mining generates a lot of rich waste in heavy metals are deposited on the surface
of the mining environment, causing these metals can enter the food chain as well
as causing an environmental impact. Among the strategies that have been
successfully used to immobilize or remove heavy metals phytoremediation
ecosystem, which also can be favored by the action of chelating agents that can
simultaneously coordinate to a metal ion and lime treatments, organic is and
nutritious in contaminated soils. Tailings of "The Boat", Zacatecas (22º46'35'' N,
102°26'21'' W), mine were obtained by sampling based on exploratory-AA-132SCFI Standard NMX-2006, achieving representative samples of the area, which
were sieved, the pH was determined and quantified mercury and arsenic, by
spectroscopy XRF (EDXRF), to subsequently be cultured in said samples, with 6
treatments of different mixtures of solutions nutritive and 4 chelating agents
(ammonium thiocyanate, citric acid, Arscórbico acid and EDTA), used for the
purpose of inducing metal hyperaccumulation. Once harvested plants arsenic (As)
and mercury (Hg) was quantified in the different sections of the plant (leaf, stem,
root) and in the rhizosphere where they were cultivated. Resulting concentrations
up to 2500 ppm As and 430 ppm of Hg in soils and pH values from 7 to 7.83, which
could result in a different biochemical comportimiento uptake of metals by plants.
Regarding the application of organic matter in combination with nutrient solutions
increased the biomass and in combination with chelating agents such as EDTA
and ascorbic acid, the concentration of As in the plant (100 ppm) and the
concentration of Hg absorbed tripled by plant increased from 45 ppm to 70 ppm.
Keywords: Brassica rapa L, Chelating agents, metals, phytoremediation
591
Introducción
Actualmente la contaminación por metales pesados esta siendo considerada uno
de los problemas ambientales, más serios principalmente en áreas que tienen una
gran actividad antrópica. (Llugany, et. al, 2007; Enciso, 1994)
En Zacatecas existen zonas mineras, en las cuales se encuentran
concentraciones elevadas de metales pesados que pueden ocasionar problemas
de salud (Flores, et al., 2014)
Los desechos producidos por las minas, desde los inicios de esta actividad
económica, contiene metales pesados que son depositados en la superficie del
entorno minero, (Flores et al., 2014). El suelo, entonces, experimenta un gran
impacto y una consecuencia directa de esta contaminación del suelo es una
ausencia inicial de vegetación o la pérdida de su productividad, la disminución de
la biodiversidad e indirectamente, la contaminación del aire y aguas superficiales y
subterráneas, (Becerril et al, 2007). Entre las estrategias que se han empleado
con éxito para extraer o inmovilizar a los metales pesados en los ecosistemas se
encuentra la fitoremediación. La fitoremediación de suelos contaminados se basa
en el uso conjunto de plantas, enmiendas al suelo y técnicas agronómicas para
eliminar, retener, o disminuir la toxicidad de los contaminantes del suelo (Chaney
et al., 2000).
La clave para el éxito de la aplicación de estas tecnologías se basa en un
fenómeno denominado hiperacumulación de metales. Desde hace algún tiempo,
con el propósito de híperacumular algunos metales pesados, se han utilizado
especies como Brassica juncea, Brassica nigra y Brassica rapa L., que han
mostrado capacidad para híperacumular Zn, Pb, As, Cd, Cu (Lasat et al., 2002)
El grupo de fitotecnologías usadas para descontaminar los suelos, reúne un
gran número de ventajas, especialmente la limpieza y la economía; no deben
utilizar reactivos químicos peligrosos, ni afectar negativamente a la estructura del
suelo, sólo aplicar prácticas agrícolas comunes; además, el proceso se realiza 'in
situ' evitando costosos transportes. Sin embargo, las condiciones adversas,
producto de la actividad minera, inhiben el crecimiento y desarrollo de especies
vegetales debido a la presencia de altas concentraciones de metales pesados que
resultan tóxicos para las plantas. Baker y Walter (1990), han propuesto que las
plantas tienen diferentes respuestas al estrés de metales pesados; ya que pueden
extraerlos, acumularlos o sólo indicar su presencia (Rascio et al., 2011). Cuando
las plantas crecen en sitios contaminados con metales pesados, el crecimiento de
raíces es fuertemente afectado (Cunningham et al., 1995)
Gracias a la adición de biofertilizantes y fertilizantes foliares la especie vegetal
tiene los requerimientos nutrimentales en el suelo, como consecuencia, su
crecimiento se acelera y en condiciones desfavorables (alto contenido de sales,
variación de pH) las plantas adquieren una mayor posibilidad de desarrollarse.
Los agentes quelantes forman compuestos de coordinación que son solubles
en agua y pueden ser asimilables por la vegetación, según Navarro et al. (2007).
La adición de nutrientes a los sustratos alojados en las presas de desechos de la
actividad minera, puede favorecer las condiciones de crecimiento de las especies
592
que a menudo son usadas con propósitos de fitoextracción de metales y otros
contaminantes (Anderson et al., 2005)
Algunas plantas y diferentes organismos desarrollan un complejo mecanismo
de homeostasis para minimizar los efectos dañinos de los metales pesados,
controlando la absorción, y la traslocación de los metales en los tejidos (Becerril,
et al., 2007).
Especies de plantas hiperacumuladoras de metales pesados como lo es la
Brassica rappa (L.), es una alternativa para la limpieza de los suelos contaminados
con metales, como la fitoextracción, la cual consiste en la absorción y acumulación
de metales pesados en la parte aérea de las plantas hiperacumuladoras
(Anderson, et al., 2005).
La materia orgánica es muy importante ya que tiene la capacidad de retener
agua, mejora la estructura del suelo, estimula el desarrollo de las plantas, su
acción quelante contribuye a disminuir los riesgos carenciales y favorece la
disponibilidad de algunos micronutrientes (Felix, et. al., 2008).
Material y métodos
Se seleccionó como área de muestreo, los jales que se encuentran ubicados en la
PLQD ³(O %RWH´ =DFDWHFDV XELFDGD D PHQRV GH NP GH OD ]RQD XUEDQD HQ OD
capital del Estado de Zacatecas, en las coordenadas 22º46´35´´ latitud N,
102º26´21´´ latitud W, realizandose un muestreo de tipo sistemático, basado en un
patrón geométrico específico donde las muestras son tomadas a intervalos
regulares a lo largo de ese patrón. Útil para cubrir en forma fácil y uniforme un
sitio, de forma que toda la zona está representada en las muestras, la primera
muestra se escogió aleatoriamente, y el resto de acuerdo al patrón asignado, con
este tipo de muestreo se asegura que la población de muestras este representada
en forma total y uniforme, es ideal para obtener volúmenes de material
contaminado (NMX-AA-132-SCFI-2006).
Las muestras fueron almacenadas en material de polietileno, se evitaron
factores como la luz, el calor, la humedad y otros factores que puedan cambiar la
naturaleza de la muestra, posteriormente las muestras se secaron en charolas de
aluminio esto se hace para facilitar el manejo de la muestra, mejorar la
homogeneización y disminuir los cambios químicos indeseables.
Procesamiento de suelos. Se realizó el tamizado de las muestras con dos
objetivos, uno liberarnos de rocas de tamaño mayor a 5 mm y el segundo es
obtener una muestra homogénea para su posible cuantificación por el método de
fluorescencia de Rayos X, donde se requieren 2 g de muestra con tamaño de
partícula menor a 200 um, calibrando el equipo con estándares NIST (National
Institute of Standards and Technology) de concentraciones de 1000 ppm, 500
ppm, 250 ppm, 125 ppm, 72.5 ppm, 36.2 ppm, 18.1 ppm y 9 ppm de los metales a
cuantificar mercurio (Hg) y arsénico (As).
Para el análisis de pH del suelo se utilizó un potenciómetro marca Orion,
preparando extractos 1:10 agua-suelo, como lo indica la Norma Oficial Mexicana
593
021±SEMARNAT-2000, posterior a la calibración del potenciómetro con
soluciones amortiguadores de la misma marca, con pH de 4, 7 y 10, lo cual es una
de las mediciones mas comunes e importantes de caracterización fisicoquímica de
suelos derivado que controla reacciones químicas y biológicas (Ramos et al.,
2012).
Cultivo de Brassica rappa L. Con el propósito de evaluar la influencia de
fertilizantes químicos, agentes calcáreos y agentes quelantes, se cultiva la planta
Brassica rappa L., bajo 6 tratamientos, 5 repeticiones y un testigo, tomando como
base del cultivo jal compuesto de todas las muestras extraídDV GH OD PLQD ³(O
%RWH´ =DFDWHFDV 8WLOL]DQGR IHUWLOL]DQWH TXtPLFR FX\D FRPSRVLFLyQ HV GH 1
8% P, 46% K, además de elementos tales como Fe, Mn, Mg, Cu, Ca y S en
cantidades traza, mezclando dicho fertilizante con una fuente de fitoreguladores
conocida como biol en una relación 1:100 (fertilizante-biol), como agente calcáreo
se empleó Ca(OH)2, los cuales se agregaron a los cultivos como parte de los
diferentes tratamientos.
Las plantas fueron cultivadas en condiciones de invernadero a una
temperatura promedio de 27ºC y humedad relativa del 75%, usando semilla Hvola
401, germinándolas previamente en charolas de germinación por un período de 30
días, las cuales contienen composta. Posteriormente se trasplantaron a
recipientes de polietileno con 1 kg de jal, aplicando riego hasta una capacidad de
campo. Una vez realizado el trasplante se comenzó con la aplicación del agente
calcáreo y fertilizante químico (tabla 1), aplicando riego 2 veces al día y 7
aplicaciones de fertilizante químico divididos en los 69 días del experimento.
TABLA 1. Base de los suelos de cultivo
Tratamiento
Contenido en maceta
1
Jal
2
Jal, fertilizante químico
3
Jal, Ca(OH)2
4
Jal, fertilizante químico, Ca(OH)2
5
Jal, materia orgánica
6
Jal, fertilizante químico, materia orgánica
Para la inducción de la capacidad de acumulación de los metales se añadieron
a los cultivos 2 veces con espacio de una semana, 100 mL de agentes quelantes
en las concentraciones que se observan en la tabla 2. Finalmente el diseño
experimental en conjunto de los tratamientos y los agentes quelantes empleados
se observa en la figura 1.
594
TABLA 2. Agentes quelantes utilizados para la inducción a la
hiperacumulación de metales
Agente químico
Concentración
quelante
Tiocianato de amonio 1 g/kg (Anderson et. al.,
(NH4SCN)
2005)
0.1 M (Robinson et al.,
Ácido Cítrico (C6H8O7)
1997)
Ácido Ascórbico (C6H8O6) 0.01 M
0.1 M (Robinson et al.,
EDTA Sal sódica
1997)
FIGURA 1. Distribución de los agentes quelantes en el diseño experimental
Posterior a la aplicación de los agentes quelantes, las plantas se secaron, por
lo que se cosecharon y además se tomó muestra de los suelos donde fueron
cultivadas, ambas muestras se molieron con el fin de obtener tamaños de partícula
homogéneos y así poder cuantificar por fluorescencia de Rayos X (MiniPal, Phillips
®). Una vez obtenidas las concentraciones de metales tanto en suelos como en
las plantas, la capacidad de acumulación de metales por las plantas se evalúa por
medio del factor de biodisponibilidad (Flores, et al., 2013).
/DV PXHVWUDV GH MDO REWHQLGDV GH OD PLQD ³(O %RWH´ =DFDWHFDV FXHQWDQ FRQ XQD
textura predominantemente franco arenosa (CSSS, 1993), con una fracción del
7.9% con tamaño de partícula menor a 74 um.
595
El pH es determinante para la disposición de los metales, la mayoría de ellos
tienden a estar disponibles a pH ácido, excepto As, Mo, Se y Cr, los cuales tiende
a estar más disponibles a pH alcalino. La absorción de los metales pesados está
fuertemente condicionada por el pH del suelo y por tanto, también la
biodisponibilidad de sus componentes (Alloway et al., 1993).
Los niveles de pH de las muestras obtenidas oscilan desde los 7.6 hasta los
3.0, al momento de realizar la muestra compuesta para el cultivo de la planta en
estudio el pH obtenido fue de 5.4 +/- 0.04, posterior al añadir el fertilizante químico
el pH no se modificó considerablemente llegando a 5.6, en el caso de los
tratamientos que se le agregó enmiendas calcáreas el valor de pH obtenido fue de
6.4 +/- 0.06. El efecto tanto de la materia orgánica como del fertilizante químico en
relación con la biomasa obtenida fue de 12 veces el tamaño y masa con respecto
a los tratamientos sin la presencia de dichas modificaciones.
La concentración de Hg y As contenidos en el material de cultivo de la planta
%UDVVLFD UDSD / SURYHQLHQWH GH ORV UHVLGXRV PLQHURV GH OD PLQD ³(O %RWH´
Zacatecas, resultaron de 185.24 +/- 3.02 ppm de Hg y 1125.49 +/- 12.42 ppm de
As., valores que superan la normatividad vigente (NOM-147-SEMARNAT/SSA-12004), donde establece que los niveles permitidos de Hg para suelo industrial es
máximo 310 ppm y para uso agrícola/residencial de 23 ppm, tomando como valor
permisible el de uso agrícola/residencial, resaltando que a escasos 200 m se
encuentran asentamientos humanos, para el caso de As, la normatividad marca
como niveles máximos permitidos en suelo industrial 260 ppm y para suelo de uso
agrícola/residencial de 22 ppm, ambos superados por los niveles encontrados.
Lo cual indica que el sitio muestreado es una zona de riesgo, tanto por los
niveles de As y Hg contenidos, como por la ubicación y el tamaña de partícula, ya
que es una zona de vientos elevados que pueden acarrear partículas de jales a
sitios donde hay una mayor densidad poblacional.
Una vez cuantificadas las concentraciones de Hg y As en las plantas obtenidas
después de los tratamientos y exposiciones a los agentes quelantes, se realizaron
los cálculos necesarios para obtener los factores de biodisponibilidad (FB) que
indican la capacidad de la planta de acumular algún elemento en particular, en
este caso los metales mencionados, realizando dicho cálculo relacionando las
concentraciones del metal en la planta con respecto a la concentración del mismo
metal contenido en el suelo donde se cultivo.
Para el caso de factores de biodisponibilidad de Hg de la planta Brassica rappa
L., se observa en la tabla 3, que la combinación del tratamiento 2 con el agente
quelante ácido ascórbico incrementa en mayor cantidad la acumulación de Hg,
seguido de la combinación del mismo agente quelante con los tratamientos 1 y 4,
para el caso del agente quelante tiocianto de amonio, éste incrementa la
capacidad de acumulación de Hg por la planta en combinación con el tratamiento
5 y 6, el EDTA incrementó la acumulación del metal en la planta únicamente en
combinación con el tratamiento 3 y en el experimento donde se utilizó ácido
ascórbico el incremento fue mínimo o incluso nulo con todos los tratamientos.
596
TABLA 3. FB de Hg de Brassica rappa L. por cada tratamiento
Agentes quelantes
Tratamientos EDTA
Control
Tiocianato
Ácido
Ácido
Cítrico
Ascórbico
1
0.266±0.039 0.236±0.026 0.166±0.022 0.284±0.017 0.190±0.025
2
0.339±0.136 0.265±0.043 0.293±0.044 0.399±0.083 0.279±0.041
3
0.227±0.029 0.203±0.018 0.155±0.010 0.153±0.021 0.224±0.025
4
0.255±0.029 0.256±0.037 0.213±0.054 0.312±0.067 0.223±0.027
5
0.206±0.048 0.309±0.063 0.187±0.041 0.271±0.051 0.251±0.033
6
0.209±0.027 0.433±0.034 0.301±0.044 0.351±0.049 0.311±0.027
La tabla 4 muestra los resultados de FB de Brassica rappa L., para el As,
donde se observa que el agente quelante que incremento mayoritariamente la
acumulación de As en la planta fue el tiocianato de amonio y el EDTA en
combinación con los tratamientos 2, 4 y 6 para el tiocianato y 1, 3 y 5, de igual
forma que con el Hg, el ácido cítrico, no incrementó la acumulación de As por la
planta, sin embargo para el As, tampoco el ácido cítrico demostró elevar
considerablemente la acumulación del metal por la planta.
TABLA 4. FB de As de Brassica rappa L. por cada tratamiento
Agentes quelantes
Tratamientos EDTA
Control
Tiocianato
Ácido
Ácido
Cítrico
Ascórbico
1
0.081±0.004 0.033±0.002 0.034±0.002 0.043±0.004 0.026±0.001
2
0.036±0.001 0.042±0.001 0.030±0.001 0.027±0.001 0.037±0.001
3
0.092±0.007 0.042±0.004 0.032±0.002 0.030±0.002 0.038±0.003
4
0.037±0.003 0.106±0.004 0.041±0.003 0.027±0.002 0.034±0.002
5
0.032±0.001 0.030±0.001 0.031±0.003 0.030±0.002 0.032±0.001
6
0.030±0.002 0.038±0.001 0.033±0.004 0.027±0.002 0.031±0.002
Anderson (2005) reporta que la especie vegetal Brassica juncea, usando como
agente quelante tiocianato de amonio, eleva la capacidad de acumular metales,
teniendo semejanzas con los resultados mostrados para el caso del As, y como lo
reporta Alloway (1993), cuando el pH es más ácido el As, se encuentra menos
biodisponible, observándose lo anterior con el uso de los dos agentes quelantes
ácidos (cítrico y ascórbico), no así para el Hg, ya que el mejor agente quelante fue
el ácido ascórbico y el mejor tratamiento el que además de jal, se le adicionó
únicamente fertilizante.
597
Conclusión
Se observa que la especie vegetal Brassica rappa L., al cultivarse con
tratamientos a base de fertilizantes químicos eleva su capacidad de acumulación
de Hg y As, sin embargo al adicionarle agentes quelantes para inducir dicha
acumulación, la planta en estudio responde favorablemente incrementado hasta
en un 25% su capacidad de retención de metales.
El agente quelante que elevó de forma mas significativa la acumulación de Hg
por la Brassica rappa L., fue el ácido ascórbico, seguido del tiocianato de amonio y
el EDTA, para el caso de la acumulación de As por la planta, funcionaron de forma
mas eficiente el tiocianato de amonio y el EDTA, descartando como inductores de
acumulación al ácido cítrico y ascórbico probablemente por la disminución del pH
que provoca su carácter ácido.
Por lo tanto la especie vegetal Brassica rappa L., se puede continuar
estudiando como un probable recurso fitotecnológico para la extracción de Hg y As
u otros metales pesados presentes en suelos contaminados, siempre y cuando se
empleen fertilizantes químicos, materia orgánica y agentes químicos capaces de
generar una mayor biodisponibilidad de los metales ya que según Pratas y
colaboradores (2013), para que una planta se considere hiperacumuladora, su
factor de biodisponibilidad debe ser igual o mayor a la unidad.
Referencias
Alloway B. J. and Ayres D. C. (1993) Chemical principles of Environmental Pollution, Blackie
Academic U.K., 140 ± 149.
Anderson C., Moreno F., and Meech J. (2005) A field demostration of gold phytoextration
technology, Minerals Engineering, 18(4), 385-392.
Becerril J. M., Barrutia O., García J. L., Hernández A., Olano J. M. y Garbisu C. (2007) Especies
nativas de suelos contaminados por metales: Aspectos ecofisiológicos y su uso en
fitoremediación, Rev Ecosistemas, 463 ± 484.
Canadian Society of Soil Science, Soil Sampling and Methods Analysis, Lewis Publishers, pp 823,
1993.
Chaney R., Malik M., LiM Y., Brown L., Brewer P., Scott Angle J. And Baker J. M. (2000)
Phytoremediation of Soils Metals, Phytoremediation State of the Science Conference U.S.
Environmental Protection Agency.
Cunningham S. D., Beri W. R. And Huang J. W. (1996) Promises and prospects of
phytoremediation, Biothecnology, 13, 1809 ± 1719.
Enciso de la Vega Salvador (1994), Crecimiento urbano de la ciudad de Zacatecas y sus
asentamientos humanos en zonas mineralizadas polimetálicas, pag 106-112.
Flores-De La Torre J.A., Medrano-Monreal A.Y., Manzanares-Acuña E., López-Luna M.A.,
Sánchez-Rodríguez S.H., Ramírez-Santoyo R.M., Vidales-Rodríguez L.E., López-Rodríguez
E.E. (2013). BIOACUMULACIÓN DE ARSÉNICO EN PLANTAS PRESENTES EN RESIDUOS
MINEROS DEL SEMIDESIERTO ZACATECANO. IX Congreso Nacional de Toxicología.
Nuevo Vallarta, Nayarit México. Revista Bio Ciencias 2(4) Supl 2: 46-47
Flores de la Torre J. A., Manzanares Acuña E., López Luna M. A., Sánchez Rodríguez S. H. And
Cuevas Flores M. R. (2014) Presencia de plomo en suelo y plantas en Concepción del Oro
Zacatecas por energía dispersiva de Fluorescencia de Rayos X, Memorias de Resúmenes 14º
Seminario de Investigación Universidad Autónoma de Aguascalientes, México, 162 ± 269.
Lasat MM. (2002) Phytoextration of toxic metals: a review of biological mechanisms, Journal of the
Environmental Quality, 31, 109-120
598
Navarro Aviño J.P., Aguilar Alonso I. Y López Maya J. R. (2007) Aspectos Bioquímicos y genéticos
de la tolerancia y acumulación de metales en plantas, Ecosistemas, 16 (2), 125 ± 132.
Pratas J., Favas P. J. C., Souza R., Varun M and Paul M. S. (2013) Phytoremedial assessment of
flora tolerant to heavy metals in the contaminated soils of an abandonaded Pb mine in Central
Portugal, Chemosphere, (90), 2216 ± 2225.
Ramos-Gómez M., Avelar-González F. J., Medel-Reyes A., Yamamoto-Flores L., Godinez L.,
Ramírez M., Guerra R y Rodríguez F. (2012) Movilidad de Metales en Jales Procedentes del
Distrito Minero de Guanajuato México, Rev Int. Contami. Ambie., 28, 49 ± 59.
Robinson B. H., Brooks R. R., Chambers M. F. and Nicks L (1998) Phytomining, Trends in Plant
Science, 3, 359-362.
Secretaría de Medio Ambiente y Recursos Naturales (SEMARNAT) / Secretaría de Salud y
Asistencia (SSA) (2007) Norma Oficial Mexicana NOM-147-SEMARNAT/SSA-1-2004, Diario
Oficial de la Federación, 2ª Ed., 1-69.
Secretaría de Comercio y Fomento Industrial (SECOFI) (2001) Norma Mexicana NMX-AA-132SCFI-2006 Norma Mexicana para el muestreo de suelos para la identificación y la
cuantificación de metales y metaloides y manejo de la muestra, Diario Oficial de la Federación.
599
Lista de co-autores.
Flores-de la Torre Juan Armando, Autor de correspondencia
Carretera Guadalajara km 6, Ejido la Escondida, Campus UAZ Siglo XXI, Edificio
1, (492) 92 5 66 90, [email protected]
Luna-Sandoval Rubí E.
1, (492) 92 5 66 90, [email protected]
Rodríguez-Hernández Guillermo
1, (492) 92 5 66 90, [email protected]
Manzanares-Acuña Eduardo
1, (492) 92 5 66 90, [email protected]
600
CHAPTER 6.5. EFFECT OF HYDROCARBONS ON GERMINATION AND
PHENOLIC PROFILE OF in vitro CULTURES OF Cyperus laxus
Morales Navarro Sarahí Itzel (1); Rivera Casado Noemí Araceli (2);
Salazar Plata Mónica Silvana (1); Gómez Guzmán Octavio (2);
Pérez Vargas Josefina (3); Calva Calva Graciano *(2)
(1) UVM, Ciencia de la Salud, México DF, México;
(2) CINVESTAV- IPN, Biotechnology and Bioengineering Department, México DF, México;
(3) TESE Div. IBQ, Edo. México, México.
Abstract
Cyperus laxus is an invasive plant species found in the tropical zone of the South
of Mexico, and is one of the pioneering species in weathered long-term oil spill
impacted sites of Tabasco. Previous studies on phytoremediation of soils of those
sites reported that this species can remove large amounts of petroleum
hydrocarbons (> 300,000 ppm), including polyaromatic (HPA), with a simultaneous
formation of conjugated plant phenolics-HPA compounds. However, the
mechanism used by this species for removal of such compounds, especially in the
absence of microorganisms, and the effect of oil on the germination and its plant
physiology is unknown. Thus, the aim of this study was to assess the effect of
anthracene (ANT), phenanthrene (FNN) and kerosene (K) over germination,
growth, physiological changes, and phenolic profile of this plant in in vitro culture
conditions. Development and physiological changes of plants was monitored
weekly for 35 days, when there was a clear removal of hydrocarbons from the
culture medium. The results demonstrate that only hydrocarbons from K at higher
concentrations than 3mM drastically affected the germination frequency; however,
after three weeks in the FNN treatments an ageotropic growth parallel to the
culture medium, with a clear increase of the thickness of leaves forming spirals or
curvatures of up to 180° was observed. The analysis of the phenolic profile in
leaves and bulb-roots, separated into an aqueous methanolic polar and a
chloroform organic fractions, showed that plants grown in the absence or presence
of hydrocarbons produced several phenylpropanoids and free flavonoids, however
in treatments with hydrocarbons also some derivatives of these compounds but
conjugated with the exogenous hydrocarbons were detected. In the organic phase,
free FNN and ANT were detected along with some of their oxidation metabolites,
mainly the trans-4,3-dihydrodiol-9-phenanthrene and the phenanthrylglucopyranoside. In conclusion the polyaromatic hydrocarbons did not significantly
affect the germination process, but it was inhibited by K at concentrations above 3
mM producing important physiological changes in the Cyperus plants. The
presence of phenolic-hydrocarbon conjugated compounds between the exogenous
hydrocarbons with the plant metabolites support the hypothesis that chemical
interactions between the plant metabolites with the HPA compounds may be
involved in the removal mechanism of HPA in phytoremediation systems.
Keywords: Cyperus laxus, germination, hydrocarbons, phytoremediation,
phenolics.
______________________
*Author for all correspondence: [email protected]
601
EFECTO DE HIDROCARBUROS SOBRE LA GERMINACIÓN Y PERFIL DE
COMPUESTOS FENOLICOS DE Cyperus laxus CRECIDAS in vitro
Resumen
Cyperus laxus es una especie vegetal invasiva que crece en la zona tropical del
sur de México y es una de las pioneras en sitios intemperizados del estado de
Tabasco después de haber sido impactados por derrames de petróleo crudo. En
estudios anteriores sobre fitorremediación de suelos de esos sitios se reportó que
esta especie puede remover grandes cantidades de hidrocarburos del petróleo
(>300,000 ppm), incluyendo poliaromáticos (HPA), acompañada de la formación
de compuestos fenólicos que se combinan con algunos hidrocarburos
poliaromáticos. Sin embargo se desconoce el efecto de estos compuestos sobre
su germinación y fisiología y cuales son los mecanismos que usa esta especie
para la remoción de ese tipo de compuestos. Así, el objetivo de este estudio fue
evaluar el efecto del antraceno (ANT), fenantreno (FNN) y queroseno (K) sobre la
germinación, crecimiento, cambios fisiológicos y los cambios en el perfil de
compuestos fenólicos de esta planta en condiciones in vitro. Se monitoreó el
desarrollo y cambios fisiológicos semanalmente por 35 días, a los cuales se
observó una clara remoción de los hidrocarburos a partir del medio de cultivo. Los
resultados demuestran que sólo el K a concentraciones mayores a 3mM afectó
drásticamente la frecuencia de geminación; sin embargo, a la tercera semana el
FNN produjo un crecimiento ageotrópico paralelo al medio de cultivo con un claro
incremento en el espesor de las hojas con formación de caireles o curvaturas de
hasta 180°. El análisis del perfil de compuestos fenólicos en las hojas y bulbo-raíz,
divididos en una fracción polar acuosa y otra orgánica, mostró que las plantas
crecidas en ausencia o presencia de hidrocarburos producen varios
fenilpropanoides y flavonoides libres, sin embargo en los tratamientos con
hidrocarburos se observaron además derivados de esos compuestos conjugados
con los hidrocarburos exógenos. En la fase orgánica, se detectó ANT y FNN libres
y algunos de sus metabolitos de oxidación, principalmente el trans-4,3-dihidrodiolfenantreno y el 9-fenantrilglucopiranósido. En conclusión los hidrocarburos
poliaromáticos no afectaron significativamente el proceso de germinación, pero
esta fue inhibida por el K a concentraciones arriba de 3 mM provocando cambios
fisiológicos importantes en las plantas. La presencia de fenoles conjugados con
los hidrocarburos aromáticos exógenos soporta la hipótesis de que la producción
de metabolitos capaces de conjugarse con xenobióticos puede ser uno de los
mecanismos de remoción en los sistemas de fitorremediación.
Palabras clave: Cyperus laxus, germinación, hidrocarburos, fitorremediación,
fenoles.
602
Introducción
La fitorremediación es un conjunto de procesos biotecnológicos a base de plantas
utilizada para la remoción, contención o transformación de contaminantes (Salt et
al., 1998). Se ha referido como una biotecnología verde que facilita la eliminación
de contaminantes a partir de aguas, suelos y sedimentos (Gunther et al., 2000).
Actualmente, se utiliza para el tratamiento de contaminantes como hidrocarburos
de petróleo, solventes clorados, pesticidas, explosivos y metales pesados. Esta
biotecnología ha permitido el diseño de estrategias tecnológicas para la
recuperación de sitios impactados por contaminantes como el petróleo crudo.
Dentro de los hidrocarburos del petróleo, los más utilizados como modelo para
estudios de fitorremediación están el fenantreno (FNN), antraceno (ANT) y
queroseno (K). El FNN se ha utilizado por sus características físico-químicas y
potencial carcinogénico (Guerrero-Zúñiga et al., 2009), mientras que el ANT se
utilizado por ser un isómero del FNN y el queroseno por ser una mezcla compleja
que incluye tanto aromáticos como alifáticos (Ritchie et al. 2003).
Aunque la fitorremediación se ha venido aplicando desde hace varios años, el
conocimiento sobre los mecanismos que utiliza la planta para la remoción de
hidrocarburos, especialmente en ausencia de microorganismos es escaso. Con
respecto a ese tema, en trabajos anteriores, nuestro grupo ha reportado que los
metabolitos secundarios producidos por plantas de Cyperus pueden actuar como
mediadores metabólicos en procesos de fitorremediación de suelos
intemperizados y con alto contenido de hidrocarburos después de haber sido
impactados por derrames de petróleo crudo (Rivera-Casado et al., 2008). Durante
esos estudios se encontró que puede haber interacciones químicas y bioquímica
entre los metabolitos y enzimas de la planta con la formación de complejos
metabolito-xenobiotico durante el proceso de remoción (Rivera-Casado et al.,
2011). En seguimiento a esos estudios en este trabajo se presentan los resultados
del cambio en la fisiología y perfil de compuestos fenólicos de plantas crecidas en
sistemas de fitorremediación in vitro para la remoción de hidrocarburos aromáticos
por Cyperus laxus. Cabe mencionar que Cyperus laxus es una especie vegetal
pionera de esos sitios intemperizados y que es capaz de crecer en suelos con más
de 325000 pm de hidrocarburos totales (Palma-Cruz et al., 2009).
Material vegetal. Se usaron semillas de plantas crecidas en invernadero sobre
suelo procedente de sitios intemperizados de Tabasco después de 10-50 años de
haber sido impactados crónicamente por derrames de petróleo crudo y
conteniendo hasta 350,000 ppm de hidrocarburos totales.
Desinfestación de semillas. Las semillas se embeben en agua por 2 horas, con
la finalidad de rehidratar y eliminar material flotante. La desinfección se realizó
dentro de una campana de flujo laminar y con material en condiciones de
esterilidad. Se realizaron tres lavados con agua destilada y desionizada estéril a
las semillas y se sumergieron en etanol al 70% por 20 segundos. Se repitió el
603
proceso de lavado y después se sumergió en hipoclorito al 5% por 20 minutos. Se
repiten los lavados y se reposan en agua desionizada estéril.
Geminación. Se vaciaron 20 mL de medio MS (Murashige y Skoog, 1962) en
placas petri. Para los tratamientos con hidrocarburos se adicionaron al medio aun
líquido la cantidad necesaria de soluciones stock para llegar a concentraciones
mM de 0, 0.01, 0.1, 0.7, 1, 3, 5 y 10, con los respectivos controles con sólo
solvente (0 mM) sin embargo solo fue para el tratamiento con FNN Y ANT y con K
las concentraciones fueron g/L. Sobre el medio gelificado se colocan las semillas y
se sella la caja para su cultivo en oscuridad por 7 días. Posteriormente se exponen
a la luz para que continúe el crecimiento de las plántulas generadas por la
germinación. En este tiempo se estima la frecuencia de germinación y se sigue
monitoreando este factor por 35 días.
Extracción y análisis de compuestos fenólicos. Se realizó a partir del método
reportado por (Martínez-Juárez et al., 2004) con las plántulas a la primera semana
de haber sido expuesta a la luz en todas sus concentraciones para los 3
hidrocarburos. Se separó raíz-bulbo y hoja (figura 1) y se determinó peso fresco
para posteriormente triturarlas por separado con N2. Se realizó la extracción con
cloroformo: metanol (1:1) por triplicado. La fase metanólica acuosa se evaporó con
N2 gaseoso y resuspendió en metanol-agua (1:1). La fase de cloroformo se
evaporo y se resuspendió en hexano-etanol (1:1). El análisis de los fenoles e cada
fracción se realizó por HPLC-UV utilizando el método reportado por MartínezJuárez et al., 2004.
FIGURA 1. Semillas germinadas de Cyperus laxus y sus órganos
fundamentales.
Resultados y Discusión
Efecto de la concentración de hidrocarburos sobre la germinación de
semillas. Se determinó el efecto fitotóxico de los hidrocarburos sobre germinación
de semillas de Cyperus laxus. El proceso de germinación se consideró completo
cuando la cubierta de la semilla se rompe completamente y emerge la radícula y
cotiledón (Amadi et al. 1993; Maila & Cloete, 2002). En ausencia de hidrocarburos
la germinación se observó después 7 días de incubación en oscuridad (figura 2).
604
FIGURA 2. Microfotografía (40X) de semillas de Cyperus laxus germinadas
después de 7 días de incubación en oscuridad.
El fenantreno y antraceno afectaron poco la frecuencia de germinación a las
concentraciones probadas, sin embargo con queroseno la germinación fue
inhibida a partir de 3mM (Figura 3). Esto puede deberse a que en presencia de
hidrocarburos los tejidos vegetales elevan su temperatura interna y se reduce o
bloquea la fotosíntesis en las partes verdes, por lo que se interrumpe la
transpiración y se cierran los estomas en todos los órganos (Pezeshki et al.,
2000). Estos resultados concuerdan con reportes que sugieren que debido a que
la mayoría de los xenobióticos son lipofílicos y pueden ser adsorbidos y/o
acumulados a niveles tóxicos, las especies vegetales llevan a cabo mecanismos
de detoxificación a través de dos procesos secuenciales: transformación química y
compartamentalización (Amadi et al., 1993). Al respecto, en ese estudio se
propone que la inhibición de la germinación está relacionada con las propiedades
hidrofóbicas de los hidrocarburos, lo que impide y reducen el intercambio de agua
y gases con el medio ambiente.
FIGURA 3. Efecto de la concentración de hidrocarburos sobre la
frecuencia de germinación de semillas de C. laxus. Los números representan
0, 0.01, 0.1, 0.7, 1, 3, 5 y 10 mM para ANT y g/L para K de hidrocarburo.
Efecto de la concentración de hidrocarburos sobre el crecimiento de C.
laxus. La influencia del tipo y concentración de FNN, ANT y K sobre el crecimiento
y fisiología de plantas se determinó evaluando los cambios en tamaño y
geotropismo de las partes aéreas y sumergidas por 35 días de cultivo in vitro. En
605
ausencia de hidrocarburos y después de dos semanas, las plantas mostraron la
orientación de crecimiento geotrópico común, es decir las partes aéreas eran
erguidas con orientación ageotrópica y las raíces sumergidas con orientación
geotrópica (figura 4A). Sin embargo, en presencia de hidrocarburos se observaron
cambios fisiológicos importantes, especialmente con FNN, donde las partes
aéreas cambiaron su geotropismo a un crecimiento ageotrópico paralelo al medio
de cultivo (Figura 14B). Además, a partir de 0.7 mM de FNN y después de 21 días
de cultivo, se detectó un claro incremento en el espesor de las hojas acompañado
de la formación de una singular curvatura acairelada de hasta 180°, en ocasiones
resultando en la presencia de hojas en forma de espirales (figura 4C). Debe
enfatizarse que este cambio fisiológico respecto al geotropismo y espesor de las
hojas se acentuó después de la tercera semana, pero el cambio en la dirección de
crecimiento geotrópico a un crecimiento paralelo al medio de cultivo se detectó
poco después de la germinación. Interesantemente, los tratamientos con ANT no
mostraron cambios apreciables en el espesor y forma de las hojas en comparación
con el control, pero su crecimiento también fue paralelo al medio de cultivo. En
contraste, el K no afectó ni la fisiología de las hojas ni la orientación en el
crecimiento, mostrando plantas con características similares a las del control.
Estos resultados con K pueden ser debidos a que esta mezcla contiene
principalmente hidrocarburos alifáticos (>80%) y sólo una pequeña cantidad de
aromáticos (<20%), por lo que el contenido de FNN u otros HPA en los
tratamientos de este experimento probablemente no fue suficiente para ejercer el
efecto sobre la fisiología de las plantas.
A
B
C
FIGURA 4. Cyperus laxus después de 14 días de cultivo en ausencia de
hidrocarburos (A) y en presencia de 0.7 mM de fenantreno (B, C). Obsérvese
el típico crecimiento geotrópico en A y la formación de hojas espirales en B
y C.
Perfil de compuestos fenólicos de plantas crecidas en hidrocarburos. El perfil
de compuestos fenólicos polares contenidos en la fase de metanol de hoja y
bulbo-raíz de plantas crecidas en ausencia de hidrocarburos mostró la presencia
de diversos fenilpropanoides y flavonoides libres (Figura 5A). Sin embargo en
presencia de 0.5 mM de FNN, además de estos compuestos en forma libre
también se observaron de forma conjugada, como el a y b señalados en la figura
5A y que por su tiempo de retención menor y conservación del espectro UV del
FNN (pico e) fueron identificados como conjugados con el FNN-dimetoxibenzoato
(a) y el FNN-p-hidroxibenzoato (b). Sin embargo, en la fase de cloroformo (Figura
5B) se observaron señales claras para de FNN libre (picos f), así como dos de sus
606
presuntos metabolitos (picos c, d), cuyo espectro UV corresponde al trans-4,3difenantrenol (pico c), y al 9-fenantril-glucopiranósido (pico d). Estos resultados
concuerdan con lo reportado en la literatura para la degradación de FNN por
hongos filamentosos (Sutherland et al., 1991). Además de los dos metabolitos
mencionados, se observó un tercero no identificado (pico e), el cual conserva el
perfil espectral del FNN y también migra a menor tiempo de retención, por lo cual
se predice que debe ser el FNN conjugado con algún otro compuesto.
A
FIGURA 5. Perfil de compuestos contenidos en la fase metanólica (220nm)
(A) y de cloroformo (250) (B) de extractos de hoja (H) y bulbo-raíz (R) de
plantas de Cyperus laxus crecidas en 0.7 mM de FNN.
Efecto de los hidrocarburos sobre la sobrevivencia de plantas de Cyperus
laxus y remoción de hidrocarburos. Al evaluar la sobrevivencia a través de la
presencia visual de clorofila se observó que a los 50 días las hojas de las plantas
de todos tratamientos se tornaron cloróticas hasta la pérdida total de su color
característico (Figura 6). Finalmente la remoción de FNN en el tratamiento con 0.7
mM, donde se detectó la presencia de sus metabolitos y conjugados, fue del 80%
mientras que para el tiramiento con 10 mM fue del 70%. Esta disminución en la
tasa de remoción puede ser debida a que a la concentración de 10 mM el FNN
afectó más severamente la germinación (Figura 3) y crecimiento de las plantas, y
por lo tanto su capacidad de degradación.
FIGURA 6. Hojas cloróticas después de 50 días.
Conclusiones:
La presencia de hidrocarburos en cultivos in vitro de Cyperus laxus afecto
severamente su fisiología y solo en queroseno su germinación. El perfil de
compuestos fenólicos fue dependiente del hidrocarburo y de su concentración,
similar a las observaciones previas de que la cantidad y tipo de hidrocarburo
afectan no solo el metabolismo de esta planta a nivel fisiológico, sino también la
biosíntesis y perfil de fenoles que pueden actuar como intermediarios metabólicos
607
en el proceso de fitorremoción. En la fase de metanol revelo la presencia de
fenoles conjugados con hidrocarburos, en comparación con la fase de cloroformo
donde se detectaron derivados de estos hidrocarburos tanto libres como
conjugados con fenilporpanoides y flavonoides. En la sobrevivencia de las plantas
después de los 35 días las hojas dependiendo de la cantidad de hidrocarburo
contenido todas se tornaron cloróticas.
Agradecimientos
El autor agradece al laboratorio de Ingeniería metabólica que forma parte del
departamento de biotecnología y bioingeniería de CINVESTAV por el apoyo
constante para realizar este trabajo.
608
Referencias
Amadi A, Dickson A.A y Mate G.O (1993) Remediation of soil:effect of organic and inorganic
nutrient suplemento in the performance of maize (zea mays L). Water Air Soil Poll 66,5976.
Guerrero-Zuñiga L.A, Rodriguez-Dorantes A.M. 2009. Efecto de la presencia de fenantreno sobre
la expresión de proteínas y la actividad enzimática de Cyperus hermaphroditus.
Polibotanica . 27:103-130.
Gunther, T., Kirsche, B., Fritsche, W., 2000. Potential of plant±microbe-interactions for in situ
bioremediation of hydrocarbon- contaminated soils. In: Wise, D.L., Trantolo, D.J., Cichon,
E.J., Inyang, H.I., Stottmeister, U. (Eds.), Bioremediation of Contaminated Soils. Marcel
Dekker, USA, pp. 285±293.
Maila M.P., Cloete T.E. 2002. Germination of Lepidium sativum as a method to evaluate polycyclic
hydrocarbons (PAHs) removal from contaminated soil. International Biodeterioration and
Biodegradation 50:107-113
Martínez-Juárez, V. M., N. Ochoa-Alejo, E. Lozoya-Gloria, M. L. Villarreal-Ortega, A. Ariza-Castolo,
F. J. Esparza-Garcia,G. Calva-Calva. 2004:Specific synthesis of 5,5'-dicapsaicin by cell
suspension cultures of capsicum annuum var. annuum (chili Jalapeno chigol) and their
soluble and NaCl-extracted cell wall protein fractions. J Agric Food Chem. 52(4), p. 972-9
Palma-Cruz, F. (2009). Ecología y fenología de especies vegetales encontradas en suelos de
Tabasco crónicamente contaminados con hidrocarburos. Departamento de Biotecnología
y Bioingeniería. México D.F., Centro de Investigación y de Estudios Avanzados del
Instituto Politécnico Nacional (CINVESTAV) Unidad Zacatenco. Doctor en Ciencias.
Pezeshki, S.R., Hester, M. W., Lin, Q. and Nyman, J.A. (2000). The effects of oil spill and clean up
on dominant US Gulf coast marsh macrophytes: A review. Environ. Pollut., 108: 129-129.
Ritchie G.D., Still R.K., Ross J., Beckkedal V.Y.M., Bobb J.A., Arfsten P.D. 2003. Biological and
health effects of exposure to kerosene-based jet fuels and performance additives. J.
Toxicology and Environ. Health. Part B. 6:357-451.
Rivera Casado, N.A., Montes-Horcasitas, M.C., Esparza-Garcia, F.J., Ariza-Castolo, A., Gómez
Guzmán, O., Pérez-Vargas, J., Calva-Calva, G., fitotratamiento de suelos impactados por
derrames de petróleo: interacción entre hidrocarburos poliaromáticos, fenoles y enzimas
oxidativas. CENIC. Ciencias Químicas, Magazine Article. p.1-11.
Rivera Casado N.A., Rodríguez-Vázquez R., Montes-Horcasitas M.C., Pérez-Várgas J., GómezGuzmán O., Calva-Calva G., (2008) Hidrocarburos Aromaticos y Fenilpropanoides
Presentes en la Rizosfera de Plantas de Cyperus laxus Crecido en Suelos Contaminados
con Hidrocarburos. Tecnológico de Estudios Superiores de Ecatepec. Tecnocultura 20: p.
4-15
Salt, D.E., Smith, R.D., Raskin, I., 1998. Phytoremediation. Ann. Rev. Plant Physiol. Plant Mol. Biol.
49, 643±668.
Sutherland, J.B., Selby, A.L., Freeman, Evans, and Cerniglia, C.E (1991). Metabolism of
Phenanthrene by Phanerochaete chrysosporium. Appl. Environ. Microbiol. 57, 33 10-3316
609
CHAPTER 6.6 BIOLIXIVIACION DE METALES DE SEDIMENTOS
ANAEROBICOS DEL RIO RECONQUISTA (ARGENTINA) COMO ESTRATEGIA
POTENCIAL DE REMEDIACIÓN
Natalia Porzionato* (1); Roberto Candal (1, 2); Gustavo Curutchet (1, 2)
(1) Instituto de Investigación e Ingeniería Ambiental, Universidad Nacional de San Martín, Campus
Miguelete 25 de Mayo y Francia, Provincia de Buenos Aires, Argentina;
(2) Consejo Nacional de Investigaciones Científicas y Técnicas, Rivadavia 1917, Buenos Aires,
Argentina.
RESUMEN
Los sedimentos anaeróbicos de cursos de agua sujetos a contaminación de origen
industrial, acumulan contaminantes persistentes, fundamentalmente metales
pesados. En condiciones oxidantes, generadas por dragado o por escorrentías de
lluvias, el reservorio de metales acumulado puede solubilizarse en la columna de
agua con el consiguiente impacto negativo al ecosistema. Los procesos
biocatalizados de oxidación y reducción de compuestos del azufre en el sedimento
son las reacciones clave, al estar involucrados en los procesos de movilización e
inmovilización de metales pesados. Son estos mismos bioprocesos, los que en
condiciones controladas podrían ser empleados para la remediación de los
sedimentos contaminados.
El objetivo de este trabajo fue estudiar la aplicación de lixiviación en biopilas
(bioheap leaching) como estrategia de remediación de sedimentos contaminados
de la cuenca del Río Reconquista y evaluar el efecto del agregado de distintas
cantidades de azufre elemental sobre la movilidad de metales desde la biopila
hacia la solución.
La potencialidad de remediación por biolixiviación fue demostrada a través de
ensayos en biopilas conteniendo 100 gr de sedimento con 340 mg Kg -1 de zinc y
116 mg Kg-1 de cobre. Los sistemas se confeccionaron en columnas de 6 cm de
alto y 21 cm de diámetro, acondicionando con perlita para mejorar del drenaje y
agregando concentraciones crecientes de azufre elemental (1%, 2%, 5% p/p) para
incrementar el nivel de acidificación. Luego de tres meses de acidificación
generada por el riego periódico, se alcanzó una extracción del 70% del Zn inicial y
43% del Cu inicial en el sistema con 5% p/p de azufre dispuesto de manera
superficial sobre la biopila. Además se lograron enriquecer, aislar e identificar las
especies bacterianas responsables de la catálisis de los procesos de movilización
e inmovilización de los metales en el sedimento.
Se concluye que, dada la factibilidad que un drenaje ácido ocurra, generando
un aumento en la disponibilidad y peligrosidad de metales en condiciones
oxidantes, es recomendable no manipular indiscriminadamente los sedimentos de
cursos de agua contaminados como el Río Reconquista, siendo posible llevar a
cabo la remediación de estos por biolixiviación en biopilas de manera simple y
económicamente viable.
Palabras Clave: cuenca del río Reconquista, estrategia de remediación, metales
pesados, lixiviación en biopilas, sedimentos anaeróbicos contaminados.
610
ABSTRACT
Anaerobic sediments of watercourses which are subject to industrial pollution have
a tendency to accumulate contaminants, mainly heavy metals. In oxidizing
conditions, generated by dredging or rainfall, the accumulated metals mobilize to
the water column with a negative impact on the ecosystem. The biocatalyzed
oxidation and reduction of sulphur compounds in the sediment are the key
reactions that lead to this solubilization. These bioprocesses of mobilization and
inmobilization of metals are the same as those that could be used for remediation
of heavily contaminated sediments.
The aim of this work was to study the application of bioheaps leaching as a
strategy for the remediation of anaerobic contaminated sediment of the
Reconquista River basin, and to evaluate the effect of adding different
concentrations of elementary sulphur on the mobility of metals from the bioheap to
the aqueous solution.
The potential for remediation by bioleaching was demonstrated through
bioheap tests of containing 100 g of anaerobic sediment with 340 mg Kg -1of zinc
and 116 mg Kg-1 of cooper. The system was compiled into columns of 6 cm high
and 21 cm diameter, using perlita to improve drainage, and adding increasing
concentrations of elementary sulphur (1%, 2%, 5% w/w) to achieve an increased
level of acidification. After 3 months of acidification generated by periodic watering,
extraction of 70% of the initial Zn and 43% of the initial copper was achieved with
the addition on the surface of the bioheap of 5% w/w of elementary sulphur.
Furthermore, we succeeded in enriching, isolating and identifying the bacterial
strains responsible for the mobilization-immobilization processes of metals from
sediment.
We conclude that, given the feasibility of acid drainage, which the
consequential increasing of bioavailable and dangerous metals in oxidizing
conditions, it should not be indiscriminately manipulate the sediments from polluted
watercourses as Reconquista River, being possible to carry out its remediation by
a simple and economically viable manner using bioheaps leaching.
Keywords: anaerobic contaminated sediments, bioheap leaching, heavy metals,
remediation strategy, Reconquista River basin.
611
Introducción
El canal José León Suarez es un afluente del río Reconquista. Si bien nace de la
confluencia de varios colectores pluviales, arrastra desde su salida a cielo abierto
altos tenores de contaminación tanto cloacal como industrial, producto de
descargas irregulares (Curutchet et al., 2012). En contraste, en estos mismos
sistemas dónde el nivel de contaminación es elevado, se ha demostrado que la
capacidad de autodepuración del agua es alta, llegando al río Reconquista,
ubicado 5 Km aguas abajo, parámetros aceptables de carga orgánica (menor a 50
mg/l DQO) (Curutchet et al., 2012). Dicha autodepuración supera la velocidad
esperable de estabilización de la materia orgánica por oxidación en un curso de
agua de estas características, por lo que se atribuye la mayor proporción del
proceso a mecanismos de sedimentación e incorporación de contaminantes en los
sedimentos. Debido a esto se observa una gran incorporación de materia orgánica
a los sedimentos, generando una alta demanda de oxígeno, y estableciéndose en
consecuencia un ambiente anaeróbico, adecuado para la formación biocatalizada
de sulfuros insolubles (Porzionato et al., 2013, 2014). En estas condiciones de
bajo potencial redox, la mayor parte de los metales pesados precipitan en forma
de sulfuros e hidróxidos de baja solubilidad y se asocian a los distintos
componentes minerales del sedimento (Eggleton y Thomas, 2004),
manteniéndose en un estado menos bioaccesible mientras las condiciones redox
no se vean alteradas. Cuando ocurre un cambio en las condiciones por exposición
al oxígeno, sea por dragado o desecación (condición producida por disminución
del nivel de agua), los componentes reducidos como los sulfuros tienden a
oxidarse a través de procesos biocatalizados por bacterias azufre oxidantes
(Rawlings, 2002; Sand et al., 2001), provocando acidificación y liberación de
metales a la columna de agua e incrementando de esta manera su
biodisponibilidad (Porzionato et al., 2014; Di Nanno et al., 2007, Lors et al., 2004).
Aunque este proceso es análogo al observado en los drenajes ácidos de minas,
en el caso de los sedimentos, la diferente proporción y reactividad de los sulfuros y
sus características especiales de difusión de oxígeno, hacen necesarios diferentes
modelos y experimentos predictivos (Förstner, 2004; Seidel et al., 2006; Di Nanno
et al., 2007).
La liberación de metales que acompaña a la oxidación, se encuentra ligada a
generalmente a un proceso de acidificación, dependiendo de las características
del sedimento. Dichos procesos fueron estudiados de manera intensiva
(Porzionato et al., 2013, 2014; Ye et al. 2013; Di Nanno et al., 2007, 2009) por
medio de experimentos estáticos (Kersten y Förstner, 1991) y en sistemas de
oxidación de sedimentos anaeróbicos en frascos agitados, con monitoreo de pH,
recuento de microorganismos azufre oxidantes, análisis de sulfatos y metales en
solución.
En este trabajo se investigó sobre la hipótesis de que el mismo proceso de
oxidación/acidificación que en condiciones no controladas produce un serio
impacto ambiental, posee a su vez, en condiciones controladas, la potencialidad
de permitir la recuperación de importantes concentraciones de metales valiosos a
partir de pasivos ambientales. Si bien el uso potencial de la biolixiviación como
612
herramienta para remediación de sedimentos contaminados con metales pesados
ya ha comenzado a ser estudiado tanto en sistemas con resuspensión como en
biopilas (Löser et al., 2006, 2007; Seidel et al., 2006a, 2006b; Guven y Akinci,
2013; Ye et al., 2013), las características complejas y diversas entre sedimentos
de diferente procedencia, remarcan la importancia de estudiar múltiples sitios con
el fin de converger en modelos de uso general. En particular el uso de sistemas de
lixiviación en pila, se manifiesta como el más factible dado el gran volumen de
sedimentos a tratar. En este trabajo se estudió fundamentalmente como se ve
afectada la movilidad de los metales desde el sedimento hacia la solución acuosa
ante el agregado de distintas concentraciones de azufre elemental a las biopilas
(sistemas de sedimento acondicionados con perlita para mejorar el drenaje) y la
potencialidad de la aplicación de la biolixiviación en estos sistemas como
estrategia de remediación de sedimentos contaminados.
Las reacciones involucrados en los procesos de biolixiviación pueden ocurrir a
través de diferentes mecanismos: ácido, oxidante o reductor, basados en la
oxidación biocatalizada de diferentes fases con contenido de azufre. En estos
procesos intervienen como agentes de reacción los productos del metabolismo de
bacterias acidófilas como Acidithiobacillus ferrooxidans y thiooxidans. Los
mecanismos específicos de oxidación de sulfuros y azufre por estas especies
fueron extensamente estudiados previamente (Sand et al., 2001; Rawlings, 2002).
En el mecanismo ácido, los microorganismos catalizan la oxidación del azufre
elemental a H2SO4, el cual incrementa la acidez del medio, solubilizando a los
metales asociados a las distintas fracciones del sedimento y fundamentalmente a
aquellos asociados a la fracción intercambiable, estabilizándolos en solución. En el
mecanismo oxidante, los microorganismos catalizan la oxidación del Fe(II) soluble
con oxígeno como aceptor final de electrones. El Fe(III) formado oxida a los
sulfuros más insolubles reconvirtiéndose a Fe(II). Algunos sulfuros mas solubles
pueden oxidarse mediante un mecanismo independiente del Fe(III), mediado por
formación de polisulfuros (Sand et al., 2001). Este mecanismo se asocia
principalmente a la fracción oxidable, asociada a sulfuros y materia orgánica. Por
último el mecanismo reductor es usual para compuestos en alto estado de
oxidación (MnO2; Fe(OH)3). La oxidación parcial del azufre forma politionatos que
favorecen la reducción y disolución de estos óxidos, incrementando la solubilidad
de metales asociados a los mismos (Sand et al., 2001).
El objetivo principal de este trabajo es el estudio de la potencialidad de estos
procesos para desarrollar una estrategia simple y económica dando lugar a una
solución viable para la remediación de sedimentos contaminados y, en particular,
de sedimentos de la cuenca del río Reconquista, con gran interés local. Además,
es de utilidad incrementar el interés en la revalorización de pasivos ambientales
como los sedimentos contaminados al resaltar sus potencialidad como fuente
secundaria de metales valiosos, su rol en procesos de depuración de aguas
contaminadas in situ y la presencia en ellos de microorganismos con
características de interés por considerarse verdaderos extremófilos y por su gran
eficiencia en procesos tales como lixiviación y remediación.
613
Sitio en estudio y toma de muestra. El sitio se ubica inmediatamente aguas
abajo del Barrio La Cárcova en José León Suarez, Partido de San Martín,
SURYLQFLD GH %XHQRV $LUHV $UJHQWLQD
6
´: )LJXUD (O
mismo se encuentra constantemente percudido por las contribuciones de
desechos sólidos urbanos, descargas provenientes de desagües domiciliarios y
cloacales, además de pasivos industriales clandestinos. Se colectaron las
muestras compuestas hasta profundidad de 30 cm que fueron almacenadas en
frascos de polietileno a 4°C. Se conservó la humedad inicial de las muestras con
el fin de conservar las condiciones redox los más inalteradas posible.
FIGURA 1: Ampliación de sitio de muestreo. Imágenes obtenidas desde
Google maps.
Caracterización inicial. Se determinó humedad relativa por gravimetría y el
contenido de materia orgánica oxidable por el método de Walkley Black (Gelman
et al., 2012). Para determinar el contenido de sulfuros solubles en ácido tipo MS
(M denota metal bivalente como Fe(II) o Zn(II)) se realizó el ensayo de purga y
trampa para sulfuros volátiles en ácido, seguido por conversión a azul de metileno
(Morse et al.,1987; Tabatabai, 1982). Previamente se realizó una determinación
estimativa de la cantidad de carbonatos en la muestra por medio de Kemess Fiss
Rating (Di Nanno, 2003). Se realizó la extracción secuencial de metales estándar
recomendada por la Comunidad Boreau de Referencia, usualmente conocida
como extracción BCR (Ure et al., 1993) para determinar la concentración y
distribución de los metales en los distintos componentes del sedimento (Fracción
1: asociada a carbonatos y silicatos; Fracción 2: asociada a óxidos de Mn y Fe;
Fracción 3: asociada a sulfuros y materia orgánica; Fracción 4: Residual). Se llevó
a cabo la determinación de la concentración de Cu, Zn y Fe por Espectrometría de
Absorción Atómica (EAA).
Además, se determinó el riesgo de acidificación y liberación de metales in situ
siguiendo la metodología propuesta por Kersten y Förstner (K&F; Kersten y
Förstner, 1991).
614
Aislamiento de flora bacteriana nativa. Por medio de sucesivos cultivos de
enriquecimiento del sedimento en medio Postgate C (Postgate, 1984) y medio 0K
(Silverman y Lundgrem, 1959) con agregado de Fe(II) o azufre como fuente de
electrones, se aislaron, caracterizaron e identificaron bacterias con actividad
sulfato reductora, hierro reductora, azufre y hierro oxidante. La identificación
molecular de las cepas aisladas fue realizada a través de secuenciación genómica
del gen RNA 16S por la empresa Macrogen, Korea.
Lixiviación en frascos agitados. Se llevaron a cabo experimentos de agitación
de sedimento en medio 0K (Silverman y Lundgrem, 1959) con concentraciones de
5, 10 y 15% (masa/100 ml) de densidad de pulpa (DP) con y sin agregado de
Azufre (5% de la DP correspondiente). Todos los sistemas fueron inoculados con
mezcla de Acidithiobacillus ferrooxidans (DSM 11477) y Acidithiobacillus
thiooxidans (DSM 11478) de aproximadamente 1,2.109 bacterias ml-1 y
acondicionados inicialmente a pH 5 con gotas de H2SO4 1:10. Se realizaron
ensayos por duplicado. Se monitoreó periódicamente pH, concentración de H +,
concentración de Zn y sulfatos en solución y recuento de bacterias en suspensión.
Ensayo de lixiviación en biopilas (bioheaps). Los sistemas se confeccionaron
en lisímetros cilíndricos de PET, en columnas de 21 cm de diámetro y 6 cm de
altura, conteniendo cada uno 91 g (peso seco) de sedimento anaeróbico. El
sedimento fue acondicionado con 5% p/p de perlita para mejorar del drenaje y
aireación, y se adicionó azufre elemental (1%, 2% 5% p/p) en la mezcla. Uno de
los sistemas fue espolvoreado superficialmente con 5% de azufre en lugar de
mezclarlo en toda la masa. Todos los sistemas sedimento-perlita fueron saturados
con una suspensión de mezcla de cultivos en medio 0K de Acidithiobacillus
ferrooxidans (DSM 11477) y Acidithiobacillus thiooxidans (DSM 11478), con una
población total aproximada de 2,4.108 bacterias ml-1.
El régimen de riego fue desarrollado en dos etapas (Seidel et al., 2004). En la
primera etapa o etapa de acidificación, con una duración de 131 días, se buscó
promover fundamentalmente la acidificación generada por la oxidación de sulfuros
y azufre elemental, mientras que en la segunda etapa o etapa de lavado, con
duración de 2 días, se logró arrastrar todo el metal soluble liberado en el
microentorno. En la etapa de acidificación se regaron los sistemas con una
frecuencia de 3 días con 50 ml de agua destilada y se tomaron periódicamente
alícuotas del agua de riego para monitorear pH y contenido de Zinc, Cobre y
Hierro. Se permitió la recirculación del agua de riego hasta su posterior
recolección y almacenamiento. A continuación, en la etapa de lavado, se procedió
a regar los sistemas con 500 ml de agua destilada. Se analizó por EAA el
contenido de Zn, Cu y Fe del agua de lavado. Finalmente, se realizó la extracción
secuencial de metales (Extracción estándar BCR, Ure et al., 1993) sobre el
sedimento ya tratado con el fin de analizar la distribución final de Zn, Cu y Fe que
se encontrara remanente en sistema.
615
Características del sedimento. Los sedimentos estudiados presentaron las
características propias de un ambiente contaminado: altas concentraciones de
materia orgánica oxidable (12%) y sulfuros volátiles (685 mg kg-1). Estas
condiciones son las predominantes en los sistemas anaeróbicos, dado que la falta
de oxigeno disponible y el exceso de un sustrato orgánico favorece el desarrollo
de microorganismos anaeróbicos degradadores de materia orgánica y reductores
de sulfatos. Se detectaron concentraciones de Zn de 340 mg Kg -1 y Cu de 116 mg
Kg-1.
Aislamiento de flora bacteriana indígena. De las identificaciones moleculares se
obtuvo que las cepas aisladas del sedimento con actividad sulfato y hierro
reductora correspondían ambas a proteobacterias, del género Desulfovibrio y del
género Citrobacter respectivamente. Se aislaron además del mismo sedimento,
bacterias acidófilas con actividad azufre oxidantes. La caracterización bioquímica y
molecular confirmó que corresponden a Acidithiobacillus ferrooxidans y
Acidithiobacillus thiooxidans. La presencia de acidófilos en un sedimento de
características neutro-alcalinas, puede parecer sorprendente en una primera
lectura, aunque ya ha sido mencionada en otros trabajos (Lors et al., 2004;
Carnicero et al., 2009) y sugiere la presencia de microentornos con características
muy diferentes a las condiciones promedio del sistema.
Lixiviación en frascos agitados. En la figura 2 se muestran el porcentaje de
extracción de zinc y el pH a lo largo del experimento. En los sistemas con azufre
agregado, se observa una clara relación entre el zinc extraído y el descenso de
pH. En los sistemas con menor densidad de pulpa de sedimento, la acidificación
producida por la oxidación de azufre biocatalizada por las bacterias azufreoxidantes es más rápida debido a la menor cantidad de compuestos neutralizantes
presentes aportados por el sedimento. Estos sistemas llegan a una extracción
cercana al de 60-80 % en 40 días alcanzándose valores de pH entre 2 y 4. Los
sistemas sin azufre agregado muestran una extracción de Zn final de 40% para
5% DP y alrededor de 20% para los sistemas de 10 y 15% DP. Se alcanzan
acidificaciones de más de una unidad de pH. La extracción de Zn sin embargo no
sigue una cinética paralela al proceso de acidificación sino que sugiere además
participación de mecanismos de lixiviación por contacto (Sand, 2001; Rawlings,
2002).
616
a
b
FIGURA 2: Ensayo de resuspensión del sedimento en modo batch,
porcentaje (%) de Zn extraído en a) y valores de pH alcanzados en b) para los
sistemas con 5, 10 y 15 % de densidad de pulpa (DP).
Estos resultados muestran la potencialidad del proceso de biolixiviación para la
extracción de metales de este sedimento contaminado. Sin embargo, dados los
enormes volúmenes de sedimento a tratar, realizar un proceso de lixiviación en
este tipo de sistemas agitados puede ser económicamente no viable, de manera
que se procedió a estudiar el sistema de lixiviación estática en biopilas.
Ensayo de lixiviación en biopilas (bioheaps). En la figura 3 se muestran los
porcentajes de extracción de Cu y Zn, y el pH final alcanzado en las diferentes
biopilas.
617
FIGURA 3: Porcentajes de Zn y Cu total extraídos por lixiviación del
sedimento en biopilas, junto con los valores de pH final de la solución, para
cada sistema estudiado.
El pH final alcanzado para los sistemas sin azufre o con 1% de azufre se
mantuvo cercano a la neutralidad, mientras que para los sistemas con 2%, 5% de
azufre y 5% de azufre superficial el valor de pH final alcanzado en el sedimento
fue de 4.4, 2.9 y 2.3, respectivamente.
Los valores de porcentajes de Zn y Cu extraídos muestran grandes diferencias
entre los diferentes sistemas y en relación con el pH final alcanzado. En los
sistemas sin agregado de azufre y con 1% de azufre no se obtuvo una lixiviación
neta. Esto se encuentra asociado a que las condiciones de pH se mantuvieron
cercanas a neutralidad hasta el final del ensayo.
En los sistemas con agregado de 2%, 5% y 5% superficial de azufre se obtuvo
una significativa lixiviación de Zn, alcanzando a un 71% de eficiencia para el caso
del sistema con azufre superficial y entre 36-53% para los sistemas con azufre
5%. Mientras tanto, se detectó una lixiviación significativa de Cu solo para los
sistemas con 5% y 5% superficial. El sistema con 5 % de azufre superficial mostró
una extracción superior al sistema con 5 % de azufre mezclado. La mejor
disponibilidad de oxígeno en este sistema aparece como la principal causa de
estas diferencias. El medio ácido-reductor generado por metabolismo de los
Acidithiobacillus sobre azufre (Viera et al., 2003; Quintana et al., 2001) es capaz
de lixiviar eficientemente cobre y zinc aún de la fracción asociada a óxidos de
hierro y manganeso, probablemente por reducción de los mismos.
Se muestra en la figura 4 la especiación (según extracción secuencial)
obtenida para Zn y Cu remanente en sedimento.
618
FIGURA 4: Distribución de Zn y Cu remanente en las fracciones asociadas a
los distintos componentes del sedimento para los Sistemas Inicial y Final
luego del tratamiento: Sin Azufre, con Azufre 1%, 2%, 5% y 5% Superficial.
En todos los sistemas, aún en aquellos sin azufre se encuentra una
disminución drástica de la fracción de cobre y zinc asociada a la fracción oxidable.
Este resultado es esperable dado que se produce la oxidación de las fases sulfuro
y materia orgánica durante el proceso. El enriquecimiento de la proporción de
ambos metales en la fracción intercambiable luego del proceso de lixiviación
podría sugerir que se subestimó en el proceso la cantidad de agua de lavado
utilizada. Sin embargo, el uso de una cantidad mayor de agua de lavado produciría
una disminución de la concentración de ambos metales en ellas con la
consecuente dificultad en el procesamiento downstream para la recuperación de
los metales.
Conclusiones
Los resultados obtenidos demuestran que el proceso de biolixiviación de metales a
partir de sedimentos contaminados posee gran potencialidad para ser aplicados a
la remediación de los mismos, con recuperación asociada de metales valiosos.
En las experiencias preliminares en frascos agitados se encontró que es factible la
FRQWULEXFLyQ VREUH HO SURFHVR GH OL[LYLDFLyQ GH WDQWR PHFDQLVPRV ³GLUHFWRV´ D
partir de compuestos sulfurados presentes en el sedimento, como mecanismos
ácidos a partir de agregados de azufre elemental. El agregado de azufre presenta
ventajas en porcentaje y velocidad de extracción.
En los sistemas de lixiviación en pila, los sistemas con azufre superficial son
los que mejor performance muestran, posiblemente por tener una mejor
transferencia de oxígeno y favorecer el crecimiento de Acidithiobacillus sobre la
superficie de las partículas, mientras que en los casos con azufre integrado en la
mezcla, el mecanismo de lixiviación tuvo una eficiencia menor.
Dado que la infraestructura requerida para la lixiviación en pila no es muy diferente
a la necesaria para un relleno de seguridad (tal es el destino previsto en la
actualidad para sedimentos altamente contaminados en Argentina), utilizar este
619
proceso para la remediación de los sedimentos con el valor agregado de recuperar
metales valiosos, se muestra como una tecnología potencial de sumo interés.
Agradecimientos
Los autores agradecen al Consejo Nacional de Investigaciones Científicas y
Técnicas (PIP 11220090100079), a la Universidad Nacional de San Martín
(Diálogo entre las ciencias) y a la Agencia de Promoción Científica y Tecnológica
(PICT 0717/2011) por el soporte financiero. NP agradece al CONICET por su
beca. RC y GC son investigadores del CONICET.
Referencias
Akinci, G.; Guven, D.E. (2011). Bioleaching of heavy metals contaminated sediment by pure and
mixed cultures of Acidithiobacillus spp. Desalination. 268: 221±226.
Curutchet, G.; Gutierrez, R.; Grimberg, S. (2012). Degradación ambiental y periferia urbana: un
estudio transdiciplinario sobre la contaminación en la región metropolitana de Buenos Aires.
Ambiente & Sociedade. 15: 173±194.
Carnicero, D.; Diaz, E.; Escolano, O.; Rubinos, D.; Ballesteros, O.; Barral, M.T.; Amils, R.; Garcia
Frutos, F.J. (2009). Preliminary study of neutralization and inhibition of chemolitotrophic
bacteria in an acid mine drainage from Río Tinto site. Proceedings of the International
Biohidrometallurgy Simposium, 2009.
Di Nanno, M. P.; Curutchet, G; Ratto, S. (2009). S, Zn, Cr, Cu and Fe changes during fluvial
sediments oxidation. Ciencias deI Suelo (Argentina.).27: 199±207.
Di Nanno, M.P.; Curutchet, G.; Ratto, S. (2007). Anaerobic sediment potential acidification and
metal release risk assessment by chemical characterization and batch resuspension
experiments. Soil Sediments. 7: 187±194.
Di Nanno, M.P. (2003). Acidificación por oxidación del azufre: aspectos microbiológicos y
geoquímicos del ciclo en rocas y sedimentos. Tesis de Maestría. Universidad de Buenos
Aires, área Ciencias del Suelo. Escuela para Graduados Alberto Soriano.
Eggleton, J.; Thomas, K. (2004). A review of factors affecting the release and bioavailability of
contaminants during sediment disturbance events. Environment International. 30: 973±980.
Förstner, U. (2004). Traceability of sediment analysis. Trends in Analytical Chemistry. 23: 217±236.
Gelman, F.; Binstock, R.; Halicz, L. (2012). Application of the Walkley±Black titration for the organic
carbon quantification in organic rich sedimentary rocks. Fuel. 96:698-610.
Kersten, M.; Förstner, U. (1991). Geochemical characterization of the potential trace metal mobility
in cohesive sediments. Geo-Marine Letters. 11: 184±187.
Lors, C.; Triffreau, C.; Laboudigue, A. (2004). Effects of bacterial activities on the release of heavy
metals from contaminated dredged sediments. Chemosfere. 56: 619±630.
Löser, C; Zehnsdorf, A; Hoffmann, P; Seidel, H. (2007). Remediation of heavy metal polluted
sediment by suspension and solid-bed leaching: Estimate of metal removal efficiency.
Chemosphere. 66:1699-1705.
Löser, C; Zehnsdorf, A; Görsch, K; Seidel, H. (2006). Remediation of heavy metal polluted sediment
in the solid bed: Comparison of abiotic and microbial leaching. Chemosphere. 65:9-16.
Morse, J.W.; Millero, F.J.; Cornwell, J.C.; Richard, D. (1987). The chemistry of hydrogen sulphide
and iron sulphide systems in natural waters. Earth Science Review. 24: 1±42.
Porzionato, N.; Mellota, M.; Candal, R.; Curutchet, G. (2013). Acid drainage and metal bioleaching
by redox potential changes in heavy polluted fluvial sediments. Advanced Materials Research.
825: 496-499. (doi:10.4028/www.scientific.net/AMR.825.496)
Porzionato, N.; Candal, R.; Curutchet, G. (2014). Biocatalyzed acidification and metal leaching
processes in sediments of polluted urban streams. International Journal of Environment and
Health. 7: 3-14.
Postgate, J. R. (1984). The sulfate reducing bacteria (2nd ed.). pp. 30±100. Cambridge University
Press, London and New York.
620
Quintana, M.; Curutchet, G.; Donati, E. (2001). Factors affecting chromium (VI) reduction by
Thiobacillus ferrooxidans. Biochemistry Engineering Journal. 9:11-15.
Rawlings, D. (2002). Heavy metal mining using microbes. Annual Review of Microbiolology. 56: 65±
91.
Sand, W. (2001). (Bio) chemistry of bacterial leaching-direct vs. indirect bioleaching.
Hidrometallurgy. 59: 159±175.
Seidel, H.; Wennrich, R.; Hoffmann, P.; Löser, C. (2006)a. Effect of different types of elementar
sulphur on bioleaching of heavy metals from contaminated sediments. Chemosphere. 62:
1444-1453.
Seidel, H.; Görsch, K.; Schümichen A. (2006)b. Effect of oxygen limitation on solid-bed bioleaching
of heavy metals from contaminated sediments. Chemosphere.65: 102-109.
Seidel, H.; Löser,C.; Zehnsdorf, A; Hoffmann, P.; Schmerold, R.(2004). Bioremediation process for
sediments contaminated by heavy metals: Feasibility study on a pilot scale. Environmental
Science Technology.38:1582-1588.
Silverman, M.P.; Lundgrem, D.G. (1959). Studies in the chemo-autotrophic iron bacterium
Ferrobacillus ferrooxidans: I. An improved medium harvesting procedures for securiting high
cell yields. Journal of Bacteriology. 77: 642±651.
Tabatabai, M.A. (1982). Sulphur. in Page, A.L. and Tabatabai, M.A. (Eds.): Methods of Soil
Analysis. Part 2. pp.501±534, American Society of Agronomy Inc., Madison.
Ure, A.M.; Quevauviller, P.; Montau, H.; Griepkin, B. (1993). Speciation of heavy metals in soil and
sediments. An account of the improvement and harmonization of extraction techniques
undertaken under the auspices of the BCR of the Commission of the European Communities.
International Journal of Environment and Analytical Chemistry. 51: 135±151.
Viera, M.; Curutchet, G.; Donati, E. (2003). A combined bacterial process for the reduction and
inmobilization of chromium. International Biodeterioration and Biodegradation. 52:31-34.
Ye, S.; Laws, E.A.; Gambrell, R. (2013). Trace element remobilization following the resuspension of
sediments under controlled redox conditions: City Park Lake, Baton Rouge; LA. Applied
Geochemistry. 28:91-99.
621
LIST OF COAUTHORS:
Dr. Gustavo Curutchet, Associated Professor.
3iA-UNSAM, Instituto de Investigación e Ingenieria Ambiental, Universidad
Nacional de San Martin; Av. 25 de Mayo y Francia - B1650HMP General San
Martín, Buenos Aires, Argentina.
Tel: (54-11) 4006-1500
Natalia Porzionato*, PhD Student.
Same address as that of Dr. Gustavo Curutchet.
Dr. Roberto Candal, Associated Professor.
Same address as that of Dr. Gustavo Curutchet
622
CHAPTER 6.7. COMPARACIÓN ENTRE BIOESTIMULACIÓN Y
BIOAUMENTACIÓN PARA LA BIORREMEDIACIÓN DE SUELOS
CONTAMINADOS CON HIDROCARBUROS
Diana Cabañas-Vargas*; Alma I. Corona-Cruz; Rafael Rojas-Herrera
Facultad de Ingeniería Química de la Universidad Autónoma de Yucatán. Campus de Ciencias
Exactas e ingenierías, Periférico Nte Km 33.5 Tablaje catastral 13615, col Chuburna de Hidalgo Inn
Mérida ,Yuc. cp 97203, tel (999)9460956 email:[email protected]
RESUMEN
Los microorganismos son el elemento catalizador en los procesos de
biorremediación, debido a que son considerados los responsables del proceso de
transformación de contaminantes como son, los hidrocarburos policíclicos
aromáticos (hpa). Debido a las características particulares de todo
microorganismo, el descubrimiento de nuevas especies, así como su
caracterización, es de gran interés, a fin de llevar a cabo una implementación
eficiente de este tipo de procesos evaluando todo bajo las condiciones locales. En
este estudio se realizó la biorremediación de un suelo contaminado con
hidrocarburos (petróleo) a escala de laboratorio, para evaluar la bioestimulación
frente a la bioaumentación con elementos propios de la región. Para la inoculación
del tratamiento por bioaumentación se utilizó un consorcio microbiano autóctono
con capacidad para degradar hidrocarburos. Se diseñaron 4 tratamientos: 1.
Blanco, que contenía Suelo (S-1Kg)+Petróleo crudo (P-40g), 2. S+P+C (residuos
putrescibles-200g), 3. S+P+F (fertilizante-200g), 4. S+P+ M (consorcio
microbiano). Se empleó suelo de la región. El consorcio de microorganismos
utilizado para estos tratamientos, fue aislado de suelo en las imediaciones de un
taller mecánico (Martín M., 2010). A las muestras tomadas se les analizó por un
lado, de manera general la presencia de mesofilicos aerobios (en medio ACS), así
como hongos y levaduras (PDA) y por otro el contenido de hidrocarburos totales
del petróleo (HTP) mediante una técnica gravimétrica en la que se realiza la
extracción (agitación-centrifugación) de los hidrocarburos del petróleo del suelo y
que está basada en los métodos 3500B y 3540C de la EPA.
La bioestimulación y la bioaumentación aumentaron la capacidad de
degradación de petróleo y las tasas de reproducción de los microorganismos,
observando que la bioaumentación reporto el mayor porcentaje de degradación
(78%). No se encontró diferencia significativa entre la bioestimulación dada por
fertilizante y el material putrescible (55% y 50%). Por otro lado los cálculos de UFC
señalan que la bioestimulación claramente favoreció el crecimiento de los
microorganismos.
Key words: Bioremediación, hidrocarburos, suelos, microorganismos.
-----------------*Author for correspondence:
623
Introducción
En Yucatán, aunque no es una región considerada de gran desarrollo industrial,
existen numerosas microempresas que utilizan derivados del petróleo o que los
manejan como resultado de los servicios que prestan. Los talleres mecánicos y de
mantenimiento de vehículos y maquinaria, representan uno de los rubros más
numeroso de la actividad comercial e industrial del estado con aproximadamente
1350 de estos negocios, de los que el 50% está situado en la ciudad de Mérida
(Cabañas et al. 2010). De este sector industrial y de servicios, se considera que el
90% son micro empresas por lo que no cuentan con instalaciones adecuadas para
la prevención de los derrames de hidrocarburos o para la descontaminación de los
sitios que por accidente se contaminen. Por tanto, se puede inferir que existen
cantidades significativas de suelos contaminados con hidrocarburos en la ciudad
de Mérida, que requieren de procesos de descontaminación .
En México se han utilizado varias alternativas de remediación pare la
contaminación
con
hidrocarburos,
por
ejemplo
incineración,
solidificación/estabilización, la extracción de vapores, lavado, etc., los cuales
pueden ser relativamente caros. En este contexto la biorremediación es una
herramienta con un gran potencial en este campo de estudio, la cual se ha
utilizado para la recuperación de suelos y sedimentos contaminados y los
ecosistemas terrestres. La mayoría de las tecnologías de biorremediación han sido
desarrolladas para tratar la contaminación de hidrocarburos derivados del petróleo
y transformarlos en productos menos peligrosos para la salud humana y el medio
ambiente (Juwarkar et al., 2010).
En el 85% de la superficie del estado de Yucatán predominan los suelos
denominados Leptosoles, que se caracterizan por su escasa profundidad, escasa
cantidad de tierra fina y con gran cantidad de piedras o afloramientos rocosos
(Bautista, et al. 2007) además de su pobre contenido de materia orgánica. La roca
que aflora frecuentemente se encuentra fracturada y el suelo es altamente
permeable, lo que ocasiona que los elementos contenidos en los suelos lixivien
fácilmente hacia los mantos freáticos.
El acuífero de Yucatán es un lente delgado de agua dulce que flota sobre agua
salada y que en la ciudad de Mérida tiene un espesor aproximado de 61 metros
(Marín et al. 2004). Es un acuífero cárstico en el que la permeabilidad es una
combinación de la permeabilidad primaria de la roca, del grado de fracturamiento y
de los conductos de disolución. Estas condiciones ocasionan que los mantos
freáticos sean altamente vulnerables y fáciles de contaminar.
Por otro lado, durante un estudio realizado en 2006 para el estado de Yucatán
se detectaron aproximadamente 1350 talleres mecánicos o de servicio para
maquinaria y equipo, de los cuales aproximadamente el 50% se ubican en el
municipio de Mérida, el 30% en los tres municipios más poblados y los restantes
repartidos en los otros 102 municipios (Cabañas, et al. 2010). El 90% de estos
talleres mecánicos vierten aceites automotrices gastados y otros hidrocarburos
directamente a los suelos, lo que representa una contaminación significativa
directamente a los suelos e indirectamente hacia los mantos freáticos cuando se
lixivian en épocas de lluvias.
624
El presente trabajo presenta la comparación entre la boestimulación y la
bioaumentación como mecanismos de biorremediación de suelos contaminados
con hidrocarburos.
Los experimentos. Se utilizaron 20 kilogramos de suelo tipo arcilloso de la región,
con tamaños de partículas entre 2 y 2.36 mm, para facilitar el proceso aeróbio y
lograr una distribución homogénea del contaminante. La zona de recolección
estuvo distante de zonas de contaminación.
Como material contaminante se utilizaron 500 gramos de petróleo crudo
obtenido en la sonda de Ciudad del Carmen Campeche.
Para la inoculación del tratamiento por bioaumentación se utilizó un consorcio
microbiano autóctono aislado de suelo en las inmediaciones de un taller mecánico
con capacidad para degradar hidrocarburos, el cual fue aislado y estudiado en
trabajos previos y que está formado principalmente por Sphingomona, Dietzia,
Amycolatopsis, Thermobifida y Microbacterium (Martin López 2011).
Se diseñaron 4 tratamientos: 1. Control, que contenía Suelo (S-1Kg)+Petróleo
crudo (P-40g); 2. S+P+C (residuos putrescibles-200g); 3. S+P+F (fertilizante200g); 4. S+P+ M (consorcio microbiano). Los experimentos se montaron en
recipientes de cristal redondos de 25 cm de diámetro. Los nutrientes empleados
fueron: Fertilizante (Vigoro ®) y Composta doméstica (residuos putrescibles). Los
tratamientos se muestrearon semanalmente durante un periodo de 5 semanas y
fueron mezclados con regularidad. A las muestras tomadas se les analizó por un
lado, de manera general la presencia de mesofílicos aerobios en medio agar
cuenta estándar (ACS), así como hongos y levaduras (PDA) y por otro el
contenido de hidrocarburos mediante una técnica gravimétrica en la que se realiza
la extracción (agitación-centrifugación) de los hidrocarburos del petróleo del suelo
y que está basada en los métodos 3500B y 3540C de la EPA. En la tabla 1, se
describen los experimentos realizados.
TABLA 1. Descripción de los experimentos de biorremediación y
bioaumentación realizados.
Experimentos Elementos presentes en el experimento
Control
Suelo(1kg) + Petróleo(40g)
M
Suelo(1Kg) + petróleo (40g)+ consorcio microbiano (490 ml)
C
Suelo(1kg) + Petróleo (40g)+ residuos putrescibles (200g)
(material de Composteo)
F
Suelo + Petróleo+ fertilizante comercial (200g)
625
Al control y los experimentos C y F se les agregaron 490 ml de agua
aproximadamente para igualar la humedad alcanzada en el experimento con los
microorganismos. Durante las 5 semanas del experimento la humedad se ajustó
para estar siempre entre 30% y 50%.
Los experimentos se realizaron por duplicado a excepción del de control.
Técnicas de análisis. El contenido de hidrocarburos se determinó mediante una
técnica gravimétrica en la que se realiza la extracción (agitación-centrifugación) de
los hidrocarburos del petróleo del suelo y que está basada en los métodos 3500B
y 3540C de la EPA.
Para el conteo microbiano, se realizaron las diluciones seriadas necesarias
(10-1 ± 10-9) en tubos de ensayo con solución salina. Se sembraron en los medios
correspondientes (PDA o ACS) por triplicado para registrar el promedio del
crecimiento de los microorganismos. A partir del crecimiento obtenido, pasadas
24-48 horas, en cada una de las cajas de Petri sembradas, se realizó el conteo de
colonias y se calcularon las unidades formadoras de colonias por gramo de suelo
(UFC/g).
Resultados y discusiones.
La concentración de Hidrocarburos totales del petróleo en cada una de las
muestras en los diferentes tiempos, usando la siguiente ecuación:
El peso de la muestra de suelo inicial fue de 2gramos para todas las muestras.
La tabla 2 muestra los resultados obtenidos en los experimentos y sus réplicas
durante el tiempo de experimentación.
TABLA 2. Concentración de los HTP durante 5 semanas de experimentación
Muestra/Semana
B
C
C+
F
F+
M
M+
0
0.00815
0.01555
0.02135
0.01
0.0143
0.0202
0.0062
1
0.00815
0.0149
0.01675
0.00855
0.01265
0.0161
0.0046
2
0.00815
0.01425
0.01225
0.0071
0.0111
0.01205
0.00315
3
0.00815
0.0139
0.0078
0.0056
0.00955
0.0079
0.00245
4
0.00815
0.01315
0.0032
0.0043
0.00805
0.00495
0.00205
5
0.00815
0.01285
0.00315
0.00415
0.00795
0.00395
0.00165
626
Environmental Biotechnology and Engineering ± 2014
La Figura 1, que presenta los valores promedio de los experimentos duplicados, muestra
la disminución de los hidrocarburos totales del petróleo a lo largo de las 5 semanas.
Los resultados muestran que tanto la bioaumentación (M) como la bioestimulación (F,
C) favorecieron la disminución de la concentración de hidrocarburos en los suelos
contaminados (figura1).
FIGURA 1. Disminución en la concentración de HTPs (valores promedio).
El experimento llamado control (C), al cual no se le adicionó ningún sustrato
permaneció sin cambio significativo durante las 5 semanas de experimentación. La
presencia de residuos putrescibles, con los cuales se simuló un proceso de Composteo,
logró hasta el 56% de remoción de los HTPs. La presencia del fertilizante logró una
remoción de HTPs de hasta un 50% (Fig.1).
En este trabajo se observó que la opción que presentó el mayor porcentaje de
degradación del petróleo fue el experimento al que se le agregó el consorcio autóctono, el
cual obtuvo un porcentaje de degradación promedio de 78%. Estos niveles de
degradación se han alcanzado con consorcios microbianos comerciales a los cuales se
les añaden biosurfactantes (Ortiz y col., 2005). En este caso no se utilizaron agentes
surfactantes.
Crecimiento de microorganismos. El crecimiento de los microorganismos en los medios
agar papa dextrosa para cuantificar hongos y levaduras y agar cuenta estándar para los
mesofilicos aerobios, se presenta en las figuras 2 y 3. Se puede observar en ambas
gráficas que el experimento con bioaumentación en el cual se incluyó el consorcio
microbiano, alcanzo niveles mayores de crecimiento, el cual fue disminuyendo en la
medida que los nutrientes disponibles de fácil metabolización se fueron agotando. Los
microorganismos son altamente especializados y juegan un papel específico en la
descomposición de compuestos orgánicos. La habilidad de una comunidad microbiana
627
para descomponer compuestos orgánicos complejos depende primeramente de la
versatilidad metabólica de sus miembros individuales. Los consorcios de especies
microbianas mixtas pueden dar un mayor espectro de posibilidades siempre y cuando
tengas las condiciones ambientales y nutrimentales adecuadas (Alvarez y Guevara,
2003).
En los experimentos de bioestimulación en donde fue adicionado por un lado residuos
putrescibles y por otro un fertilizante comercial, se puede observar en las figuras 2 y 3 que
la respuesta fue similar, iniciando con una disminución en el crecimiento probablemente
debido a la presencia del contaminante y posteriormente presentaron recuperaciones en
sus tasas de crecimiento, en particular se puede observar una mejor respuesta ante la
presencia de residuos putrescibles que ante la presencia de solo fertilizante (Castelblanco
y Niño Avendaño, 2011).
FIGURA 2. Respuesta del crecimiento de hongos y levaduras en medio agar papa
dextrosa, ante las condiciones de bioaumentación o bioestimulacion con residuos
putrescibles (RP) o fertilizante (F) en un suelo contaminado con petróleo.
628
FIGURA 3. Respuesta del crecimiento de mesofilicos aerobios en medio agar cuenta
estándar, ante las condiciones de bioaumentación o bioestimulación con residuos
putrescibles (RP) o fertilizante (F) en un suelo contaminado con petróleo.
Conclusión
x La bioestimulación y la bioaumentación aumentaron la capacidad de degradación de
petróleo y las tasas de reproducción de los microorganismos.
x La bioaumentación con microorganismos autóctonos reportó el mayor porcentaje de
degradación (78%).
x No se encontró diferencia significativa entre la bioestimulación dada por fertilizante y
el material putrescible (50 a 56%).
x Los cálculos de UFC señalan que la bioestimulación claramente favoreció el
crecimiento de los microorganismos.
Agradecimientos
Se agradece a la Facultad de Ingeniería Química de la Universidad Autónoma de Yucatán
por el financiamiento del proyecto Bioremediación de Suelos Contaminados con
Hidrocarburos, ya que este trabajo se derivó del mencionado proyecto.
629
Referencias
Alvarez Alvarado P. J. y Guevara Perez, E., Biorremediación y atenuación natural de acuíferos
contaminados por sustancias químicas peligrosas. Consejo de desarrollo científico y humanístico de la
universidad de Carabobo, Venezuela. 2013.
%DXWLVWD)$JXLODU<5LYDV+\3iH]5/RVVXHORVGHOHVWDGRGH<XFDWiQ(Q,PSRWDQFLDGHO
%LQRPLR6XHOR±0DWHULD2UJiQLFDHQOH'HVDUUROORVRVWHQLEOHPp 11-43.
&DEDxDV'' 5H]D* 6DXUL050pQGH]5, %DXWLVWD) 0DQULTXH:5RGUtJXH]( %DODQFiQ $0 \
0HGLQD 5 ,QYHQWDULR GH IXHQWHV SRWHQFLDOHV GH UHVLGXRV SHOLJURVRV HQ HO HVWDGR GH <XFDWiQ
0p[LFRRevista Internacional de Contaminación Ambiental. 26 (4) 269-277, 2010.
Castelblanco Casallas I.F. y Niño Avendaño J.A., Manejo y tratamiento actual de residuos aceitosos en la
industria petrolera colombiana. Universidad Industrial de Santander. Bucaramanga, Co. Tesis. 2011.
Juwarkar, A., Sanjeev, S. y Ackmez, M. (2010). A comprehensive overview of elements in
bioremediation.Reviews in Environmental Science and Biotechnology. Vol.9, pp. 215-288.
Martín López M. Aislamiento y caracterización de un consorcio microbiano degradador de hidrocarburos..
Facultad de Ingeniería Química. Universidad Autónoma de Yucatán. Tesis 22 de noviembre del 2012.
Ortiz, E.; R. Núñez.; E. Fonseca; J. Oramas; V Almazán; Y. Cabranes; A. Miranda; O .Barbán, C. Martínez;
Y. Díaz; y G. Borges. BIORREMEDIACION DE SUELOS CONTAMINADOS CON HIDROCARBUROS.
Revista Contribución a la Educación y la Protección del Medio Ambiente. (6): 51-60. Pág. 51-60 2005
630
LISTA DE COAUTORES
Diana Cabañas-Vargas (presenting), PhD
Facultad de Ingeniería Química de la Universidad Autónoma de Yucatán. Campus de
Ciencias Exactas e ingenierías, Periférico Nte Km 33.5 Tablaje catastral 13615, col
Chuburna de Hidalgo Inn Mérida, Yuc. CP 97203, tel (999)9460956
email:[email protected]
Alma I. Corona-Cruz; Dr.
Chuburna de Hidalgo Inn Mérida, Yuc. CP 97203, tel (999)9460956 email:[email protected]
Rafael Rojas-Herrera, Dr.
Chuburna de Hidalgo Inn Mérida, Yuc. CP 97203, tel (999) 9460956 .
631
CHAPTER 6.8. BIODEGRADACIÓN DE ATRAZINA EN SUELOS AGRÍCOLAS DEL
ESTADO DE MORELOS MÉXICO
Ana Gabriela Ortega-Hernández1; Luis Gerardo Treviño-Quintanilla1;
Carlos Díaz-Gómez1; Rosa Angélica Guillén-Garcés1*
(1) Dirección de Biotecnología y Tecnología Ambiental, Universidad Politécnica del Estado de Morelos. Blvd.
Cuauhnáhuac 566, Col. Lomas del Texcal, 62550. Jiutepec, Morelos, México.
Resumen.
La Atrazina es el tercer herbicida más utilizado en México en la eliminación de maleza de
hoja ancha en cultivos de sorgo, caña de azúcar, trigo y piña (Whitacre &Ware, 2004). Es
un compuesto poco soluble en agua y posee una alta movilidad en el suelo, aumentando
el riesgo de contaminación de los cuerpos de agua. El uso de Atrazina está prohibido en
la Unión Europea y restringido en Estados Unidos (Beth y Colangelo 2006). En México no
existe una regulación para su uso. El estado de Morelos es productor de maíz, sorgo y
caña de azúcar y este herbicida es ampliamente usado en el estado. Objetivo general
Caracterizar la biodegradación de Atrazina en diferentes muestras de suelo agrícola del
Estado de Morelos. Metodología Se hicieron muestreos en tres municipios del estado,
Moyotepec, Yecapixtla, Huautla, con base al grado de permeabilidad. El suelo fue
caracterizado fisicoquímicamente y se realizaron cinética de biodegradación en las
diferentes muestras de suelo, aplicando una concentración similar a lo recomendado en
campo. Se aplicó Atrazina al 100 % y el producto comercial con 43 % del ingrediente
activo. Las condiciones experimentales fueron 50 % de la capacidad de campo y se
incubaron a 32 ±1°C durante 29 días. Resultados Las velocidades de biodegradación de
la Atrazina al 43 en los suelo de Moyotepec, Yecapixtla y Huautla fueron de 0.004d -1,
0.0375d-1 y 0.0174 d-1; con vidas medias de 173.28 18.48 y 39.84 días con porcentajes de
biodegradación de 23.2, 71.8 y 50.1%, respectivamente, las cuales fueron más lentas que
la biodegradación de atrazina al 100 %. En suelos donde había mayor concentración de
materia orgánica y una aplicación recurrente del herbicida las velocidades de degradación
fueron mayores, debido probablemente a la presencia de microorganismos capaces de
degradar el herbicida. Conclusión, Con base en la velocidad de degradación de Atrazina
al 100 % y Atrazina comercial se pudo observar que la presencia de aditivos y
coadyuvantes contenidos en el producto comercial disminuye la velocidad y porcentaje de
biodegradación del herbicida.
632
Abstract
Atrazine is the third most widely used herbicide in Mexico in eliminating broadleaf weeds in
sorghum, sugar cane, wheat and pineapple (Whitacre & Ware, 2004). Is slightly watersoluble compound and has high mobility in soil, increasing the risk of pollution of the
surface water and ground. The use of atrazine is banned in the European Union and
restricted in the U.S. (Beth y Colangelo 2006). In Mexico there is no regulation for use.
The state of Morelos is use in corn, sorghum and sugarcane. Objetive. Characterize the
biodegradation of atrazine in different samples of agricultural ground in the State of
Morelos. Methods. Samplings were made in three municipalities in the state, Moyotepec,
Yecapixtla, Huautla, based on the degree of permeability. The soil was characterized
physicochemically and biodegradation kinetics were made in different soil samples,
applying in similar field recommended concentration. Atrazine was applied at 100% and
the commercial product with 43% active ingredient. The experimental conditions were 50%
of field capacity and incubated at 32 ± 1 ° C for 29 days. Results. The rates of
biodegradation of atrazine (43 %) in the ground Moyotepec, Yecapixtla and Huautla were
0.004d-1, 0.0375d-1 and 0.0174 d-1; with half-lives of 173.28, 18.48 and 39.84 days and
percentage biodegradation of 23.2, 71.8 and 50.1% respectively, results which were
slower than biodegradation atrazine (100%). Soils which had higher concentrations of
organic matter and continuous application of herbicide degradation rates were higher,
probably due to the presence of microorganisms capable of degrading the herbicide.
Conclusion. The presence of additives contained in the commercial product decreases
the percentage biodegradation rate of the herbicide in the soils studied.
Palabras clave: Atrazina, Adsorción, Aditivos, Biodegradación, Disipación, Disponibilidad.
Introducción.
La atrazina (6-Cloro-N2-etil-N4-isopropil-1,3,5-triazina-2,4-diamina) con número CAS:
1912-24-9 es un herbicida triazinico selectivo pre y post emergente ampliamente usado
para la eliminación de maleza de hoja ancha en cultivos de sorgo, caña de azúcar, trigo,
soya, frijol, plátano y piña, entre otros. Su principal nombre comercial es Gesaprim el cual
contiene desde 46 a 90 % de ingrediente activo según su presentación como suspensión
o granular, respectivamente y sustancias humectantes y dispersantes. A nivel mundial su
uso se inició en 1958 y se ha incrementado en los últimos 50 años (Golla et al., 2011) en
México su uso inició 1975 y actualmente se sigue empleando (SAGARPA 2007). El uso
de atrazina está relacionado con efectos agudos como irritación de ojos, piel, nariz y
garganta; efectos crónicos como alteraciones en las funciones del corazón, hígado y riñón
y problemas de reproducción y alterando los niveles de hormonas que afectan la
ovulación; malformaciones en seres humanos además de que se considera como posible
cancerígeno en seres humanos (Rusiecki, 2004, ATSDR 2003); asimismo estudios han
determinado que es un inductor de hermafrodismo en ranas en concentraciones tan bajas
et al. 2002). Debido a sus características fisicoquímicas, movilidad
y persistencia en el ambiente se ha considerado un contaminante orgánico persistente,
aunque su clasificación e integración al anexo de Estocolmo está en revisión. Es un
compuesto poco soluble en agua y posee una alta movilidad en el suelo, aumentando el
633
riesgo de contaminación de los cuerpos de agua superficiales y subterráneos debido a los
procesos de infiltración y escorrentía.
En su estructura cuenta con un anillo heterocíclico un radicales cloro, etil y isopropil
(Worthing y Phill, 1987) y los procesos que intervienen en la atenuación natural de la
atrazina, son la biodegradación o degradación biológica como el principal mecanismos de
disipación en el ambiente (Ma y Selim 1996), y este es afectado por factores como por
tipo de suelo, el contenido de materia orgánica (Ling et al., 2005) y pH (Tyess et al.,
2006). Considerando que existen diversos factores que afectan la biodegradación de
atrazina se ha encontrado un intervalo amplio entre las vidas media de este compuesto
obteniendo valores que va desde 0.99 hasta 5,824 días (Guillén-Garcés et al., 2007,
Schoen y Winterlin, 1987).
En el caso del tipo de suelo se ha encontrado que en suelos arenosos la disponibilidad
de los contaminantes aumenta en comparación con los suelos arcillosos, esto debido
principalmente a que las arcillas tienen mayor capacidad de adsorción que las arenas, por
lo que las primeras pueden limitar los procesos de biodegradación debido a que los
compuestos no se encuentran disponibles para ser degradados por los microorganismos
(Huang et al., 2013). Con respecto a la materia orgánica se ha encontrado que suelos con
alto contenido puede limitar la biodegradación de atrazina, ya que el herbicida tiende a
adsorberse en esta fracción orgánica del suelo (González-Márquez y Hansen, 2009),
aunque esta limitación también dependerá de la presencia de microorganismos capaces
de degradar la atrazina, debido a que se ha demostrado que la aplicación frecuente del
herbicida en el suelo tiende a modificar la población microbiana presente (Ross et al.,
2006), permitiendo así la inducción de enzimas en los microorganismos y la adaptación de
las comunidades microbianas a la presencia de atrazina, favoreciendo de esta manera la
biodegradación del herbicida en el suelos (Mahía y Díaz-Raviña 2007). Otro factor que
influye sobre la biodegradación de la atrazina es el pH del suelo, al ser la atrazina un
compuesto que presenta característica de base débil este protona a pH´s por debajo de 4,
por lo que tiende a adsorberse en suelos con pH ácidos (Wang et al. 2011).
Por estas características es un compuesto que se ha caracterizado como móvil
encontrándose en diversos cuerpos de agua tanto superficiales como subterráneos (Farré
et al. 2007). Por lo que su límite máximo permisibles de concentración en agua de
-1.
Además de que se ha prohibido su uso en la Unión Europea y restringido en países como
Estados Unidos y Australia (Beth y Colangelo 2006), mientras que en Asia, Canadá,
América Latina, el Caribe y México es permitido (FAO 2005, PMRA 2007, UNEP 2002).
En México es el tercer herbicida más utilizado, se ha reportado que sólo en los
distritos de riego del país se utilizan 1,078 toneladas anuales (Villada ±Canela 2006) y su
uso se tiene registrado ante la Comisión Intersecretarial para el Control del Proceso y Uso
de Plaguicidas y Sustancias Tóxicas (CICOPLAFEST), pero no existe una regulación de
su aplicación. El estado de Morelos es productor de maíz, sorgo y caña de azúcar y este
herbicida es ampliamente usado en este lugar, hasta la fecha no se ha caracterizado la
dinámica (biodegradación, adsorción e infiltración) de este compuesto en suelos agrícolas
de la zona antes mencionada, por lo que el objetivo del presente trabajo fue caracterizar
la biodegradación de atrazina en diferentes muestras de suelo agrícola del Estado de
Morelos.
634
Metodología.
Se hicieron muestreos en tres municipios del estado: Yecapixtla, Moyotepec y Huautla,
con base al grado de permeabilidad alta, media y baja, respectivamente. El método de
muestreo fue de tipo superficial considerando sólo la capa arable (30 cm de profundidad)
y de los terrenos se tomaron 21 muestras simples y se mezclaron para obtener una
muestra compuesta. La muestra compuesta se tamizó a través de una malla No. 6
correspondiente a 4.76 mm de apertura para homogenizarla. Los suelos fueron
caracterizados fisicoquímicamente determinando parámetros tales como: pH, humedad,
materia orgánica, capacidad de campo, densidad aparente y textura. Una vez que se
caracterizaron los suelos se realizaron cinética de biodegradación por triplicado en 250 g
de suelo, se aplicó atrazina en concentración similar a la recomendada para cultivos de
maíz (2 kg ha-1), la aplicación se realizó en dos presentaciones, la primera como
ingrediente activo al 100 % y la segunda como producto comercial que contiene 43 % del
ingrediente activo. Las condiciones experimentales fueron 50 % de la capacidad de
campo y se incubaron a 32 ±1°C durante 29 días. A diferentes tiempos se tomó una
muestra de 25 g del suelo y se le adicionó 25 mL de una mezcla de solventes compuesta
por acetona: metanol: ácido acético en una proporción de 63:25:12 v/v para la extracción
del ingrediente activo, se colocó en agitación durante 17 horas a 120 rpm, posteriormente
las muestras se centrifugaron a 6000 rpm para obtener el sobrenadante, el cual fue
evaporado hasta 1 mL en un rotaevaporador a una temperatura de 70 a 75°C a 150 rpm,
posteriormente la atrazina contenida en el sobrenadante se extrajo con diclorometano, el
cual se evaporó hasta alcanzar 1 mL, después de esta extracción se le adicionó 30 mL de
metanol y se evaporó hasta sequedad, por último los residuos obtenidos se disolvieron en
10 mL de la mezcla de acetona: metanol y ácido acético y de este extracto se inyectaron
40 µl al equipo de cromatografía de líquidos de alta resolución (HPLC), la cuantificación
de la atrazina se realizó en una columna C-18 con un tamaño de partícula de 5 µm y
dimensiones de 4.5 x 250mm , con un detector de UV-VIS a 220 nm; flujo de 1 mL min-1,
la composición de las fases utilizadas fue (A) fosfato dibásico de potasio (K 2HPO4) 20 mM
pH 7 y (B) acetonitrilo los gradientes de solventes se detalla en la tabla 1.
TABLA 1. Gradientes de fases para cuantificación de atrazina por medio de HPLC
Tiempo (min)
0
10
12
18
A(%)
80
70
70
50
B (%)
20
30
30
50
635
Resultados
En la caracterización se obtuvieron suelos ligeramente ácido (Yecapixtla y Huautla) y
superior a un pH neutro como Moyotepec (Tabla 2). Considerado que la atrazina es una
EDVHGpELOTXHSURWRQDDEDMRGHS+VHIDYRUHFHVXDGVRUFLyQHQVXVWDQFLDVK~PLFDVD
pH ácidos (Ahmad y Rahman 2009), por lo que en los pH encontrados en las muestras de
suelo del estado de Morelos la atrazina puede estar disponible para los proceso de
biodegradación.
TABLA 2. Caracterización fisicoquímica de las muestras de suelo procedentes del
Estado de Tabasco.
Parámetro
Muestra de suelo procedentes de:
Moyotepec
Yecapixtla
Huautla
pH1
7.55
6.48
6.20
2
Humedad [%]
15.03
19.50
15.90
Materia orgánica [%]2
2.16
6.60
4.67
3
Capacidad de campo [%]
16.30
24.52
14.48
Porosidad [m3 m-3]3
0.35
0.37
0.426
5
Textura
Franco
Franco
Franco
limoso
arcilloso
arenoso
1 ASTM D4972-89, 2 ASTM D2974 00, 3F1815-97, 4ASTM F1815-97, 5 F1632-00
Con respecto a la materia orgánica el suelo que presentó mayor porcentaje fue el de
Yecapixtla con un 6.60 %, así como de arcilla presentando una textura franco arcillosa
(Tabla 2). Esto puede indicar que el proceso de adsorción de atrazina tiene mayor
influencia sobre la biodegradación en comparación con los suelos de Huautla y
Moyotepec, ya que estudios han reportado que en presencia de materia orgánica y arcilla
el efecto de la adsorción limita procesos como la biodegradación de la atrazina (Ling et
al., 2005).
Las cinéticas de biodegradación para los suelos donde se aplicó atrazina como
ingrediente activo (100 %) se presentan en la figura 1.
636
FIGURA 1. Cinética de degradación de atrazina 100 %
A partir de los experimentos de degradación se determinaron los parámetros cinéticos
tales como: velocidades de biodegradación, vidas medias y porcentajes de degradación
obteniendo valores para los tres suelos (Tabla 3).
TABLA 3. Parámetros cinéticos obtenidos en los experimentos de degradación de
atrazina al 100 % en suelos agrícolas del Estado de Morelos
Parámetro
Velocidad de
biodegradación [d-1]
Vidas medias [d]
Degradación [%]
Moyotepec
0.038±0.0021
Yecapixtla
0.045±0.0036
Huautla
0.162±0.023
18.34±1.08
57.4±4.08
15.55±1.21
72.86±2.37
4.36±0.712
65.79±1.28
En los tres suelos se presentó biodegradación del herbicida, es decir que las muestras
contenían microorganismos capaces de asimilar a la atrazina como fuente de carbono y
nitrógeno. Por otro lado, se encontró que en el suelo de Huautla el herbicida se encontró
más biodisponible para el proceso que en los otros dos, esto debido probablemente a que
la concentración de materia orgánica fue media en comparación de los otros dos suelos y
el suelo se clasificó como franco arenosos, ya que entre menor sea la presencia de
materia orgánica y más arena se favorece la disponibidad del herbicida a los
microorganismos, y se limita el proceso de adsorción del herbicida al suelo (Ling et al.
2005).
En el caso del suelo de Yecapixtla la velocidad de biodegradación fue menor, esto
puede ser debido a que la concentración de arcilla y materia orgánica (Ling et al., 2005)
fue mayor que en los otros dos suelos, limitando la biodegradación del herbicida.
Obteniendo una vida media cuatro veces mayor (15.55 días) al suelo de Huautla (4.36
días).
637
Con respecto a las cinéticas de biodegradación con atrazina comercial al 43 % estas
fueron más lentas en los suelos de Moyotepec y Huautla en comparación con los
experimentos realizados con atrazina pura al 100 % (Figura 2). Mientras que en el suelo
de Yecapixtla el proceso de bidegradación fue similar al obtenido en la biodegradación
con atrazina como ingrediente activo (100 %). Lo anterior puede ser debido a que la
atrazina comercial contiene en su composición, aditivos y coadyuvantes, que afectaron la
velocidad y porcentaje de biodegradación del herbicida. Aunque esto sólo se observó en
los suelos donde la cantidad de materia orgánica y arcilla fue menor, esto debido
probablemente a que la presencia de estos compuestos favorece otros procesos de
disipación como la adsorción y limita la disponibilidad del herbicida al proceso de
biodegradación en suelos con bajo contenido de materia orgánica y arcilla. Mientras que
en el caso de las muestras de Yecapixtla el proceso de degradación no fue afectada,
debido posiblemente a que la concentración de materia orgánica y arcilla no permitió
visualizar el efecto de los aditivos contenidos en la atrazina comercial, por lo que los
parámetros cinéticos obtenidos fueron similares a los experimentos donde se aplicó
atrazina al 100%.
FIGURA 2. Cinética de degradación de atrazina comercial (43 %)
Los parámetros cinéticos obtenidos para la atrazina comercial (43 %) se presentan en
la tabla 4.
638
TABLA 4. Parámetros cinéticos obtenidos en los experimentos de degradación de
atrazina al 43 % en suelos agrícolas del Estado de Morelos
Parámetro cinético
Velocidad de degradación [d-1]
Vidas medias
Degradación [%]
Moyotepec
Yecapixtla
0.005±0.0006 0.0316±0.003
138.64±15.47
22.13±2.26
21.4±2.52
71.77±1.43
Huautla
0.0239±0.005
30.67±7.89
50.10±3.65
Con respecto a la clasificación de permeabilidad para cada uno de los suelos y su
relación con las velocidades de biodegradación de atrazina al 100 % el suelo con
permeabilidad alta y media (Moyotepec y Yecapixtla) puede tener mayor riesgo de
infiltración si el único proceso de disipación fuera la biodegradación debido a que fueron
las velocidades más lentas. Con respecto a la atrazina comercial (43 %), el mayor riesgo
de infiltración es en el suelo de Moyotepec debido a que la biodegradación fue más lenta
aumentado el riesgo de infiltración cuando se aplica en esta presentación.
Conclusiones.
Los experimentos de biodegradación permitieron obtener los parámetros cinéticos de
atrazina pura y comercial en muestras de suelo agrícola del Estado de Morelos.
Las velocidades de biodegradación de atrazina pura fueron cuatro veces mayor en los
suelos de Huautla (0.162 d-1) en comparación con los suelos de Moyotepec (0.038 d -1) y
Yecapixtla (0.045 d-1).
Aun cuando la biodegradación de atrazina al 100 % en el suelo de Yecapixtla fue la
más lenta alcanzó hasta un 72 % en los primeros 29 días. Mientras que aunque en los
otros suelos la degradación fue más rápida el porcentaje de degradación fue menor al 65
%.
La presencia de aditivos y coadyuvantes contenidos en la atrazina comercial limita la
degradación de atrazina por un orden de magnitud por debajo de la atrazina pura en
suelos franco limosos y franco arenosos. Pero no en suelos arcillos con alta concentración
de materia orgánica.
639
Referencias.
Ahmad R. y Rahman A. (2009). Sorption characteristics of atrazine and imazethapyr in soils of New Zealand:
Importance of independently determined sorption data. J. Agric. Food Chem. 57, 10866±10875.
ASTM (American Society for Testing and Materials). 1989. Standard test method for pH of soils. D4972-89.
En annual book of ASTM standards, Vol 04.08. Philadelphia, PA, United State American, pp 1035-1037.
ASTM (American Society for Testing and Materials). 1997. Standard test method for saturated hydraulic
conductivity, water retention, porosity, particle density, and bulk density of putting green and sports turf
root zones. F1815-97. En Annual book of ASTM standards, Vol 15.07. Philadelphia, PA, USA.
ASTM (American Society for Testing and Materials). 2000. Standard test methods for moisture, ash, and
organic matter of peat and other organic soils. D2974-00. En Annual Book of ASTM Standards, Vol
04.08. Philadelphia, PA, USA.
ASTM (American Society for Testing and Materials). 2000c. Standard test methods for particle size analysis
and sand shape grading of golf course putting green and sports field rootzone mixes. F1632-00. In
Annual Book of ASTM Standards. Vol. 15.07. West Conshohocken, PA. 11 p.
ATSDR (2003). Resumen de Salud Pública (Atrazina) CAS#: 1912-24-9. Agencia para Sustancias Toxicas y
Registro de Enfermedades. Departamento de Salud y Servicios Humanos de los EUA disponible en
línea http://www.atsdr.cdc.gov/es/phs/es_phs153.pdf, consultado 29/08/2013.
Beth J. y Colangelo A. (2006). European Union Bans Atrazine, While the United States Negotiates
Continued Use, Int. J. Occup. Environ. Health.12:260±267.
Buttiglieri G, Migliorisi L, Malpei F, 2011, Adsorption and removal at low atrazine concentration in an MBR
pilot plant. Water Sci Technol. 2011;63(7):1334-40.
FAO (2005). Proceedings of the Asia regional workshop on the implementation, monitoring and observance
of the international code of conduct on the distribution and use of pesticides. Food and Agriculture
Organization. Bangkok, Tailandia. 249 p.
Farré M., Martínez E., Ramón J., Navarro A., Radjenovic J., Mauriz E., Lechuga L., Marco M. y Barceló
D.(2007). Part per trillion determination of atrazine in natural water samples by a surface plasmon
resonance immunosensor. Anal Bioanal Chem. 388, 207±214.
Golla V., Nelms J., Taylor R. y Mishra S. (2011). Pesticide concentrations in drinking water from farm homes:
Variation between community water supply and well-water. J. Environ. Sci. Eng. 5, 955-961.
González-Márquez L.C. y Hansen A.M. (2009). Adsorción y mineralización de atrazina y relación con
parámetros de suelo del DR 063 Guasave, Sinaloa. Rev. Mex. Cienc. Geol. 26, 587-599.
Guillén-Garcés R. A., Hansen A. M. y Afferden van M. (2007). Mineralization of atrazine in agricultural soil:
Inhibition by nitrogen. Environ. Toxicol. Chem. 26, 844-850.
Hayes T.B., Collins A., Lee M., Mendoza M., Noriega N., Stuart A. y Vonk A. (2002). Hermaphroditic,
demasculinized frogs after exposure to the herbicide atrazine at low ecologically relevant doses. PNAS
99, 5476-5480.
Huang Y., Liu Z., He Y., Zeng F., 2013, x| et al. Quantifying effects of primary parameters on adsorption±
desorption of atrazine in soils. Journal of Soil and Sediments, 13:82-93.
Krutz L. y Shaner D. (2008) Atrazine Dissipation in s-Triazine±Adapted and Nonadapted Soil from Colorado
and Mississippi: Implications of Enhanced Degradation on Atrazine Fate and Transport Parameter, J.
Environ. Qual 37, 848±857.
Ling, W.T., Wang, H.Z., Xu, J.M., Gao, Y.Z. (2005). Sorption of dissolved organic matter and its effects on
the atrazine sorption on soils. J. of Env. Sci. 17, 478-48.
Ma L., Selim, H.M. (1996). Atrazine retention and transport in soils. Rev. Env. Cont. Tox. 145, 129-73.
Mahía J. y Díaz-Raviña M. (2007). Atrazine degradation and residues distribution in two acid soils from
temperate humid zone. J. Environ. Qual. 36, 826-831.
PMRA (2004). Re-evaluation Decision Atrazine (Environmental Assessment). Pest Management Regulatory
Agency,
Health
Canada.
Ottawa,
Ontario.
Page
1.
En
línea
http://publications.gc.ca/collections/collection_2008/pmra-arla/H113-28-2007-5E.pdf
Ross M., Goberna M., Moreno J.L., Hernández T., García C. e Insam H. (2006). Molecular and physiological
bacterial diversity of a semiarid soil contaminated with different levels of formulated atrazine. Appl. Soil
Ecol. 34, 93-102.
640
Rusiecki, J.A., De Roos, A., Lee, W.J., Dosemeci, M., Lubin, J.H., Hoppin, J.A., Blair, A.,Alavanja, M.C.,
2004. Cancer incidence among pesticide applicators exposed to atrazine in the agricultural health study.
Journal of the National Cancer Institute 96, 1375±1382Villada-Canela M. (2006). Estimación del riesgo
de contaminar mantos freáticos por infiltración de herbicida atrazina en distritos de riego en México.
Tesis de Maestría. Universidad Nacional Autónoma de México, Morelos, México. 128 p.
SAGARPA (2007). Anteproyecto de Modificación de la Norma Oficial Mexicana NOM-032-FITO-1995. Por la
que se establecen los requisitos y especificaciones fitosanitarios para la realización de estudios de
efectividad biológica de plaguicidas agrícolas y su dictamen técnico.Secretaría de Agricultura,
Ganadería, Desarrollo Rural, Pesca y Alimentación.[en línea]. http://www.cofemermir.gob.mx/mir/
uploadtests/10046.59.59.1.NOM-%20032-FITO1995%20(MODIF.).doc.(Último acceso mayo 13, 2013).
Schoen S. R. y Winterlin E. P. 1987. The effects of various soil factors and amendments on the degradation
of pesticides mixtures Journal Environmental Science Health, B22:347-377.
Tyess, D.L., Shea, P.J., Parkhurst, A.M., 2006, Mineralization potential of atrazine and degradation
intermediates from clustered characteristics in inoculated soils. Soil Sediment Contam. 15, 87-102.
UNEP (2002). Central America and the Caribbean Regional Report, Regionally Based Assessment of Persistent Toxic Substances. United Nations Environment Programme. Ginebra, Suiza. 133 p.
Wang X., Guo X., Yang Y., Tao S. y Xing B. (2011). Sorption mechanisms of phenanthrene, lindane, and
atrazine with various humic acid fractions from a single soil sample. Environ. Sci. Technol. 45, 21242130.
Worthing, C. R., y M. D. Phill (ed.). 1987. Atrazine, In The pesticide manual, 8th ed. British Crop Protection
Council, Fainham, United Kingdom, p. 36±37.
641
LIST OF COAUTHORS:
Rosa Angélica Guillén Garcés Profesor de tiempo completo
Ingeniería en Tecnología Ambiental
Universidad Politécnica del Estado de Morelos
Paseo Cuauchnáhuac no. 566 col. Lomas del Texcal Jiutepec Mor. C.P.62550
Tel. 777 2293533
Email. [email protected]
Ana Gabriela Ortega Hernández, Egresada de la carrera de Ingeniería en Tecnología
Ambiental
Ingeniería en Tecnología Ambiental
Tel. 777 2293533
Carlos Díaz Gómez Profesor de asignatura
Tel. 777 2293533
642
CHAPTER 6.9. EFECTO DEL 1,2,3-TRICLOROBENCENO EN LA BIODEGRADACIÓN
DE UN SUELO CONTAMINADO CON HEXACLOROCICLOHEXANO EN UN REACTOR
DE SUELOS ACTIVADOS EN AMBIENTE DESNITRIFICANTE
Gonzalo M. Escobedo-Acuña (1); B. Camacho-Pérez (1), E. Ríos-Leal (1);
M. T. Ponce-Noyola (1); J. Barrera-Cortés (1); Héctor M. Poggi-Varaldo* (1);
H. B. C. Poggi-Rinderknecht (2); N. Rinderknecht-Seijas (2)
(1) CINVESTAV- IPN, México DF, México;
(2) ESIQIE-IPN, México DF, México
RESUMEN
(OȖ-hexaclorociclohexano (lindano) es uno de los plaguicidas que más se han utilizado en
el mundo. Es un compuesto altamente halogenado, tóxico y persistente en el medio
ambiente. Durante la ruta de degradación anaeróbica del lindano, se producen diferentes
metabolitos intermediarios, uno de ellos es el 1,2,3-Triclorobenceno (TCB), el cual ejerce
un efecto sobre el desempeño del proceso de biorremediación de un suelo contaminado
con lindano y obteniendo así una remoción pobre del lindano. Existe poca información
referente al efecto que los metabolitos intermediarios en la ruta de degradación de un
compuesto, puedan tener sobre la biodegradación del compuesto madre así como la falta
de métodos para obtener dicha información.
Los objetivos de este trabajo fueron (i) determinar las velocidades iniciales de
degradación del lindano a diferentes concentraciones, (ii) determinar las constantes
cinéticas de degradación (Qi, Ks y Ki) y (iii) determinar qué efecto ejerce el TCB en la
biodegradación del lindano en un RSA con ambiente desnitrificante que trata un suelo rico
en materia orgánica y contenido de arcilla (suelo pesado).
Se obtuvieron las velocidades iniciales de degradación (Qi), posteriormente las
constantes Ks y Ki, determinando que el efecto que ejerce el TCB sobre el consorcio
microbiano lindanoclástico es un efecto de inhibición acompetitiva.
Palabras Clave: constantes cinéticas, inhibición, metabolito intermediario, anaerobiosis.
------------------------------------------*Autor de correspondencia
643
Introducción
El hexaclorociclohexano (HCH) es un compuesto altamente clorado, que se sintetiza a
partir de la fotocloración del benceno obteniéndose teóricamente 8 estereoisómeros, los
predominantes en el HCH técnico sonĮ-ȕ-Ȗ-, y į-isómero. El Ȗ-isómero es el único con
propiedades insecticidas, también llamado lindano. El lindano es un insecticida orgánico,
ampliamente usado a nivel mundial, particularmente en México. Se ha usado
principalmente en la agricultura y salud pública (Elango at al., 2011; Guillen-Jiménez et
al., 2012; INE, 2004; Lal et al. 2010).
El lindano es un compuesto altamente halogenado, tóxico y persistente en el medio
ambiente. Durante la ruta de degradación anaeróbica del lindano, se han identificado
diferentes metabolitos intermediarios generados gracias a las reacciones de decloración.
Uno de ellos es el 1,2,3-Triclorobenceno (TCB) el cual es un compuesto organoclorado,
liposoluble, presenta alta volatilidad, es ambientalmente estable y resistente a
biodegradación, por lo que se le considera como un contaminante prioritario. Se sabe que
la ruta de degradación es llevada acabo por enzimas clave responsables de las
reacciones de dehidrocloración, dicloración, dehalogenación y decloración reductiva,
hasta llegar al producto final, clorobenceno y diclorobenceno. (Camacho-Pérez et al.,
2012;, et al., 2010; Quintero et al., 2005; Zhao et al. 2014).Debido a lo anterior, la cinética
de degradación anaeróbica del lindano se puede analizar como una cinética enzimática,
ya que ademásambas describen una trayectoria hiperbólica, y así determinar el efecto de
los metabolitos intermediarios.
Por otro lado, la biorremediación constituye un técnica adecuada para el tratamiento
de suelos contaminados con lindano debido a sus ventajas sobre otras alternativas tales
como incineración, almacenamiento o lavado de suelo (Eweis 1998; Varo-Arguello et al.,
2012). Una de las técnicas más eficaces e importantes ad situ y ex situ en el tratamiento
de suelos y sedimentoscontaminados con xenobióticos recalcitrantes, específicamente
suelos de textura arcillosa y estratificados, con baja conductividad hidráulica, baja
permeabilidad y alto contenido de materia orgánica (Suelos Pesados) son reactores de
suelos activados (RSA) (Bachmannet al., 1998; Robles-González et al., 2008). Se han
estudiado diferentes aceptores de electrones (SO4=, NO3-, CO2) en los RSA con
diferentes rendimientos (Tabla 1) (Robles-Gonzáles et al., 2006; Robles-González et al.,
2012). En el presente trabajo se estudia la biodegradación del lindano en un RSA con
NO3- como aceptor de electrones.
Hasta nuestro conocimiento actualmente existe poca información referente al efecto
que los metabolitos intermediarios en la ruta de degradación de un compuesto, puedan
tener sobre la biodegradación del compuesto madre así como la falta de métodos para
obtener dicha información (Segel, 1979).
Los objetivos de este trabajo fueron (i) determinar las velocidades iniciales de
degradación del lindano a diferentes concentraciones,(ii) determinar las constantes
cinéticas de degradación (Qi, Ks y Ki) y (iii) determinar qué efecto ejerce el TCB en la
biodegradación del lindano en un RSA con ambiente desnitrificante que trata un suelo
pesado.
644
Materiales y Métodos
Suelo. El suelo usado en los experimentos fue colectado de San Miguel Tequixtepec,
Oaxaca. Con características fisicoquímicas como un pH de 7.2, un alto contenido de
materia orgánica (8.1%). Contiene 37, 21 y 42% de arena, limo y arcilla respectivamente,
de textura arcillosa. Es un suelo de tipo Cambisol con una conductividad hidráulica baja.
(Varo-Arguello et al. 2012).
TABLA 1. Características fisicoquímicas del Lindano.
Características del Lindano
Masa Molecular (g·mol-1)
290.83
-3
Densidad (g·cm )
1.87
Punto de ebullición (ºC)
112.8
Punto de fusión (ºC)
323.4
Solubilidad a 20º C (mg·L-1) 7.3 ± 7.8
Benceno (g·L-1)
289
-1
Acetona (g·L )
435
Constante de Henry
2.92x10-6
(atm·m3·mol-1)
Presión de vapor (atm)
7.33x10-8
Coeficiente de partición
3.72
(Octanol/agua)
Reactivos Químicos. Se utilizó Ȗ-hexaclorociclohexano (Lindano) grado reactivo de
Sigma-Aldrich (97% de pureza). Las propiedades fisicoquímicas del Lindano se exponen
en la tabla 1. El Lindano es un compuesto organoclorado, moderadamente lipofílico,
caracterizado por un alto coeficiente de partición octanol-agua (Kow=4x103). Baja
solubilidad en agua (7 mgL-1) y de carácter polar bajo debido a la fuerte electronegatividad
de los átomos de Cl unidos al anillo alifático.
El 1,2,3-Triclorobenceno (TCB) grado analítico de Sigma-Aldrich (99.9% de pureza)
fue seleccionado por ser el metabolito intermediario más clorado comercialmente
disponible y debido a sus propiedades fisicoquímicas (Tabla 3) representa bien a los
demás metabolitos intermediarios. El hexano y acetona utilizados fueron grado HPLC.
Preparación y contaminación del suelo. El suelo fue esterilizado por calor seco (3
veces a 121º C durante 60 min. con un periodo de incubación de 24 h). El suelo se
contaminó con una mezcla de diferentes concentraciones de Lindano y TCB (Tabla 2). La
cantidad correspondiente a cada concentración de Lindano y TCB fue disuelta en 500 ml
de acetona grado HPLC y mezclada con el suelo en un rotavapor a 100 rpm durante dos
horas, posteriormente se evaporó la acetona a 56º C. (Camacho-Peréz et al., 2013).
Diseño Experimental. Consistió de una hipermatriz, compuesta de 3 matrices (I, II y III),
cada una con concentraciones diferentes de Lindano (50, 100 y 180 mg/Kg) y 4
645
concentraciones distintas de TCB (0, 32.5, 62.5 y 120 mg/Kg) (Tabla 3).
Se realizó un muestreo compuesto para mayor homogeneidad en las muestras y se
muestreó los días: 0, 2, 3, 5, 6 y 7.
Reactores de suelo activado en ambiente desnitrificante. Las unidades
experimentales consistieron en botellas serológicas de 125 ml, a las cuales se les agregó
20 g de suelo estéril contaminado de acuerdo a su matriz. Posteriormente se les agregó
60 mL de medio mineral de composición: 300 mg/L NaNO3, 100 mg/L K2HPO4, 0.5 mg/L
NHCO3, 5 mg/L CaCl2 (Modificado de Bachmannet al., 1989) y 10 mL de una solución de
resazurina (1mg/L) y 500 mg SSV L-1 de inóculo DN aclimatado al lindano.
Las unidades experimentales se dividieron en:
1.Unidad Experimental Activa (UEA): suelo estéril contaminado, inóculo aclimatado al
lindano y ambiente desnitrificante sin esterilizar.
2.Control Abiótico (CA): Suelo estéril contaminado, inóculo aclimatado al lindano y
ambiente desnitrificante esterilizado, además de esterilización química mediante
acida de sodio.
Las cinéticas realizadas consistieron en 16 UEA y 16 CA por cada matriz. Donde una
muestra compuesta se consistió en dos unidades experimentales en las mismas
condiciones.
Marco Teórico Para esta actividad, la cinética microbiana de degradación del lindano por
el consorcio microbiano previamente aclimatado al lindano en condiciones desnitrificantes,
fue asimilado, como una analogía a un cinética enzimática, lo cual es una práctica común
en modelos de tratamiento de aguas (Grady et al., 2011).
A pesar del hecho que estamos trabajando un cinética microbiana en lugar de una
cinética enzimática, es posible hacer dicha analogía debido a que ambas cinéticas tienen
el mismo patrón hiperbólico. Además la ruta metabólica de degradación del HCH es
conducida por enzimas dominantes dentro de un consorcio microbiano (Bushan et al.,
2000; Camacho-Pérez et al., 2012; Lodha et al., 2007).
TABLA 2. Propiedades fisicoquímicas del TCB.
Características del 1,2,3-Triclorobenceno
Masa Molecular (g·mol-1)
Densidad (g·cm-3)
Punto de ebullición (ºC)
Punto de fusión (ºC)
Presión de vapor (Pa a 25º C)
Coeficiente de partición
(Octanol/agua)
181.45
1.45
218.5
53.5
17.3
4.05
646
En la práctica de tratamiento de aguas residuales, es un práctica común construir
cinéticas microbianas de inhibición por analogía a los cuatro casos más comunes de
cinética enzimática (Grady et al., 2011).
Con esto en consideración, el coeficiente de la tasa específica de consumo de sustrato
Qi reemplaza al coeficiente de velocidad inicial máxima Vi, mientras que Ks y Ki reemplaza
al coeficiente de afinidad Km y al coeficiente de inhibición Kirespectivamente.
En primer lugar existe una fuerte analogía entre las cinéticas microbiana y enzimática,
ambas son hiperbólicas en la concentración del sustrato.
En segundo, el metabolismo de un consorcio microbiano está dominado por una o unas
enzimas clave con cinéticas hiperbólicas, por lo tanto el metabolismo microbiano puede
manejarse con una analogía hacia una cinética enzimática.
Adicionalmente, la presencia de un compuesto tóxico en una cinética microbiana es un
análogo de un inhibidor enzimático (Lodhaet al., 2007; Grady et al., 2011 and 1996; VaroArguello et al., 2012).
Por lo anterior podemos asumir los principios básicos de los 4 tipos más comunes de
inhibición enzimáticas y cómo vamos a medir el efecto que tienen los metabolitos
intermediarios selectos en la tasa de biodegradación del lindano. (Cappiello et al., 2014;
Segel, 1976).
Métodos Analíticos. La concentración del lindano fue analizada por Cromatografía de
gases usando volatilización al espacio gaseoso (Headspace) en un cromatógrafo de
gases Perkin-Elmer equipado con detector de captura de electrones.
TABLA 3. Hipermatriz de concentraciones de Lindano y TCB.
MATRIZ I
CINETICA Lindano
TCB
(mg/Kg)
(mg/Kg)
A
50
120
B
50
62.5
C
50
31.5
D
50
0
MATRIZ II
MATRIZ III
Lindano
TCB
Lindano
TCB
(mg/Kg) (mg/Kg) (mg/Kg) (mg/Kg)
100
120
180
120
100
62.5
180
62.5
100
31.5
180
31.5
100
0
180
0
El procedimiento de extracción de los residuos de Lindano en el reactor de suelos
activados fue realizado de acuerdo con el método reportado por Quintero et al. 2005. La
concentración de TCB fue analizada en un cromatógrafo de gases acoplado a masas
Agilent Technologies con un automuestreador Gerstel(MPS-2 Twister), la temperatura del
horno fue programada de la siguiente manera: tiempo de retención 40º C, 2 min., una
rampa de 3º C/min hasta 180º C; una rampa de 8º C/min hasta 270º C. El volumen de
inyección fue de 1 ȝOYtD 6SOLW-less a 280º C. Se utilizó Helio como gas acarreador a un
flujo de 1.0 ml/min.
El pH del suelo fue determinado en una solución agua desionizada/lodo del RSA 1:2
(w/w). La textura fue medida por el método hidrométrico, la demanda bioquímica de
oxígeno (DQO) fue estimado de acuerdo al método 507 y el contenido de materia
orgánica fue estimado por el método de oxidación con K2Cr2O7.
647
Resultados y Discusiones
De acuerdo a los resultados obtenidos para las velocidades iniciales (Tabla 4) se observa
como al aumentar la concentración de TCB la velocidad de degradación del lindano
disminuye en todas las matrices, siendo la cinética A en las tres matrices la que tiene el
desempeño más pobre, al ser la que contiene la concentración más alta de TCB (120
mg/kg).
En la matriz III (180 mg/Kg de Lindano) las velocidades de degradación fueron las
mínimas, esto puede deberse tanto al TCB como a la concentración de lindano, ya que
recordemos que el lindano (análogo del sustrato) es un compuesto tóxico persistente y a
cierta concentración inhibe actividad de degradación, lo que conlleva a estrés y la muerte
del consorcio microbiano. Hasta nuestro conocimiento no se ha estudiado dicha
concentración en estos sistemas (RSA) y con la cantidad de inóculo utilizado, que es de
500 mg SSV, por lo que podemos confirmar que nuestro consorcio microbiano
lindanoclástico soporta concentraciones superior a las reportadas para RSA, sin embargo
la actividad de biodegradación es mínima.
TABLA 4. Velocidades iniciales de degradación de lindano.
CINÉTICA
MATRIZ I
Qi
R2
MATRIZ II
Qi
R2
MATRIZ III
Qi
R2
A
3.20
0.90
2.42
0.73
2.65
0.88
B
3.67
0.90
2.73
0.81
2.48
0.72
C
4.34
0.81
2.91
0.74
2.74
0.93
D
5.43
0.99
4.65
0.84
3.23
0.71
Matriz I: Concentración de Lindano 50 mg/Kg; Matriz II: Concentración de
Lindano 100 mg/Kg; Matriz III: Concentración de Lindano 180 mg/Kg
A: TCB 120 mg/Kg; B: TCB 65 mg/Kg; C: TCB 32.5 mg/Kg; D: TCB 0 mg/Kg
Qi: velocidad inicial de degradación del lindano.
TABLA 5. Recíprocos (Qi y S) para gráfica Lineweaver-Burk.
A
B
C
D
0.3125 0.02 0.2725 0.02 0.2304 0.02 0.1842 0.02
0.4132 0.01 0.3663 0.01 0.3436 0.01 0.2151 0.01
0.3774 0.0056 0.4032 0.0056 0.3650 0.0056 0.3096 0.0056
Los datos obtenidos al ser graficados presentan una tendencia hiperbólica o de
Monod, (con la excepción de los valores para la cinética A) con lo cual una vez obtenido el
parámetro Qi (degradación de mg de lindano / Kg suelo / 500 mg SSV de biomasa / día)
se realizó una gráfica recíproca de Lineweaver-Burk (1/Qi vs1/s). La figura 1 muestra
como las curvas recíprocas se corresponden paralelas, lo que significa que, de acuerdo
con el método gráfico seleccionado el efecto de inhibición del TCB es acompetitivo,
también mediante este método fue posible determinar las constantes de afinidad (Ks)y la
648
constante de inhibición(Ki).Las cuales fueron:23.61mg/Kg y 5.31 mg/Kg respectivamente.
Como se puede apreciar en la tabla 6 la constante de inhibición se corresponde con la
concentración del TCB, es decir van aumentando en la misma proporción.
TABLA 6. Parámetros cinéticos de la degradación
microbiana.
Cinética
A
Ecuación de la recta
Y= -5.4585x+0.4325
R2
Qi
0.62249
2.31
2.19
2.34
3.05
B
C
Y= -9.1314x+0.4557
Y= -9.6829x+0.4279
0.99936
0.97623
D
Y= -7.7472x+0.3282
0.7652
Ks
Ki
12.62 19.85
20.04 10.99
22.63 5.31
23.61 0.00
El efecto que ejerce el TCB sobre la remediación del Lindano en un RSA-DN es
inhibición acompetitiva, gráficamente se puede observar las curvas de las cinéticas de
manera paralela en la gráfica Lineweaver-Burk (Gráfica 1). El TCB se une al complejo
Consorcio-Lindano inactivándolo y de esta manera los microorganismos lindanoclástico
reducen su actividad de biodegradación, ya que el consorcio microbiano tiene diferente
afinidad tanto por el TCB como por el Lindano y parte de éste se desvía a degradar al
TCB. El efecto de inhibición competitivaes especialmente común en sistemas
multienzimáticos,como es el caso del sistemaestudiado en el presente trabajo (Segel,
1975; Wang et al., 2013).
FIGURA 1.Lineweaver-Burk de cinéticasmicrobianas de degradación de lindano.
Se ha reportado en microorganismos aerobios que la ruta de degradación del lindano
se puede llevar hasta Acetil-CoA y Succinil-CoA que posteriormente entra al ciclo TCA,
siendo el lindano únicamente la fuente de carbón, para esto se requiere una maquinaria
enzimática compleja, que hasta nuestro conocimiento no ha sido reportada para
microorganismos anaerobios. Debido al efecto que ejerce el TCB sobre la degradación del
lindano, el producto final reportado por distintos autores en sistemas anaerobios es el
Clorobenceno y Triclorobenceno. Por lo que esto representa un cuello de botella en la
biodegradación del Lindano (Camacho-Pérez et al., 2012).
649
Conclusiones
Se determinó que el efecto que ejercen el metabolito intermediario es Acompetitivo, lo que
provoca que el consorcio microbiano no biodegrade adecuadamente el lindano y tenga un
desempeño pobre en el RSA-DN.
El ambiente desnitrificante del RSA se mantuvo durante toda la cinética de
degradación. Aunque los RSA son una de las mejores alternativas in situ y ad situ para el
tratamiento de suelos pesados fuertemente contaminados, el nitrato como aceptor de
electrones no favorece la biodegradación del lindano, puede deberse tanto al efecto de los
metabolitos intermediarios como al ambiente desnitrificante, lo que quizá interfiera con el
correcto funcionamiento de las enzimas involucradas tanto en la dehalogenación
reductiva. Por lo cual resulta necesario realizar estudios enzimáticos en dicho ambiente
anaerobio.
Uno de los criterios importantes de las cinéticas enzimáticas es que el inhibidor
siempre debe estar presente en durante ésta. Lo que se observó es que el TCB siempre
estuvo presente durante la cinética de degradación del lindano, además su concentración
en el medio fluctúa, debido a que existe degradación y producción por parte consorcio
microbiano lindanoclástico.
El tiempo de adsorción abiótica variaba dependiendo la concentración del Lindano y
TCB sin embargo en promedio fue de 3 días, al realizar el cuarto muestreo éste ya no
presentaba remoción abiótica en los CA.
Referencias
Bachmann A, De Bruin W, Jumelet JC, Rijnaarts HH, Zehnder AJ (1988): Aerobic biomineralization of alphahexachlorocyclohexanein contaminated soil. Appl Environ Microbiol, 54(2):548-554.
Bachmann A, Walet P, Wijnen P, De Bruin W, Huntjens JL, Roelofsen W, Zehnder AJ (1988):
Biodegradation of alpha and beta-hexachlorocyclohexane in a soil slurry under different redox
conditions. Appl Environ Microbiol, 54(1):143-149.
Bhushan B, Chauhan A, Samanta S, Jain R (2000): Kinetics of Biodegradation of p-Nitrophenol by Different
Bacteria. Biochemical and Biophysical Research Communications 274, 626±630.
Camacho-Pérez B., Ríos-Leal E, Rinderknecht-Seijas N, Poggi-Varaldo HM. (2012) Enzymes involved in the
biodegradation of hexachlorocyclohexane: A mini review., Journal of Environmental Managemen
2011.06.047.
Camacho-Pérez B., Ríos-Leal E, Rinderknecht-Seijas N, Poggi-Varaldo HM (2011): Desorption
Enhancement and Biostimulation in the Bioremediation of a Heavy Agricultural Soil Polluted with
Lindane in Sequential Slurry Bioreactors. Bioremediation and Sustainable Environmental
Technologies²2011. International Symposium on Bioremediation and Sustainable Environmental
Technologies. ISBN 978-0-9819730-4-3.
Camacho-Pérez B.; Ríos-Leal E.; Solorza-Feria O.; Vazquez-Landaverde P.; Barrera-Cortés J.; PonceNoyola M.; Garcia-Mena J.; Rinderknecht-Seijas N.; Poggi-Varaldo H.M. (2013). Performance of
anelectrobiochemical slurry reactor for the treatment of a soil contaminated with lindane. Journal of
New Materials for Electrochemical Systems 16, 217-228.
Cappiello M, Moschini R, Balestri F, Mura U, Del-Corso A (2014): Basic models for differential inhibition of
enzymes. Biochemical and Biophysical Research Communications 445, 556±560.
Elango V; Kurtz H, Andersonc C; Freedman D (2011): Use of -hexachlorocyclohexane as a terminal
electron acceptor by an anaerobic enrichment culture. Journal of Hazardous Materials 197 (2011)
204±210.
650
Eweis JB, Ergas SJ, Chang DPY, Schroeder ED (1998): Bioremediation Principles McGraw-Hill. USA; 207236.
Grady Jr, C.P. L.; Barth F.; Barbeau, S. (1996). Variability in kinetic parameter estimates: a review of
possible causes and a proposed terminology. Water Res.30 (3), 742-748.
Grady Jr. C.P.L.; Daigger, G.T.; Love, N. G.; Filipe, C.D.M. (2011). Biological wastewater treatment. 3rd.
edition IWA Publ. and CRC Press, Boca Raton, FL. ISBN-13: 978-0849396793
Guillen-Jiménez F, Cristiani-Urbina E, Cancino-Díaz J, Flores-Moreno J, Barragán-Huerta E (2012): Lindane
biodegradationbythe Fusarium verticillioides AT-100 strain, isolatedfrom Agave tequilanaleaves:
Kineticstudy and identi௱cation of metabolites. International Biodeterioration&Biodegradation 74, 36-47.
Lal R; Pandey G; Sharma R; Kumari K; Malhotra S; Pandey R; Raina V; Kohler H; Holliger C; Jackson C;
Oakeshott J (2010): Biochemistry of MicrobialDegradation of Hexachlorocyclohexane and
ProspectsforBioremediation. Microbiology and Molecular BiologyReviews, Mar. 2010, p. 58±80 Vol. 74,
No. 1.
Lodha B.; Bhat P.; Suresh K.; Vaidya N.; Mudliar S.; Killedar D.; Chakrabarti T.; (2007). Bioisomerization
kinetics of gamma-HCH and biokinetics of Pseudomonas aeruginosa degrading technical HCH.
Biochem. Eng. J. 35, 12±19.
4XLQWHUR -& 07 0RUHLUD * )HLMRR DQG -0 /HPD ³$QDHURELF GHJUDGDWLRQ RI
hexachlorocyclohexane isomers in liquid and VRLOVOXUU\V\VWHPVĆhemosphere. 61:528-536.
Robles-González I, Fava F, Héctor M Poggi-Varaldo HM (2008): A review on slurry bioreactors for
bioremediation of soils and sediments. Microbial Cell Factories 2008, 7:5; doi:10.1186/1475-2859-7-5.
Robles-González I, Ríos-Leal E, Sastre-Conde I, Fava F, Rinderknecht-Seijasd N, Poggi-Varaldo HM (2012):
Slurry bioreactors with simultaneous electron acceptors for bioremediation of an agricultural soil
polluted with lindane. Process Biochemistry 47, 1640±1648.
Robles-González I, Ríos-Leal E, Sastre-Conde I, Ferrera-Cerrato R, Caffarel-Méndez S, Poggi-Varaldo HM
(2008): Effect of supplementary organic carbon source and electron acceptors on the removal of 2,4Dichlorophenoxyacetic acid in slurry bioreactors. In Proceedings 2nd International Conference on
Remediation of Contaminated Sediments, Venice, September 30 2003; Venice, Italy Edited by: Pellei M,
Porta A, Hinchee E. Battelle Press, Columbus, OH; 2003. Paper G-08, ISBN. 1-57477-143-4.
Segel I (1975): Biochemical calculations ± How to solve mathematical problems in biochemistry. John Wiley
& Sons Inc. 2nd. edition, New York, USA.
Varo-Arguello W, Camacho-Pérez B, Ríos-Leal E, Poggi-Varaldo HM (2012): Aerobic Degradation of
Lindane in Clay Soil with High Organic Matter Content in a Triphasic SlurryBioreactor. Remediation of
Chlorinated and Recalcitrant Compounds²2012. Eighth International Conference on Remediation of
Chlorinated and Recalcitrant Compounds (Monterey, CA; May 2012). ISBN 978-0-9819730-5-0,
©2012 Battelle Memorial Institute, Columbus, OH, www.battelle.org/chlorcon.
Varo-Arguello W, Camacho-Pérez B, Ríos-Leal E, Vazquez-Landaverde PA, Ponce-Noyola MT, BarreraCortés J, Sastre-Conde I, Rindernknecht-Seijas NF, Poggi-Varaldo HM (2012): Triphasic slurry
bioreactors for the bioremediation of lindane-impacted soil under aerobic and anaerobic conditions.
Environmental Engineering and Management Journal Vol. 11, No. 10, 1811-1823.
Wang G, Post W (2013): A note on the reverse Michaelis-Menten kinetics. Soil Biology & Biochemistry 57,
946-949.
Zhao X.; Li J.; Sun X.; Yao Z. (2014): Remediation of 1,2,3-trichlorobenzene contaminated soil using a
combined thermal desorption±molten salt oxidation reactor system. Chemosphere 97;125±129.
651
LIST OF COAUTHORS:
Gonzalo Manuel Escobedo-Acuña, M. Sc. Candidate, presenting author. Same address
and e-mail as those of Dr. H.M. Poggi-Varaldo.
Beni Camacho-Pérez, Sc. D. Candidate. Same address as Dr. H.M. Poggi-Varaldo. Email: [email protected]
Elvira Ríos-Leal, full Profesor. Same address as Dr. H.M. Poggi-Varaldo. E-mail:
[email protected]
María Teresa Ponce-Noyola, Sc. D., Full Professor, Same address as Dr. H.M. PoggiVaraldo. E-mail: [email protected]
Josefina Barrera-Contreras, Sc. D., Full Professor, Same address as Dr. H.M. PoggiVaraldo. E-mail: [email protected]
Héctor B. Cuauhtémoc Poggi-Rinderknecht, B. Sc.;
Noemí Rinderknecht-Seijas, Full Professor.
ESIQIE-IPN, Division of Basic Sciences, Campus Zacatenco, México D.F., Mexico. E-mail:
[email protected]
Héctor M. Poggi-Varaldo, Sc D, Full Professor. Author to whom all correspondence
should be addresed
CINVESTAV, Dept. Biotechnology and Bioengineering, EnvironmentalBiotechnology and
RenewableEnergies R&D Group; P.O. Box 14-740, México DF, 07000, México.
Tel: 5255 5747 3800 ext 4324, Fax. 5255 5747 3313,
652
CHAPTER 6.10. COMPARISON OF METHODS FOR FITTING KINETIC DATA OF
INHIBITION OF THE BIODEGRADATION OF A MAIN POLLUTANT FOR
REMEDIATION ASSESSMENT
Gonzalo M. Escobedo-Acuña (1); B. Camacho-Pérez (1), E. Ríos-Leal (1);
M. T. Ponce-Noyola (1); J. Barrera-Cortés (1); Héctor M. Poggi-Varaldo* (1);
H. B. C. Poggi-Rinderknecht (2); N. Rinderknecht-Seijas (2)
(1) CINVESTAV- IPN, México DF, México;
(2) ESIQIE-IPN, México DF, México;
ABSTRACT
The Ȗ-hexachlorociclohexane (also called lindane) has been one of the most used
pesticides worldwide. It is a highly halogenated compound, toxic and persistent in the
environment. During its anaerobic biodegradation many intermediate metabolites are
formed; some of them are more resistant to be degraded by the microbial population and
can influence or inhibit the bioremediation process performance. So far, there is little
information on the kinetic impact of intermediate metabolites on the degradation of the
parent pesticide (likely inhibition) as well as scarce information on the likelihood and
criteria selection of fitting methods for determining the kinetic coefficients. Kinetic
parameter values, in turn, could significantly influence the time frame of a bioremediation
project. Therefore we assessed the effectiveness and accuracy of several methods for
fitting kinetic data assuming a scenario of competitive inhibition with one inhibitor
(intermediate metabolite) and a parent pesticide (substrate). We found that no single
method showed the minimum bias in all kinetic parameters. Considering the scenario of
only the dependent variable with error a first compromise solution is to choose the fitting
based on Levenberg-0DUTXDUG 63 µD¶ ZKHUHDV WKH VHFRQGPHWKRGRI FKRLFH ZRXOGEH
the weighed linear regression with linearized data Lineweaver-Burk (LB-:/5µD¶,QRUGHU
to improve the estimation of the 3 parameters, the use of two methods is recommended as
a second trade-RII :LWK WKH PHWKRG 63 µD¶ JRRG HVWLPDWLRQV RI ERWK qmax and Ks are
obtained, whereas the fitting with method LB-:/5 µD¶ ZLOO JLYH WKH EHVW HVWLPDWLRQ RI Ki.
The batch method Batch LR-2Var, in spite of the advantage of smaller experimental work,
leads to large biases of Ks and Ki. The weighed regression based on Batch was worse.
This is unfortunate because from the experimental point of view the batch method required
significant less experimental work than the other methods.
To the best of our knowledge, this is the first comprehensive, critical evaluation of
fitting methods to a microbial inhibition kinetics, based on a synthetic experiment (in silico).
With caution due to that, our results reveal that some assumed advantages of weighed
linear regression and less biased Levenberg-Marquardt based software were not
confirmed, particularly when both variables were subject to error, because the fitting
methods did not produce more accurate estimations of the kinetic parameters.
653
Introduction
Halogenated organic substances constitute one of the most important groups of
environmental pollutants as a result of their widespread use as herbicides, insecticides,
fungicides, solvents, hydraulic and heat transfer fluids, plasticizers, and intermediates for
chemical syntheses. Because of their toxicity, bioconcentration, persistence, and ubiquity,
the halogenated compounds have raised concern over the possible effects on the quality
of life (Camacho-Pérez et al., 2012).
7KH Ȗ- KH[DFKORURF\FORKH[DQH Ȗ-HCH; also called lindane) is a highly halogenated
organic pesticide that has been used worldwide, particularly in Mexico, in spite of its
banning in first world countries. Lindane has been used for crop protection and prevention
of vector-borne diseases for many decades. Negative impacts of lindane on the
environment and human health have been reported worldwide. Due to their
hydrophobicity, lindane is tightly bound to the organic matter and clay of soils. This, in turn,
decreases their bioavailability (Camacho-Pérez et al., 2013).
Bioremediation constitutes a feasible approach to clean up soils and water systems
FRQWDPLQDWHG E\ Ȗ-HCH because of its advantages over other alternatives such as
incineration, storage or soil washing (Varo-Arguello et al., 2012). Treatment of soils and
sediments in SBs has become one of the best options for bioremediation of soils polluted
by recalcitrant pollutants under controlled environmental conditions. In fact, under slurry
conditions, the pollutant depletion rates depend mainly on the degradation activity of the
microorganisms available in the system, and the results obtained generally reflect the
actual biological depuration potential of the soil (Robles-González et. al, 2012). Is well
known that the anaerobic degradation of lindane is attractive, since reductive
dehalogenation pathways can be very effective for the dechlorination of the pollutant
(Camacho-Pérez et. al, 2011).
Nevertheless during the pathway biodegradation of lindane some intermediate
metabolites (IM) are produced such as pentachlorocyclohexane isomers (PCCH),
tetrachlocyclohexane (TCCH), 1,2,3-trichlorobenzene (1,2,3-TCB), 1,3-dichlorobenzene
(1,3-DCB), and chlorobenzene (CB) (Camacho-Pérez et al., 2012). Due to the biological
depuration potential of each system of bioremediation, the IM can represent a barrier to
achieve better efficiencies in the bioremediation process. In fact, the toxicity and the poor
biodegradability exhibited by these compounds could be bottlenecks in the application of
microbial degradation of these compounds because they could act as inhibitors. Also this
IM could represent another environment risk due to a major chemical stability than the
parent pesticide (Guiren, 2008; Robles-González et al., 2012).
So, inhibition scenarios in the degradation of lindane can be expected. For a given
microbial kinetic inhibition model, there are several methods to fit the experimental data in
order to determine values for qmax, Ks, and Ki. From the statistical theory, fitting methods
lead to errors and bias on the kinetic parameters (Spector et al., 1981). However, to the
best of our knowledge, there is no critical comparison of the most common fitting
procedures and their associated uncertainties of inhibition models. Thus, the selection of
the method remains a matter of preference or software availability rather than a sound
statistical criterium.
We have developed a synthetic set of experimental data with errors. Since there are at
least four simple inhibition models (Competitive-like, Uncompetitive-like, Noncompetitivelike, Mixed type-like; Grady et al., 1996; Segel, 1976), and each model could be fitted by
several methods, the complete task is overwhelming. Thus, we have chosen the
654
competitive inhibition scenario for evaluating the likelihood of fitting procedures. So, the
conclusions of this work will be valid only for competitive inhibition kinetics; in future
contributions the likelihood of fitting procedures to uncompetitive, noncompetitive, and
mixed type inhibition kinetics models will be assessed
So far, there is little information on the likelihood and criteria selection of fitting
methods for determining the kinetic coefficients. In this work we assessed the
effectiveness and accuracy of five methods for fitting kinetic data assuming a scenario of
competitive inhibition with one inhibitor (intermediate metabolite) and a parent pesticide
(substrate). Furthermore, we discuss the advantages and disadvantages of each fitting
method.
Methods
Theoretical background. In this study, the degradation microbial kinetics of a substrate
(the main pollutant) by the whole microbial population was assimilated, as an analogy, to
an enzymatic kinetics as it is a common practice in wastewater treatment models (Grady
et al., 2011). Despite the fact that we are treating with a microbial kinetic instead of an
enzymatic kinetic, it is possible because both kinetics has the same hyperbolic pattern.
Also the organochlorine pesticide degradation is conducted in a metabolic pathway that is
leading by dominant enzymes in the microbial population (Camacho-Pérez et al., 2012;
Lodha et al., 2007).
In the practice of wastewater treatment, it has been very common to build microbial
inhibition kinetics by analogy to the four cases of enzymatic kinetics (Grady et al., 2011). In
this regard, the specific substrate uptake coefficient qmax replaces the maximum rate of
enzymatic substrate consumption Vmax, whereas Ks and Ki replace the affinity coefficient
Km and the inhibition coefficient Ki, respectively. First, there is a strong analogy between
microbial Monod kinetics and enzymatic Michaelis-Menten for simple cases; both are
hyperbolic in substrate concentration. Second, a complex microbial metabolism could be
dominated by a key or a couple of key enzymes with hyperbolic kinetics, so it is not a
surprise that the microbial metabolism could be approximated itself by an analogical
kinetics. Furthermore, the presence of a toxic compound in a microbial kinetic is the
analogue of an enzymatic inhibitor (Bharat et al., 2000; Grady et al., 2011 and 1996;
Varo-Arguello et al., 2012).
We can assume the basic principles of a competitive inhibition and how we are going
to measure the effect of intermediate metabolites in the biodegradation rate of lindane.
(Segel,1976). Five methods of fittings were tested, namely Lineweaver-Burk simple linear
regression (LB), Lineaweaver-Burk weighed linear regression (LB-WLR); Batch (Batch LR2Var), Batch weighed linear regression 2 variables (Batch WLR-2V), and Software
GraphPad Prism 5.0 (SP) based on Levenberg-Marquardt algorithm.
Kinetic data. We worked with synthetic data based on a known kinetic model modified by
low-to-moderate noise, that is, we departed from known values of qmax, Ks, and Ki. The
obtained values of the maximum specific uptake rate of substrate, affinity and inhibition
coefficients (qmax, Ks, and Ki, respectively) were compared with the true values for analysis
of accuracy and precision.
Lineweaver-Burk linear regressions (LB). It is well known that in order to facilitate the
determination of the kinetic constants by linear regression of a hyperbolic model, the data
655
should be previously linearized following the method by Lineweaver-Burk (Segel, 1976), in
this form:
(1)
where:
;
; q0 = initial rate of substrate consumption; qmax = maximum rate
of substrate consumption; So,p = initial concentration of substrate (or main pollutant), m
levels 1 p m; [I]o,j LQLWLDOFRQFHQWUDWLRQRILQKLELWRUZLWKµQ¶OHYHOV j n.
It is worth highlighting that the experimental data required to fit the linearized model Eq.
LVDVHWRIµP¶PDWULFHV)RUHDFKYDOXHRI>,@0,j WKHUHLVDPDWUL[RIµQ¶FRXSOHVȖDQGıWKLV
JLYHVDWRWDORIµPQ¶FRXSOHVRIYDOXHV60, q0)
The kinetic parameters were obtained from the regression coefficients after performing the
linear regressions of Eq. 1 and Eq. 5b (with only changes of notation in Eq. 2):
(2)
where:
qmax = 1/b0
(3a);
(4a);
(3b)
Ks,app = Ks*ࢥ
(4b); Ks,app = b1*qmax (4c)
7KH SURFHGXUH FRQVLVWHG RI SHUIRUPLQJ µP¶ UHJUHVVLRQV RQH IRU HDFK YDOXH RI >,@ 0,j.
From the first regression of this set, without inhibitor (control), we obtained the values of
qmax and Ks (Eq. 4b), since ࢥ = 1 in Eq. 3a
Afterwards, we determine the apparent Ks,app for each regression of the first set. With the
VHWRIµP¶YDOXHVKs,app, we perform a second regression Ks,app, j versus [I]0,j by re-arranging
Eq. 4b:
(5a);
where c0 = Ks
(6a);
Ks,app = c1*[Io] + c0
c1 = Ks/Ki
(6b),
(5b)
Ki = co/c1 (6c)
Classical linear regression is based on the assumption that the independent variable is
exact, whereas the dependent variable has an error and this error is represented by a
uniform variance (homoscedasticity) (Kreyszig, 1971). However, in practice, there are
violations to these assumptions that generally lead to parameter estimates whose values
can be quite different to the corresponding true values (bias). So, it is of interest to study
the effect of a non uniform standard deviation in the dependent variable, and the effect of
both variables presenting errors, on the estimated values of the kinetic parameters in this
work.
Therefore, in our work synthetic errors were introduced in both q 0,p and S0,p with the
random number routine of Data Analysis of Excel (Microsoft Office, Microsoft, Seattle WA)
aVVXPLQJ D FRQVWDQW FRHIILFLHQW RI YDULDWLRQ GHFLPDO &RQVHTXHQWO\ Ȗ DQG ı KDG
also errors that were calculated with the well known equations of error propagation
(Kreyszig, 1971).
656
We distinguished two sub-cases here: (i) ILUVWRQO\ȖKDVDQHUURU ZKHUHDVıLVH[DFW
and (ii) VHFRQGERWKȖDQGıKDYHHUURUV
In this way, we will be able to assess not only the influence of the fitting method, but also
the influence of the errors in the experimental variables, on the estimated values of the
kinetic parameters.
Weighed linear regressions of Lineweaver-Burk equation (LB-WLR). It has been
reported that when the dependent variable has a variable error (it is heteroscedastic,
against the requirement of homoscedasticity of linear regression), the estimation of the
regression coefficients is improved if a weight is introduced to each experimental
observation (Kutner et al.7KLVZHLJKWµZ¶LVDPXOWLSO\LQJIDFWRUWKDWFRQVLVWVRIWKH
inverse of the variance of the experimental value (Kutner et al., 2004), So, an experimental
datum with high standard deviation will have a very low weight, and consequently, the
uncertain datum will have a lower influence in the estimation of the regression coefficients
and the derived kinetic parameters.
The regression coefficients can be found by solving the system of linear equations (by
minimization of Qw, that is, the sum of the weighed of the quadratic differences that is, the
Gauss technique (Kutner et al., 2004).
Qw = 6(wj*(Yj,experimental ± Yj,predicted)2)
(7)
For weighed simple linear regression (WLR-1V), it can be shown that the system is:
A
b
=
c
(8)
where A represents the experimental matrix 2x2, b is the vector of regression coefficients,
2x1; and c is the vector of independent terms, 2x1.
First batch method for the batch degradation kinetics of pesticide with respect to
time (Batch LR-2Var). For this method, 2 runs are necessary; the first one is performed in
absence of any introduced initial inhibitor, whereas the second run is carried out in
presence of a known initial concentration of inhibitor.
The departing batch kinetic equations are:
(9)
(10)
It can be shown that the integrated forms of the Eq. 9 and 10 are:
(11);
(12);
Y = ao + a1 X1 + a2 X2
(11a)
YI = b0 + b1 XI1 + b2 XI2
(12a)
657
(12b)
Please note that the batch kinetics with inhibitor is based on the assumption that the
inhibitor is a recalcitrant compound, therefore its concentration is assumed to remain
essentially constant during the batch run (because of this reason Eq. 10 can be integrated
to give Eq. 12). Incidentally, this could be a weakness of the batch experiment for inhibition
determination.
For fitting the data of the batch method, we only need two sets of values S versus t.
For the normal problem or procedure, the amount of experimental work is much reduced
compared to the set of matrices of the method of LB. The regression coefficients of Eqs.
11a and 12a are calculated by linear bi-variable regressions with Excel. Unfortunately, in
our synthetic experiment, we do not know the values of S as a variable with time. So, we
had to start by solving the integrated Eqs. 11a and 12a for S for given times and for the
true values of qmax, Ks, and Ki, using the bisection or regula falsi method. qmax, Ks, and Ki
(Poggi-Varaldo, 2013, private communication, Salvadori and Baron, 1959) (Fig. 1).
Afterwards, we introduced a synthetic error into S using the Random Number routine of
Data Analysis of Excel, thus obtaining Serr. We built two sets of data Serr vs. t, one without
inhibitor and the second for data with inhibitor.
It can be shown that the kinetic parameters can be calculated from the regression
coefficients of both equations, as follows:
(13a);
(13b);
(13c)
Alternatively, Ki can be calculated as:
(14)
Second batch method with weights (Batch WLR-2V). For this method, the kinetic data
were obtained in a similar way than for the first batch method above, except that we
FRQVLGHUHGWKHZHLJKWµZ¶RIHDFKREVHUYDWLRQVLPLODUO\WRZKDWZDVH[SODLQHGLQWKHFDVH
of LB-WLR before. The minimization of the function Qw of Gauss for a two independent
variables regression with three regression coefficients leads to two linear system of
equations (3x3 each) of the form shown below (Poggi-Varaldo, 2013b, private
communication):
A
a or b
=
k
(15)
658
From one system of equations (with data without inhibitor) we obtain a 0, a1, a2, and
from the second (data with inhibitor) we estimate b0, b1, and b2. Afterwards, we estimate
qmax, Ks, and Ki similarly to Eqs. 11a and 12a in the first batch method.
Software for optimization of kinetic parameters using the Levenberg-Marquardt
algorithm (SP). Kinetic data from the set of matrices S0-q0 for each inhibitor concentration
was entered to a software (GraphPad Prism 5.0) that performs parameter optimization and
fitting data without forcing any linearization. It is advocated that this approach results in
lower bias in the coefficient estimations than linearization of hyperbolic models (Motulsky
et al., 2003).
Please note that methods based on LB, LB-WLR, and SP use experimental data
arranged as a set of matrices. In our synthetic experiment each matrix had 7 pairs (S 0, q0)
and there were 7 matrices because 7 different inhibitor concentrations were set for our
synthetic experiment. On the other hand, the batch method used only two tables S-t, with 7
values of t in each table.
Table 1 shows the compilation of the values of the kinetic parameters and their biases for
the methods tested in this work.
Effect of the error in both dependent and independent variable on parameter
estimates. Regarding the simple linear regression using the equations of LB (comparison
/% µD¶ vs. /% µE¶ 7DEOH WKHUH ZDV QR QRWLFHDEOH LQFUHDVH RI ELDV ZKHQ ERWK YDULDEOHV
were subject to error, as compared to the case with only the dependent variable with error
and exact independent variable. Actually, the effect was parameter-dependent. For Ks, we
found an increase of its bias. However, the bias of qmax remained similar to the.
Unexpectedly, the bias of Ki decreased significantly (Table 1).
For LB-WLR, the bias of qmax and Ks increased when both variables are subject to error
(comparison LB-:/5µD¶ vs. LB-:/5µE¶7DEOH +RZHYHUWKH ELDV RI Ki decreases to
nearly one third of that with the method where only the dependent variable had error.
659
Start
software
on
A
True values
qmax, Ks, Ki
t=t
j
Eqs. 11 and
12 solve for
S
B
t, S
CVar
SP
based
Calculation
of bias eq.
16
End
In the
case
of the
a0, a1, a2; b0, b1,
b2
Estimated qmax,
Ks, Ki
Regression
2V
Eqs. 11a and
12a
LevenbergMarquardt
SURFHGXUH FRPSDULVRQ 63 µD¶
vs. 63 µE¶ WKH
biases of all the kinetic parameters increased when both variables exhibited errors (Table
1), which was consistent with the expectations and previous discussion in the open
literature (Kutner et al., 2004). Indeed, the bias increased by factors between 2 to 3,
depending on the kinetic parameter.
Start
A
B
t, S
CVar
and sd
End
t, Serr and
Calculation
of bias eq. 16
Linear system
eq. 15
a0,w; a1,w; a2,w; b0,w; b1,w;
b2,w
Estimated qmax,w ; K s,w,
K i,w.
FIGURE 1. Flow diagram of the synthetic batch weighed experiment and its analysis.
Above, simple two-variable linear regression; below, weighed two-variable linear
regression.
Keys: CVar: coefficient of variation; sd: standard deviation of substrate concentration; subscript w:
belonging to the weighed linear regression procedure; aj: regression coefficients in the kinetics
without inhibitor; bj: regression coefficients with inhibitor
Effect on weighing the uncertainty of heterscedastic dependenet variable on
parameter estimations. With the method based on LB when only the dependent variable
VKRZHG HUURU FDVH µD¶ XVLQJ WKH ZHLJKHG OLQHDU UHJUHVVLRQ OHG WR GHFUHDVHG ELDV RI DOO
WKH NLQHWLF SDUDPHWHUV FRPSDULVRQ /% µD¶ vs. LB-:/5 µD¶ 7DEOH 7KLV ZDV FRQVLVWHQW
with the claims in the classical textbooks regarding the benefits of using weighed linear
regression (Kutner et al., 2004; Kreyszig, 1971).
However, when both variables were subjected to error, the weighed linear regression
GLG QRW LPSURYH WKH SDUDPHWHU HVWLPDWLRQ LQ JHQHUDO FRPSDULVRQ /% µE¶ vs. LB-:/5 µE¶
Table 1): only the bias of Ki decreased, whereas the bias of qmax and Ks were similar to
WKRVHRIWKHEDVDOFDVHµD¶
660
When weighed linear regression was applied to data from the Batch model
(comparison Batch LR-2Var vs. Batch WLR-2Var, Table 1), the bias of Ks significantly
decreased. However, the bias of qmax and Ki were exceedingly high and superior to those
of the basal case by factors of 7 to 18, depending on the kinetic parameter. This was in
disagreement with expected results and the classical justification on the use of the
weighed regression.
General likelihood of the fitting methods. No single method exhibited minimum bias in
all kinetic parameters. Considering the scenario of only the dependent variable with error a
first compromise solution would consist of choosing the method with at least two smallest
biases: in this regard thHILUVWPHWKRGRIFKRLFHLV63µD¶ZKHUHDVWKHVHFRQGPHWKRGRI
choice would be LB-:/5µD¶7DEOH7KLVGLVFXVVLRQSRLQWVRXWWRDVHFRQGFRPSURPLVH
VROXWLRQ WKH XVH RI WZR PHWKRGV:LWK WKH PHWKRG 63 µD¶ JRRG HVWLPDWLRQV RI ERWK qmax
and Ks are obtained, whereas the fitting with method LB-:/5 µD¶ ZLOO JLYH WKH EHVW
estimation of Ki. Both methods are compatible because they use the same experimental
database, that is, the same set of matrices of initial rates and initial substrate
concentration).
:KHQ ERWK YDULDEOHV DUH VXEMHFW WR HUURU WKH PHWKRG 63 µE¶ GRHV QRW MXVWLI\ WKH
expectations of a better estimation compared to the traditional LB based on linearization of
WKHRULJLQDOGDWD/%µE¶,QIDFWWKHODWWHUH[KLELWHGORZHUELDVHVIRUqmax and Ki WKDQ63µE¶
(Table 1) and would be the method of choice, against expectations.
The batch method Batch LR-2Var, in spite of the advantage of smaller experimental
work, unfortunately leads to large biases in Ks and Ki (Table 1). The weighed regression
based on Batch is out of consideration. Why the large biases, particularly in Ki?. Likely due
to the biases of both a2 and b2 from regressions Eqs. 11 and 12, and their use in Eq. 13c.
The ratio b2/a2 propagates the bias into the calculation of Ki, in a manner similar to
propagation of errors (Kreyszig, 1971). There is a difference, though, between propagation
of bias and propagation of errors; the first is simply algebraic, whereas the second is
based on a pythagorean weighing of the variances of the variables. Moreover, batch
regressions either without or with weighing Eqs. 11 and 12, are statistically weak since the
dependent variable S in the left member appears also as logarithm in the right member, as
a pseudo independent variable (ln S). This is a violation of basic assumptions of the
multivariable linear regression model (Kutner et al., 2004), because some degree of
autocorrelation between Y (S) and X2 (ln S) could exist and badly distort the results of the
regression coefficients.
The Batch method could still be used whenever more quantitative information on the
reproducibility and structure of the bias of the kinetic parameters could be unveiled. This
knowledge could be obtained from a more exhaustive simulation under different scenarios
of initial substrate concentration, initial inhibitor concentration, different sets of true
parameter values, etc. The rationale is that if we know the structure of the bias, the bias
could be taken into account to correct the estimated values of the kinetic parameters. The
corrected values would be closer to the corresponding true values. Relative bias was
calculated as (estimated value ± true value)/ true value * 100 (eq. 16).
661
TABLE 1. Estimation of kinetic parameters of competitive-like microbial kinetics by
several fitting methods
Method
22.0
22.0
22.0
22.0
22.0
22.0
47.0
47.0
47.0
47.0
47.0
47.0
300.0
300.0
300.0
300.0
300.0
300.0
22.8
22.8
22.7
22.8
20.8
43.6
Estimated values and bias
Ks
Ki
rel
rel
bias
bias
(%)
(%)
3.64
57.2
21.70
352.6
3.53
59.3
26.11
285.4
3.18
56.7
20.64
317.9
3.61
59.4
26.40
303.5
-5.65
34.4
-26.76 335.8
98.16
21.6
-1.85
66.6
22.0
22.0
47.0
47.0
300.0
300.0
22.4
23.1
1.82
5.09
qmax
/%µD¶
/%µE¶
LB-:/5µD¶
LB-:/5µE¶
Batch LR-2Var
Batch
WLR2Var
63µD¶
63µE¶
True values
Ks
Ki
qmax
51.7
58.2
10.00
23.85
330.6
362.4
rel
bias
(%)
17.53
-4.87
5.98
1.18
11.95
-77.80
10.20
20.80
µD¶)LWWLQJPHWKRGZLWKHUURULQWKHGHSHQGHQWYDULDEOHRQO\µE¶)LWWLQJPHWKRGZLWKHUURUVLQERWKYDULDEOHV
LB: Lineweaver-Burk simple linear regression; LB-WLR: Lineaweaver-Burk weighed linear regression; Batch LR-2Var:
batch based simple linear 2-variable regression; Batch WLR-2Var: batch based 2-variable weighed linear regression; SP:
Software GraphPad Prism 5.0.
Conclusion
The outcome of this work is focused on the fitting of a competitive-like inhibition microbial
kinetics.
-As expected, all the methods reflected the features of competitive inhibition, that is, the
value of qmax was not affected by the inhibitor whereas the apparent affinity coefficient
increased with the concentration of inhibitor.
-No single method showed the minimum bias in all kinetic parameters. Considering the
scenario of only the dependent variable with error a first compromise solution the first
PHWKRGRIFKRLFHLV63µD¶ZKHUHDVWKHVHFRQGPHWKRGRIFKRLFHZRXOGEH/%-:/5µD¶-In order to improve the estimation of the 3 parameters, the use of two methods is
recommended as a second trade-RII:LWKWKHPHWKRG63µD¶JRRGHVWLPDWions of both qmax
and Ks are obtained, whereas the fitting with method LB-:/5 µD¶ ZLOO JLYH WKH EHVW
estimation of Ki. Both methods are compatible because they use the same experimental
database, that is, the same set of matrices of initial rates and initial substrate
concentration.
-:KHQ ERWK YDULDEOHV DUH VXEMHFW WR HUURU WKH PHWKRG 63 µE¶ GRHV QRW MXVWLI\ WKH
expectations of a better estimation compared to the traditional LB based on linearization of
WKHRULJLQDOGDWD/%µE¶,QIDFWWKHODWWHUH[KLEited lower biases for qmax and Ki WKDQ63µE¶
and it would be the method of choice, against expectations.
-The batch method Batch LR-2Var, in spite of the advantage of smaller experimental work,
leads to large biases in Ks and Ki (Table 1). The weighed regression based on Batch was
worse. This is unfortunate because from the experimental point of view the batch method
required significant less experimental work than the other methods.
-In general and against expectations, weighed linear regression of LB and batch models
did not improve significantly the accuracy of estimated kinetic parameters.
To the best of our knowledge, this is the first comprehensive, critical evaluation of
fitting methods to a microbial inhibition kinetics, based on a synthetic experiment (in silico).
662
With caution due to that, our results reveal that some assumed advantages of weighed
linear regression and less biased Levenberg-Marquardt based software were not
confirmed because they did not produce more accurate estimations of the kinetic
parameters.
Finally, the significance of kinetic parameter values cannot be underestimated. For
instance, if the estimated values of qmax and Ki are much higher than the true values, there
is a risk to underestimate the time to reach the remediation goals. So, kinetic parameter
values could significantly influence the time frame of a bioremediation project.
References
Bhushan B.; Chauhan A.; Sudip K.; Rakesh K. (2000). Kinetics of biodegradation of p-nitrophenol by
different bacteria. Biochem. Biophys. Res. Commun. 274, 626±630.
Camacho-Pérez B.; Ríos-Leal E.; Poggi-Varaldo H.M.; Rinderknecht-Seijas N. (2011). Desorption
enhancement and biostimulation in the bioremediation of a heavy agricultural soil polluted with lindane
in sequential slurry bioreactors. H.V. Rectanus and R. Sirabian (Chairs) Bioremediation and Sustainable
Environmental Technologies²2011. 1st. International Symposium on Bioremediation and Sustainable
Environmental Technologies (Reno, NV; June 27±30). Book in CD-ROM. ISBN 978-0-9819730-4-3.
Battelle, Columbus, OH.
Camacho-Pérez B.; Ríos-Leal E.; Rinderknecht-Seijas N.; Poggi-Varaldo H.M. (2012) Enzymes involved in
the biodegradation of hexachlorocyclohexane: A mini review. J. Environ. Manage. 95, S306-S318.
Camacho-Pérez B.; Ríos-Leal E.; Solorza-Feria O.; Vazquez-Landaverde P.; Barrera-Cortés J.; PonceNoyola M.; Garcia-Mena J.; Rinderknecht-Seijas N.; Poggi-Varaldo H.M. (2013). Performance of an
electrobiochemical slurry reactor for the treatment of a soil contaminated with lindane. Journal of New
Materials for Electrochemical Systems 16, 217-228.
Grady Jr, C.P. L.; Barth F.; Barbeau, S. (1996). Variability in kinetic parameter estimates: a review of
possible causes and a proposed terminology. Water Res. 30 (3), 742-748.
Grady Jr. C.P.L.; Daigger, G.T.; Love, N. G.; Filipe, C.D.M. (2011). Biological wastewater treatment. 3rd.
edition IWA Publ. and CRC Press, Boca Raton, FL. ISBN-13: 978-0849396793
Kreyszig, E. (1971) Introduction to mathematical statistics. John Wiley & Sons, New York, USA.
Kutner, M.; Nachtsheim, C.; Neter, J. (2004). Applied linear regression models. McGraw-Hill Book Co., 4th.
edition, New York.
Lodha B.; Bhat P.; Suresh K.; Vaidya N.; Mudliar S.; Killedar D.; Chakrabarti T.; (2007). Bioisomerization
kinetics of gamma-HCH and biokinetics of Pseudomonas aeruginosa degrading technical HCH.
Biochem. Eng. J. 35, 12±19.
Motulsky H.; Christopoulos A.; (2003). Fitting models to biological data using linear and nonlinear regression.
A practical guide to curve fitting. GraphPag Sowtfware Inc., San Diego CA.
Poggi-Varaldo, H.M. (2013 a). Solving the batch competitive inhibition kinetics by the methods of NewtonRaphson, iterative substitution, and regula falsi. Excel file, CINVESTAV del IPN, Mexico.
Poggi-Varaldo, H.M. (2013b). Development of the equations for estimating regression coefficients in the 3variable weighed linear regression problem. CINVESTAV del IPN, Mexico.
Robles-González I.; Ríos-Leal E.; Sastre-Conde I.; Fava F.; Rinderknecht-Seijas N.; Poggi-Varaldo H.M.
(2012). Slurry bioreactors with simultaneous electron acceptors for bioremediation of an agricultural soil
polluted with lindane. Process Biochem. 47; 1640±1648.
Salvadori, M.G.; Baron, M. L. (1959). Numerical methods in engineering. Prentice-Hall, 4th. edition,
Englewood-Cliffs, NJ.
Segel, I. H. (1976). Biochemical calculations ± How to solve mathematical problems in biochemistry. John
Wiley & Sons Inc. 2nd. edition, New York, USA.
Spector T.; Hajian G. (1981). Statistical Methods to Distinguish Competitive, Noncompetitive, and
Uncompetitive Enzyme Inhibitors. Analytical Biochemestry 115, 403-409.
Turlough F. (2008). Ex-situ bioremediation of chlorobenzenes in soil. J. Haz. Mater. 154, 9±20.
Varo-Arguello W.; Camacho-Pérez B.; Ríos-Leal E.; Vazquez-Landaverde P.; Ponce-Noyola M.; BarreraCortés J.; Sastre-Conde I.; Rindernknecht-Seijas N.; Poggi-Varaldo H.M. (2012). Triphasic slurry
663
bioreactors for the bioremediation of lindane-impacted soil under aerobic and anaerobic conditions.
Environ. Eng.Manage. J. 11(10), 1811-1823.
Wang, M.J.; Jones K. (1994). Behaviour and fate of chlorobenzenes (cbs) introduced into soil-plant systems
by sewage sludge application: a review. Chemosphere, 28 (7), 1325-1360.
664
LIST OF COAUTHORS:
Gonzalo Manuel Escobedo-Acuña, M. Sc. Candidate, presenting author. Same address
and e-mail as those of Dr. H.M. Poggi-Varaldo.
Beni Camacho-Pérez, Sc. D. Candidate. Same address as Dr. H.M. Poggi-Varaldo. Email: [email protected]
Elvira Ríos-Leal, full Profesor. Same address as Dr. H.M. Poggi-Varaldo. E-mail:
[email protected]
María Teresa Ponce-Noyola, Sc. D., Full Professor, Same address as Dr. H.M. PoggiVaraldo. E-mail: [email protected]
Josefina Barrera-Contreras, Sc. D., Full Professor, Same address as Dr. H.M. PoggiVaraldo. E-mail: [email protected]
Héctor B. Cuauhtémoc Poggi-Rinderknecht, B. Sc.;
Noemí Rinderknecht-Seijas, Full Professor.
ESIQIE-IPN, Division of Basic Sciences, Campus Zacatenco, México D.F., Mexico. E-mail:
[email protected]
Héctor M. Poggi-Varaldo, Sc D, Full Professor. Author to whom all correspondence
should be addresed
CINVESTAV, Dept. Biotechnology and Bioengineering, Environmental Biotechnology and
Renewable Energies R&D Group; P.O. Box 14-740, México DF, 07000, México.
Tel: 5255 5747 3800 ext 4324, Fax. 5255 5747 3313,
665
CHAPTER 6.11. BIOREMEDIATION OF A SOIL CONTAMINATED WITH LINDANE IN
AN ELECTROBIOCHEMICAL SLURRY REACTOR: ROLE OF SURFACTANT AND
CHARACTERIZATION OF MICROBIAL COMMUNITY
Beni Camacho-Pérez (1); Jaime García-Mena (2); Omar Solorza-Feria (3);
Elvira Ríos-Leal (1); Selvasankar Murugesan (4); Alberto Piña Escobedo (2);
Héctor Mario Poggi-Varaldo * (1)
(1) Dept. Biotechnology and Bioengineering, CINVESTAV ± IPN
(2) Dept. of Genetics and Molecular Biology, CINVESTAV ± IPN
(3) Dept. of Chemistry, CINVESTAV ± IPN
(4) Dept. of Pharmacology, CINVESTAV ± IPN
Lindane is an organochlorinated pesticide that has been extensively used worldwide for
the control of agricultural and medical pests. Bioslurry reactors are the most common exsitu reactor configuration for soil bioremediation because of several advantages. On the
other hand, electrobiochemical slurry reactor (EBCR) constitutes a promising technology
that could remove organic and recalcitrant pollutants in contaminated soils with
simultaneous electricity output.
The aims of this research were (i) to evaluate the desorption of lindane from soil with
Tween 80 at different concentrations; (ii) to determine the power output and removal of
lindane from soil in presence of Tween 80 in the EBCR; and (iii) to characterize the
microorganisms in the electrobiochemical slurry reactor using the Ion Torrent Personal
Genome Machine.
In vitro adsorption and desorption tests were run with the model soil polluted with
lindane and varying concentrations of Tween 80. The EBCR consisted of a Plexiglass
cylinder approximately 6 cm in diameter and 8 cm in length, fitted with two anodes and two
cathodes. Tween 80 was added at a dose of 2000 mg /L in the batch run of biotreatment of
soil in the EBCR.
Approximately 10% of lindane was removed in the in vitro desorption experiments. The
LQWHUQDOUHVLVWDQFHRIWKH(%&5GHWHUPLQHGE\SRODUL]DWLRQFXUYHZDVȍDPRGHUDWH
volumetric power activity was recorded (374 mW/m3) along with a potential of 600 mV
when the two-electrode sets were connected in parallel. During the batch operation, the
EBCR showed a 56% lindane removal (Klindane) whereas the reduction in the abiotic control
was only 3%. Unexpectedly the Klindane in the EBCR spiked with Tween 80 was lower than
that in an EBCR operated without surfactant in a previous experiment. This could be
ascribed to the increased degradable organic matter supply related to Tween in our EBCR
that possibly shifted microbial metabolism from lindane degradation to degradable matter
uptake. An average volumetric power of 685 mW/m3 and average voltage of 420 mV were
achieved.
-------------*Author for all correspondence, [email protected] 666
The energy output obtained in our EBCR allows for the recovery of 20% of the power
required for mixing, thus paving the way to sustainable bioremediation of soils. The
bacterial community was analyzed in the slurry of EBCR the 0, 15, 22, and 30 days of
operation. The most representative phyla were Actinobacteria, Bacteroidetes, Firmicutes,
and Proteobacteria, that was consistent with findings reported in the open literature. The
most dominant genera were Arcobacter (40.8%), Thrichococcus (41.6%) at 22 days of
operation, when the EBCR exhibited the maximum voltage. Arcobacter was reported to
harbor in electrodes of MFC, and rapidly generates a strong electronegative potential as a
pure culture when it is supplied with acetate. Thrichococcus has been reported in
autotrophic biocathodes of other bioelectrochemical systems.
Keywords: electrobiochemical slurry reactor, Ion Torrent Personal Genome Machine,
lindane, Tween 80,
667
Introduction
Agriculture is an essential activity of modern societies. While agricultural products provide
most of the food consumed by the world population, they also impose great pressures on
global natural resources. Reliance on pesticides and fertilizers because of intense
agricultural practices could also contribute to environmental degradation (Macary et al.,
2014; Sattler et al., 2007). Lindane (1, 2, 3, 4, 5, 6 - hexachlorocyclohexane, Ȗ-HCH) is a
moderately lipophilic, organochlorinated substance characterized by a high partition
coefficient octanol-water Kow § 4*103 with a low solubility in water, approx. 7 mg L-1 at
20ºC, and slight polarity due to the strong electronegative effects of chlorine atoms bound
to the aliphatic ring (Robles-González et al., 2012; Varo-Arguello et al., 2012; CamachoPérez et al., 2012). The Ȗ-HCH has been used in important agricultural applications as on
fruits, vegetables crops, and forestry for control of leaf-eating insects, and it is used as a
fungicidal seed treatment worldwide (Li et al., 1998; Li et al., 1999; Prakash et al., 2004). It
is a Persistent Organic Pollutant (POP), potent carcinogen and teratogen and classified by
WKH :RUOG +HDOWK 2UJDQL]DWLRQ :+2 DV µPRGHUDWHO\ KD]DUGRXV¶ KXPDQ RUDO /'R mg kg-1). Paknikar et al. (2005), have been reported potential adverse health effects
include neurological problems and immunosuppression in human and liver cancer in rats
and mice (Willet et al., 1998). The use of lindane has been restricted since the 1970s and
banned later on because of their toxicity, but the problem of residues of lindane remains
because of the high persistence and inter-conversion of lindane in soil (Prakash et al.,
2004; Willett et al., 1998; Quintero et al., 2005a). Under natural conditions, the high
adsorption of the hydrophobic compounds in the soil solid phase (mainly in the organic
matter) and its restricted availability for the biological action of endogenous or exogenous
microorganisms limits the extent and rate of degradation. The size of particles and
aggregates, and soil structure, also contribute to making the desorption rate (mass transfer
rate) the limiting stage in their degradation (Quintero et al., 2005b). One possibility to favor
the mobility of the pollutants is their transfer from the soil to the liquid phase; this possibility
would enhance the pollutant bioavailability. Bioavailability is considered a dynamic
process, determined by the rate of substrate-mass transfer to microbial cells relative to
their intrinsic catabolic activity
One alternative to facilitate lindane desorption from the soil, and thus to increase
lindane bioavailability in to use ex-situ technology such as soil slurry reactors. This
technology consist of treating mixtures of contaminated soil and water in controlled stirred
reactors (Cookson, 1995).The advantages of the slurry reactor is the feasibility of
controlling operating parameters such as mixing (facilitates aeration and enhances the rate
of chemical exchange between soil particles), controlling pH, temperature, providing
nutrients and other specialized amendments such as surfactants (Robles-González et al.,
2008; Venkata-Mohan et al., 2007; Robles-González et al., 2006).
The addition of surfactants to increase the mass-transfer rate of PAHs in soils has
received attention (Di Gennaro et al., 2008). Surfactants consist of organic molecules with
a hydrophobic and a hydrophilic part and can interact with polar as well as nonpolar
surfaces. At low concentrations, surfactants exist solely as monomers; above a critical
aqueous concentration, which is specific for each surfactant monomer, called critical
micellar concentration (CMC), the surfactant monomers are aggregated in solution to form
entities made up of a hydrophobic core and a hydrophilic shell (Mulligan et al., 2001).
668
These micelles result in increased pseudo-water-solubilities of pollutants, thereby
increasing the concentration gradient and mass transfer rates (Quintero et al., 2005b; Di
Gennaro et al., 2008; Mulligan et al., 2001).
On the other hand, microbial fuel cells (MFC) constitute a promising technology for the
biodegradation of several organic substrates such as glucose, acetate, xylose, cysteine,
cellulose, leachates from solid substrate fermentation of municipal wastes and
waterwastes, and other organic pollutants with simultaneous power generation (ValdezVázquez et al., 2006; Poggi-Varaldo et al., 1999; Du et al., 2007; Liu et al., 2004; Logan et
al., 2006; Morris et al., 2008; Ortega-Martínez et al., 2012; Pant et al., 2010; Poggi-Varaldo
et al., 2009; Rezaei et al., 2009; Sathish-Kumar et al., 2012; Vázquez±Larios et al., 2010;
Vázquez±Larios et al., 2011; Camacho-Pérez et al., 2013). Recently, it has been proposed
that soil microbial fuel cell (SMFC) technology could be applied to enhance the removal of
organic matter, lindane, phenol, and petroleum hydrocarbons in contaminated soil, while at
the same time allowing electric energy generation (Camacho-Pérez et al., 2013; Huang et
al., 2011; Wang et al., 2011). The aims of this research were (i) to evaluate the desorption
of lindane from soil with Tween 80 at different concentrations; (ii) to determine the power
output and removal of lindane in an electrobiochemical slurry reactor; and (iii) to
characterize the dominant microorganisms in the electrobiochemical slurry reactor using
16S rRNA and denaturing gradient gel electrophoresis; and (iv) to characterize the
microorganisms in the electrobiochemical slurry reactor using the Ion Torrent Personal
Genome Machine. We used a difficult to remediate soil. Since it had high contents of
organic matter and clay.
Desorption tests. Desorption assays were carried out in serum bottles at 100 rpm for 6
days in order to evaluate the effect of surfactant on lindane desorption. In each serum
bottles was loaded with 5 g of sterilized soil polluted with lindane (100 mg kg -1 dry soil), 25
mL of a solution Tween 80 at different concentrations (0, 75, 500, 2000, 5000 mg L -1).
Sodium azide (300 mg L-1), was added as a microbiological inhibitor, and CaCl2 (0.01M).
The supernatant solution was analysed for lindane concentration.
Electrobiochemical slurry reactor. EBCR consisted of a Plexiglass cylinder
approximately 6 cm in diameter and 8 cm in height (308 mL capacity), fitted with two
anodes and two cathodes. The anodes were graphite discs (5cm D x 0.5 cm) whereas the
cathodes were of Toray carbon cloth, the cathodes were in contact with atmospheric air
(Figure 1). The electrodes were separated by a cation exchange membrane (Nafion 117,
coated with 0.5 mg cm-2 platinum catalyst, Pt 10wt%/C-ETEK) and was inoculated with a
sulfate reducing inoculum acclimated to lindane (Camacho-Pérez et al., 2013).
The EBCR was batch-operated for 30 day at room temperature. The concentration of
soil was 33% w/v and of Tween 80 was 2000 mg L -1. The EBCR was fed a solution stock
of sucrose: sodium acetate: lactate to give a final concentration of 2 g COD L -1 in the
EBCR at 0 y 15 d. The contents of the EBCR were performed in an orbital shaker at 100
rpm. Measurements of the power output were performed using a Multimeter ESCORT 3146A. The
process control was EBCR operated under open-circuit with live inoculum and soil.
669
FIGURE 1. Schematic diagram of electrobiochemical slurry reactor.
Determination of internal resistence of the electrobiochemical slurry reactor. The
internal resistance was determined using the polarization curve method, by varying the
external resistance (100- ȍ DFFRUGLQJ WR SURFHGXUHV RXWOLQHG E\ HOVewhere
(Logan et al., 2006; Poggi-Varaldo et al., 2009; Vázquez-Larios et al., 2010; SathishKumar et al., 2012), this was carried out 0 y 7 d of operation.
Chemicals. The Ȗ-HCH isomer (97% purity) was purchased from Sigma-Aldrich. Lindane
is a moderately lipophilic, organo-chlorinated substance characterized by a high partition
coefficient octanol-ZDWHU .RZ § ZLWK ORZ VROXELOLW\ LQ ZDWHU DSSUR[ PJ / -1 at
20ºC, and slightly polar due to the strong electronegative effects of chlorine atoms bound
to the aliphatic ring. Hexane and acetone were of analytical grade. Tween 80 was
purchased from Sigma-Aldrich.
Soil. An agricultural soil with high contents of organic matter and clay, was sieved through
a 20 mesh, and sterilized by tyndallisation. The soil was contaminated with 100 mg lindane
kg-1. Soil pH was determined in a slurry soil/deionized water 1:2 (w/w) (Robles-González
et al., 2012), soil texture was measured by the hydrometer method, soluble BOD and
soluble COD were determined in soil water extracts, according to the Standard Methods
(1992). The main characteristics of the soil used in this work were: soil type, Cambisol; pH,
7.2; organic matter (%), 8.1 ± 0.1; soluble COD (mg COD kg-1), 5100 ± 436; soluble BOD
(mg BOD5 kg-1), 3725 ± 350; clay content (%), 42.3 ± 0.8; sand (%), 36.5 ± 2.7; silt (%),
21.2 ± 3.3. The texture was clayish and the hydraulic conductivity was low.
Metagenome sequencing using the Ion Torrent (PGM) Platform. Total genomic DNA of
the collected samples was extracted using PowerSoil® DNA Isolation kit (MoBio)
according WR WKH PDQXIDFWXUHU¶V LQVWUXFWLRQV 7KH H[WUDFWHG '1$ ZDV DPSOLILHG ZLWK WKH
primers targeting V3 region of the 16S rDNA. The forward primer was V3-341F containing
a 12bp Golay barcode (Fierer et al., 2008) and adapters for massive sequencing, and
antisense V3-518R containing A and Truncated P1 adapters (Whiteley et al., 2012). The
PCR mixture consisted of Buffer (1X), MgCl2 (2 mM), dNTP´s (0.2 mM), Taq DNA
polymerase (0.025U µL-1), forward primer V3-341F (0.2µM), reverse primer V3-518R
670
(0.2µM), and template DNA (10 ng). Amplification was performed in a GeneAmp PCR
System 2700 Thermocycler (Applied Biosystems). The PCR amplification was performed
with an initial denaturation (95°C for 5 min) followed by 30 cycles of denaturation (94°C for
15 s), annealing (62°C for 15 s) and extension (72°C for 15 s) and final extension (72°C
for 10 min). The PCR products were purified by electrophoretic separation on 2% agarose
gel and using Wizard SV Gen PCR Clean-Up System (Promega). The concentration of
PCR amplicons were measured by NanoDrop spectrophotometer (ThermoScientific).
After that the samples were sent out for the barcoded libraries preparation and sequencing
on an Ion Torrent PGM with 316 chip using the Ion Sequencing 200 bp kit ( Life
Technologies) according to the standard protocol (Ion Xpress_ Plus gDNA and Amplicon
Library Preparation, Life technologies).
After sequencing, sequence reads were filtered by the PGM software to remove low
TXDOLW\DQGSRO\FORQDOVHTXHQFHV'XULQJWKLVSURFHVVVHTXHQFHVPDWFKLQJWKH¶-adapter
were automatically trimmed and filtered. Sequenced data-Microbiota analysis Ion torrent
PGM software, Torrent_Suite v 4.0.2 was used to demultiplex the sequenced data based
on their barcodes and poor quality reads were eliminated from the datasets, i.e. quality
score <20, containing homopolymers >6, length <200 nt, and containing errors in primers
and barcodes. Filtered datas were exported as Fastq files. Demultiplexed sequencing data
were analized using QIIME version 1.8.0 software pipeline (Caporaso et al., 2010). Fastq
files were converted into Fasta files, and all the demultiplexed files were concatenated into
a single fasta file. Closed reference Operational taxonomic units (OTU) were determined
at 97% similarity level with UCLUST Algorithm (Edgar, 2010).
Chimeras were detected and removed from the datasets using the Chimera Slayer
(DeSantis et al., 2006). Sequence alignments were done against the Greengenes core set
(Haas et al., 2011). Statistical analysis of the OTU table Microbial diversity had been
assessed through both alpha and beta diversity. Using rarefied OTU tables, alpha
diversities were calculated using various matrices like Shannon, PD whole tree, chao1,
observed species. The beta diversity analysis was calculated using UniFrac analysis
(Vázquez-Baeza et al., 2013), by phylogenetic tree computed with FastTree and a
rarefiled biom table as inputs. Abundance of the bacterial groups at different taxonomic
levels (phylum, order, and genus) was separately explored with a Principal Component
Analysis (PCA) and Unweighted Pair Group Method with Arithmetic mean (UPGMA)
Clustering.
Calculations of ecological indices. Shannon-Weaver diversity index has been a popular
index in the ecological literature (Shannon et al., 1948; Ortega-Martínez et al., 2013).
Shannon-Weaver index is defined as:
(4)
Where:
pi= is the proportion of characters belonging to the i type of letter in the string of interest;
yet, in ecology, pi is often the proportion of individuals belonging to the i species in the
dataset of interest, S is the total number of species.
Pielous evenness index is a measure of a biodiversity which quantifies how equal
the community is numerically (Mulder et al., 2004; Ortega-Martínez et al., 2013).
This index is defined as:
671
(5)
where:
H´ = is the Shannon-Weaver index
S = is the total number of species
J´ = is constrained between 0 and 1, 1 indicates máximum evenness or the equitative
distribution of species.
It is worth emphasizing that the log in Equations (4) and (5) should be on base 2, since
the ShannoneWeaver index was originated on the grounds of the theory of information that
uses binary system of numbers. Yet, in the literature there is a variety of bases. Most
researchers have chosen base 10, others have selected base e (natural logarithms) for the
calculation of H´. Because of this, we distinguish and report here the índices H´2, Hé, and
H´10 to indicate that the base of the logarithm is 2, e, and 10, respectively. The base of log
S in the denominator of the Pielou index should be consistent with base of the log used in
the calculation ShannoneWeaver index. Furthermore and interestingly, Pielou index is
invariant with respect to the base of the logarithm, as long as the base of logs in the
numerator and denominator are the same (Shannon, 1948).
Lindane and metabolite analysis. The procedure for the extraction and determination of
HCH and intermediate metabolites in the electrobiochemical reactor were performed
according by Quintero et al. (2006). The lindane was analyzed in a Perkin Elmer gas
chromatograph equipped with an electron capture detector. The intermediate metabolites
were analyzed in a Perkin Elmer GC/MS (Camacho-Pérez, 2013).
Desorption tests. Figure 2 shows the lindane desorbed at different initial concentrations
of Tween 80 in the water phase. The maximum desorption effect was observed at a dose
of 2000 mg L-1 of Tween 80, 9 mg of lindane removed kg-1 of dry soil. Because of this, we
applied an initial concentration of 2000 mg L-1 of Tween 80 in our EBCR.
672
FIGURE 2. Lindane desorbed from the soil to the water phase. The initial
concentration of lindane in soil was 100 mg kg-1 of dry soil.
Characterization of the electrobiochemical slurry reactor. The values obtained of
internal resistance from the polarization curves method were, 5660, \ȍ
for face A, face B, connection in series and parallel, respectively, at time 0 day (Table 1).
The maximum volumetric power was obtained when the connection was in parallel (60
mW m-3) followed by connection in series, face A, and face B with 38, 12 and 11 mW m-3
respectively. After 7 days of operation another electrochemical characterization was
carried out. The internal resistances decreased very much compared to those of the first
characterization. Their values were approximately o 4652, 3867, 1041, 804 ȍ for face A,
face B, connection in series and connection in parallel, respectively (Table 1). The
maximum volumetric power was obtained for parallel connection (373 mW m -3, Table 1); it
was sixfold the volumetric power obtained with characterization at 0 days.
TABLE 1. Values of several variables of electrobiochemical slurry reactor
characterization at 0 and 7 day.
Parameter
Time (days)
Rint ȍ
PAn-max (mW m-2)
0
5660
0.95
Face A
7
4652
12.33
0
6022
0.86
Face B
7
3687
8.01
0
12419
3.04
Series
7
1041
29.58
0
3845
4.84
Parallel
7
804
29.90
PV-max (mW m-3)
IEBCR-max (mA)
EEBCR-max (V)
PEBCR-max (mW)
PAn-ave (mW m-2)
PV-ave (mW m-3)
IEBCR-ave (mA)
EEBCR-ave (V)
PEBCR-ave (mW)
11.86
0.11
0.31
0.003
0.37
4.57
0.02
0.13
0.001
154
0.67
0.57
0.05
4.54
56.82
0.11
0.30
0.02
10.72
0.08
0.32
0.001
0.39
4.86
0.02
0.13
0.001
100
0.55
0.58
0.03
3.93
49.20
0.09
0.30
0.02
38.08
0.08
0.50
0.01
1.04
13.01
0.03
0.21
0.004
369
0.72
0.60
0.11
9.98
124.79
0.18
0.38
0.04
60.55
0.11
0.54
0.02
1.72
21.46
0.05
0.25
0.007
373
0.94
0.59
0.12
9.92
123.96
0.18
0.39
0.04
Notes: Rint: internal resistance; PAn, surface area power density; PV, volumetric power; EEBCR, voltage;
IEBQR, current intensity; PEBCR, power delivered. Subindices: max, maximum; ave, average.
673
The improved characteristics might be a consequence of the increased microbial
activity resulting from enrichment of the biofilm on the anode (Lu et al., 2009). Figure 3
shows the time course of the potential delivered by the device when the anodes and
cathodes of the EBCR were connected in parallel; results of the average performance are
exhibited in Table 2. The voltage with the EBCR in open circuit conditions (at the early 11
h) was approximately 666 mV (phase I). The voltage remained stable when the cell was
RSHUDWHGZLWKDQH[WHUQDOUHVLVWDQFHRIȍDYHUDJHP9ILUVWQLQHWHHQRISKDVH
II); however the voltage decreased to less than 160 mV afterwards. So, open circuit
conditions were re-established in phase III where an expected increase of voltage
occurred. Subsequently, in phase IV, the cell was operated with an external resistance of
ȍWKHYROWDJHUHPDLQHGVWDEOHP9The EBCR at 16 d was supplemented with 2
g L-1 substrate (sucrose: sodium acetate: lactate) and Tween 80 (2 000 mg L -1) that was
used as the fuel. The electricity generation began to increase and reached a voltage
output of approximately 488 mV (Figure 3).
The maximum voltage output of the EBCR (580 mV) and maximum power (108 mW m -2)
were higher than those reported by Wang et al. (2011) (155 mV and maximum surface
power 0.85 mW m-2) for a cell loaded with soil polluted with total petroleum hydrocarbons.
Our results also compared very favorably to those observed by Yan et al. (2012) who
reported a voltage as low as 17 mV in the treatment of sediment contaminated with
phenanthrene and pyrene. On the other hand, Huang et al. (2011) registered a power
density ca. 30 mW m-2 and a lower voltage (150 mV) in the treatment of a waterlogged soil
polluted with phenol.
800
I
II
IV
III
700
EEBCR (mV)
600
500
400
300
200
100
0
0
5
10
15
20
25
30
35
Time (d)
FIGURE 3. Electricity generation in electrobiochemical slurry reactor during batch operation
for 30 d. The addition of substrate is indicated by the blue . Phase I, open circuit; phase II, closed
circuit with external resistance 3300ȍ; phase III, open circuit; phase IV, external resistance 820ȍ.
674
Our EBCR exhibits a bonus besides soil remediation: the bioelectricity generation. The
latter was estimated with Equation 1 below
Êharvested (MJ/tonne soil) = Pave (W)*(30 d*24 h/d*3600 s/h)/(0.0001 tonne soil * 1000000
J/MJ)
(6)
where Êharvested is the energy produced in MJ tonne soil-1, Pave is the average power in 30 d
of operation, 24 h d-1 and 3600 s h-1 are conversion factors for converting time from
days to seconds; 0.0001 tonne soil is soil mass in the lab scale EBCR.
The Êharvested was estimated at 5.9 MJ tonne-1 soil during 30 days of operation. This
energy can partially offset the power required for mixing of both the EBCR and typical SBs.
For instance, power requirements for mixing are determined empirically and can be
estimated from manufacturer´s equipment specifications. Indeed, typical power
requirements for complete mixing are in the range 20 to 50 kW/1000 m 3 for moderately
thick suspensions (Eweis et al., 1998); a mid-point value of 35 kW/1000m3 was chosen.
Performing the calculations with similar assumptions to those of Equation 6, the energy
required for mixing during the 30 d batch would be given by Equation 7 below
Êmixing = 35 (W/m3)*(0.33 tonne soil/m3)*(30 d*24 h/d*3600 s/h) = 29.9 MJ/tonne soil
(7)
That is, the EBCR allows for a bioelectricity harvest that could represent ca. 20 % of the
energy required for mixing.
TABLE 2. Average performance of electrobiochemical slurry reactor
Parameter
PAn-max (mW m-2)
PV-max (mW m-3)
EBCR-max (V)
IEBCR-max (mA)
PEBCR-max (mW)
PAn-ave (mW m-2)
PV-ave (mW m-3)
EEBCR-ave (V)
IEBCR-ave (mA)
PEBCR-ave (mW)
108
1357
0.58
0.71
0.42
54 ± 30
685± 377
0.42± 0.12
0.47 ± 0.17
0.19 ± 0.02
Notes: PAn, surface area power density; PV, volumetric power; EEBCR, voltage; IEBCR, current intensity;
PEBCR, power delivered. Subindices: max, máximum; ave, average. Standard desviation is with respect to
time
Lindane removal and intermediate metabolites. Lindane removal achieved in the EBCR
was 56 %, whereas the removals of the biotic (live) control and abiotic control EBCRs were
72 and 3%, respectively. Main metabolites due to lindane degradation in the EBCR were
detected by analysis by GC/MS in the EBCR: 1,2,3-trichlorobenzene (1,2,3 TCB), 1,4
dichlorobenzene (1,4-DCB), 1,2-dichlorobenzene (1,2-DCB), and chlorobenzene (CB)
675
(Figure 4). Lindane removals observed in our EBCR compared similarly with lindane
removals reported for standard slurry bioreactors in the literature. Okeke et al. (2002)
carried out experiments with SB inoculated with Pandorea sp., with a presumably
anaerobic operation of 9 weeks duration. Initial lindane concentration was 100 mg kg -1;
they found removals of 59.6% Ȗ-HCH.
Unexpectedly the Klindane in EBCR was lower than that in EBCR operated without
surfactant, 78% (Camacho-Pérez et al., 2013). This could be ascribed to the influence of
increased degradable organic matter supply in the experiment. Quintero et al. (2005a)
treated a sandy soil polluted with a mixture of isomers Į, ȕ, Ȗ and į-HCH (100 mg kg-1
each) in anaerobic SB. Starch was supplemented at 2 g L-1 every 3 days. High removals of
nearly 100% for Į and Ȗ isomers of HCH and 65 to 70% for ȕ and į HCH were found. On
other hand, Robles-Gonzalez et al. (2012) studied the bioremediation of a heavy soil
polluted with 100 mg lindane kg-1 in full sulfate reducing SB. Removal was 88% whereas
the detected metabolites after 30 d operation were PCCH; 1,2,4-TCB; 1,2,3-TCB; CB, and
benzene; they also demonstrated that in methanogenic SBs the removal of lindane was
between 41-47%. The SB was operated with similar soil operational conditions. In other
studies of our Group, Varo-Arguello et al. (2012) carried out experiments with SB in
sulphate-reducing conditions; they reported removals of 78% Ȗ-HCH in 30 days.
FIGURE 4. GC-MS detection of intermediate metabolites in electrobiochemical slurry reactor
at the end of operation (30 d)
Microbial consortia characterization. Figure 6 shows the composition of bacterial
population obtained from sulphate inoculum and EBCR at different days of operation. The
more representative phylum were actinobacteria, bacteroidetes, firmicutes, proteobacteria
and synergistetes.
Kim et al. (2013) performed a phylogenetic analysis of microbial samples that revealed
a diverse bacterial community consisting of Proteobacteria, Firmicutes, Bacteroidetes and
Spirochaetes. The authors worked with a single MFC with air cathode membrane
electrode assembly combined into duplicate two-module longitudinal tubular reactors. The
reactor was inoculated with anaerobic digester sludge.
In our work, the most dominant genera belonged to Arcobacter, 40.8%, and trichococcus
(41.6%) at 22 days of operation (Table 6), when the EBCR exhibited the maximum voltage.
Arcobacter was reported to harbor in electrodes of MFC, and rapidly generates a strong
electronegative potential as a pure culture when it is supplied with acetate (Fedorovich et
al., 2009). Thrichococcus has been reported in autotrophic biocathodes of other
bioelectrochemical systems (Zaybak et al., 2013). Trichococcus was identified as
facultatively anaerobic bacteria by Kim et al. (2013)
676
FIGURE 6. Phylum level identification by massive semiconductor sequencing of
microbial community in our electrobiochemical slurry reactor.
Analysis of ecological indices. Shannon-Weaver index and Pielou´s indices for the
sulphate reducing inoculum were 1.54 and 0.38 respectively (Table 5). Ortega-Martínez et
al. (2013) characterized a sulphate reducing inoculum, they reported 1.84 and 0.66 for
Shannon-Weaver index and Pielou´s index respectively.
In the case of the
electrobiochemical slurry reactor , Shannon-Weaver index was 2.42, 2.00, 1.33, and 1.39
(Table 5) for 0, 15 , 22, and 30 days of operation respectively. The Pielou´s index for
electrobiochemical slurry reactor was 0.59, 0.49, 0.32, and 0.39 (Table 5) for 0, 15, 22,
and 30 days of operation respectively. The above results concluded that the different
distributions of species were identified in the electrobiochemical slurry reactor at diferents
days of operation.
TABLE 5. Ecological indices of the electrobiochemical slurry reactor based on
phylum
Sample
SR-Ind
0de
15de
22de
30de
Shannon-Weaver index based on phylum
Héa
H´10b
H´2c
1.07
1.68
1.39
0.92
1.10
0.46
0.73
0.60
0.40
0.48
1.54
2.42
2.00
1.33
1.59
Pielou´s index based on phylum
J´
0.38
0.59
0.49
0.32
0.39
Notes: a, calculated with logarithm base e; b, calculated with logarithm base 10; c, calculated with logarithm
base 2; d, sulphate reducing inoculum; e. operation time
677
TABLE 6. Genus level identification of all the sequences (Only relative abundances
of identified Genus higher than 1 % are listed and all the other sequences are
LQFOXGHGLQ³RWKHUV´
Genus/Time (Days)
Actinobacteria Acidimicrobiia Acidimicrobiales
Actinobacteria Actinobacteria Actinomycetales Actinomycetaceae __N09
Actinobacteria Actinobacteria Actinomycetales Nocardioidaceae
Propionicimonas
Actinobacteria Actinobacteria Actinomycetales Propionibacteriaceae
Bacteroidetes Bacteroidia Bacteroidales
Bacteroidetes Bacteroidia Bacteroidales Porphyromonadaceae
Bacteroidetes Bacteroidia Bacteroidales Prevotellaceae Prevotella
Bacteroidetes Bacteroidia Bacteroidales Rikenellaceae Blvii28
Firmicutes Bacilli Bacillales Bacillaceae Bacillus
Firmicutes Bacilli Lactobacillales Carnobacteriaceae Trichococcus
Firmicutes Bacilli Lactobacillales Lactobacillaceae Lactobacillus
Firmicutes Bacilli Lactobacillales Streptococcaceae Lactococcus
Firmicutes Clostridia Clostridiales
Firmicutes Clostridia Clostridiales Clostridiaceae
Firmicutes Clostridia Clostridiales Clostridiaceae Clostridium
Firmicutes Clostridia Clostridiales Lachnospiraceae
Firmicutes Clostridia Clostridiales Lachnospiraceae Anaerostipes
Firmicutes Clostridia Clostridiales Ruminococcaceae
Firmicutes Clostridia Clostridiales Ruminococcaceae Ethanoligenens
Firmicutes Clostridia Clostridiales Ruminococcaceae Faecalibacterium
Firmicutes Clostridia Clostridiales Ruminococcaceae Oscillospira
Firmicutes Clostridia Clostridiales [Mogibacteriaceae] Anaerovorax
Firmicutes Clostridia Clostridiales [Tissierellaceae] Sedimentibacter
Proteobacteria Alphaproteobacteria Sphingomonadales
Sphingomonadaceae
Proteobacteria Alphaproteobacteria Sphingomonadales
Sphingomonadaceae Sphingomonas
Proteobacteria Deltaproteobacteria Desulfovibrionales Desulfovibrionaceae
Desulfovibrio
Proteobacteria Deltaproteobacteria Desulfuromonadales Pelobacteraceae
Proteobacteria Epsilonproteobacteria Campylobacterales
Campylobacteraceae Arcobacter
Proteobacteria Epsilonproteobacteria Campylobacterales Helicobacteraceae
Proteobacteria Gammaproteobacteria Enterobacteriales Enterobacteriaceae
Proteobacteria Gammaproteobacteria Enterobacteriales Enterobacteriaceae
Enterobacter
Proteobacteria Gammaproteobacteria Pasteurellales Pasteurellaceae
Haemophilus
Proteobacteria Gammaproteobacteria Thiotrichales Piscirickettsiaceae
Thioalkalimicrobium
Synergistetes Synergistia Synergistales Dethiosulfovibrionaceae HA73
Synergistetes Synergistia Synergistales Thermovirgaceae
Thermotogae Thermotogae Thermotogales Thermotogaceae Kosmotoga
Others
SR In
0
15
22
30
0
0
6.2
1.3
1.2
0
0.6
0.6
0.1
0.2
0.2
0
0.3
0.9
0.1
6.2
0
0
6.2
6.2
6.2
6.2
0
6.2
12.5
0
0
6.2
6.2
0
0
12.5
6.2
0
0
0
0.2
4.2
0.9
2.1
0.1
2
1.9
1.6
0.2
0.8
1.3
8.8
1.9
0
1.8
1.8
1.5
0.4
0.2
0.1
0.2
0
17.6
3.5
0.4
0.5
1.9
7.4
0
0
3.2
0.6
3.9
6.9
0
2.9
0
0.5
0.2
1.6
3.5
1.1
0
0.3
5.4
0
0
0.6
41.6
0
0
0.3
0.6
1.1
0.1
0
0.6
0
0
0.2
0.3
1
0
0.2
1.7
4.9
0.8
0
0.8
45
0.5
0.1
1.4
1
3.9
0.7
0
1.6
0.3
0.6
0.4
0.8
1.2
0
0
0.1
1.8
0
0
0
2
0.6
0.1
0.7
0
0
0.1
0.2
1.3
1
0.1
40.8
0.2
9.7
0
6.2
0
2.3
6.1
1.5
6.2
3.6
2.1
0.8
0.1
0
0.5
0.8
0.2
6.2
0
0
0
0
0
2.1
0
0
0.1
0
0
0
0.6
1.7
4.7
2.8
41.9
1.9
2.4
0.5
21.6
0
0
0
5.6
1.8
0.1
0.7
18
678
Conclusion
x A dose of 2000 mg L-1 Tween 80 exhibited the best desorption of lindane from soil to
water phase, although its absolute value was low-to-moderate
x The bioremediation of lindane in soil could be achieved in an EBCR with similar
removals to those reported in anaerobic slurry bioreactors loaded with lindaneacclimated, methanogenic inoculum, as well as other conventional slurry
bioreactors.
x The EBCR not only provided bioremediation of a toxic, recalcitrant organochlorinated pesticide, but also supplied (as bioelectricity) ca. 20 % of the energy
required for mixing the device.
x The results of bacterial community analysis of the electrobiochemical slurry reactor,
the phylum more representative were actinobaceria, bacteroidetes, firmicutes,
proteobacteria, and synergistetes.
Thus, the EBCR is a significant step towards the green/sustainable remediation of
contaminated soils.
Acknowledgements
The authors wish to thank Mr. Rafael Hernández-Vera (GBAER, DBB CINVESTAV del
IPN), Mr. A. Rodríguez-Castellanos (Dept Chemistry), Mr. A. Barbosa-Fernández
(Mechanical Shop), Mr. J. Barrera Rojas (Dept. of Biochemistry), and Mr. Gustavo Medina
(chromatographic analyses), for their excellent technical help. CONACYT granted a
graduate scholarship to BC-P. SECITI-GDF (formerly ICYTDF) supported the research
with grant PICCO 10-28.
679
References
APHA (1992) Standard Methods for the Examination of Water and Wastewater. 18th ed. American Public
Health Association (APHA), American Water Works Association (AWWA), Water Environment
Federation (WEF). Washington DC, EEUU. 1085 pp.
Camacho-Pérez, B.; Ríos-Leal, E.; Rinderknecht-Seija, N.; Poggi-Varaldo, H.M. 2012 Enzymes involved in
the biodegradation of hexachlorocyclohexane: A mini review. Journal of Environmental Management,
95:S306-S318.
Camacho-Pérez, B.; Ríos-Leal, E.; Solorza-Feria, O.; Vazquez-Landaverde, P.A.; Barrera-Cortés, J.; PonceNoyola, M.T.; García-Mena, J.; Rinderknecht-Seijas, N.; Poggi-Varaldo, H.M. 2013. ³3HUIRUPDQFHRIDQ
Electrobiochemical Slurry Reactor for the TreatmenW RI D 6RLO &RQWDPLQDWHG ZLWK /LQGDQH´ Journal of
New Materials for Electrochemical Systems. 217-228.
Caporaso, J.G.; Kuczynski, J.; Stombaugh, J.; Bittinger, K.; Bushman, F.D.; Costello, E.K.; Fierer, N.;
Gonzalez-Peña, A.; Goodrich, J.K.; Gordon, J.I.; Huttley, G.A.; Kelley, S.T.; Knights, D.; Koenig, J.E.;
Ley, R.E.; Lozupone, C.A.; McDonald, D.; Muegge, B.D.; Pirrung, M.; Reeder, J.; Sevinsky, J.R.;
Turnbaugh, P.J.; Walters, W.A.; Widmann, J.; Yatsunenko, T.; Zaneveld, J.; Knigh, R. 2010. QIIME
allows analysis of high-throughput community sequencing data. Nature Methods, 7:335±336.
Cookson J.T. Bioremediation engineering: design and application New York, USA: McGraw-Hill Publishing
Co.; 1995.
DeSantis, T.Z.; Hugenholtz , P.; Larsen, N.; Rojas, M.; Brodie, E.L.; Keller, K.; Huber, T.; Dalevi, D.; Hu, P.;
Andersen, G.L. 2006 Greengenes, a chimera-checked 16S rRNA gene database and workbench
compatible with ARB. Appl Environ Microbiol, 72: 5069±5072.
'L *HQQDUR 3 )UDQ]HWWL $ %HVWHWWL * /DVDJQL 0 3LWHD ' &ROOLQD ( ³6OXUU\ 3KDVH
%LRUHPHGLDWLRQ RI 3$+V LQ ,QGXVWULDO /DQGILOO 6DPSOHV DW /DERUDWRU\ 6FDOH´ Waste Management.
28:1338±1345.
'X=/L+*X7³$6WDWHRIWKH$Ut Review on Microbial Fuel Cells: Apromising Technology for
Wastewater Treatment and Bioenergy´. Biotechnology Advances. 25:464±482.
Edgar, R.C. 2010 Search and clustering orders of magnitude faster than BLAST. Bioinformatics, 26 (19):
2460-2461.
Eweis JB, Ergas SJ, Chang DPY, Schroeder ED: Bioremediation Principles McGraw-Hill. USA; 1998:207236.
Fedorovich, V.; Knighton, M.C.; Pagaling, E.; Ward, F.B.; Free, A.; Goryanin, I. 2009 Novel electrochemically
active bacterium phylogenetically related to arcobacter butzleri, isolated from a microbial fuel cell.
Applied and Environmental Microbiology, 7326±733.
Fierer, N.; Hamadyc, M.; Lauberb, C.L.; Knight, R. 2008 The influence of sex, handedness, and washing on
the diversity of hand surface bacteria.
Haas, B.J.; Gevers, D.; Earl, A.M.; Feldgarden, M.; Ward, D.V.; Giannoukos, G.; Ciulla, D.; Tabbaa, D.;
Highlander, S.K.; Sodergren, E.; Methe´, B.; DeSantis, T.Z.; Petrosino, J.F.; Knight, R.; Birren, B.W.
2011 Chimeric 16S rRNA sequence formation and detection in Sanger and 454-pyrosequenced PCR
amplicons. Genome Res, 21: 494-504.
+XDQJ '< =KRX 6* &KHQ 4 =KDR % <XDQ < =KXDQJ / ³(QKDQFHG $QDHURELF
'HJUDGDWLRQ RI 2UJDQLF 3ROOXWDQWV LQ D 6RLO 0LFURELDO )XHO &HOO´. Chemical Engineering
Journal. 172: 647± 653.
Kim, H.M.; Chae, N.; Jung, J.Y.; Lee, Y.K. (2013) Isolation of facultatively anaerobic soil bacteria from Ny-A°
lesund, Svalbard. Polar Biol 36:787±796.
/L <) ³*OREDO 7HFKQLFDO +H[DFKORURF\FORKH[DQH 8VDJH and its Contamination Consequences in
(QYLURQPHQWIURPWR´ The Science of the Total Environment. 232:123±160.
Li, Y.F.; Cai, D.J.; Singh, A. 1998. Technical Hexachlorocyclohexane use trends in China and their Impact on
WKH(QYLURQPHQW´ Archives of Environmental Contamination and Toxicology. 35:688±697.
Liu, H.; Ramnarayanan, R.; Logan, B.E. 2004 ³Production of Electricity during Wastewater Treatment using a
Single Chamber Microbial Fuel Cell´. Environmental Science & Technology. 38(7):2281±2285.
Logan, B. E.; Hamelers, B.; Rozendal, R.; Schroder, U.; Keller, J.; Freguia, S.; Aelterman, P.; Verstraete, W.;
Rabaey, K. 2006 ³Microbial Fuel Cells: Methodology and Technology´. Environmental Science &
Technology. 40(17): 5181-5192.
Lu, N.; Zhou, S.G.; Zhuang, L.; Zhnag, J.T.; Ni, J.R. 2009 ³Electricity Generation from Starch Processing
Wastewater using Microbial Fuel Cell Technology´. Biochemical Engineering Journal. 43: 246-251.
680
Macary, F.; Morin, S.; Jean-Luc Probst, J.L.; Saudubray, F. 2014 ³A Multi-scale Method to assess Pesticide
&RQWDPLQDWLRQ5LVNVLQ$JULFXOWXUDO:DWHUVKHGVÉcological Indicators. 36: 624± 639.
0RUULV -0 -LQ 6 ³)HDVLELOLW\ RI XVLQJ 0LFURELDO )XHO &HOO 7HFKQRORJ\ LQ %LRUHPHGLDWLRQ RI
+\GURFDUERQVLQ*URXQGZDWHU´Journal of Environmental Science and Health Part A. 43:18-23.
Mulder CPH, Bazeley-White E, Dimitrakopoulos PG, Scherer-Lorenzen AHM, Schmid B. 2004 Species
evenness and productivity in experimental plant communities 104:50-63.
0XOOLJDQ &1 <RQJ 51 *LEVV %) ³6XUIDFWDQW-enhanced Remediation of Contaminated Soil: a
5HYLHZÉngineering Geology. 60:371-380.
Okeke, B.& 6LGGLTXH 7 $UEHVWDLQ 0& )UDQNHQEHUJHU :7 ³%LRGHJUDGDWLRQ RI Ȗ+H[DFKORURF\FORKH[DQH OLQGDQH DQG Į-Hexachlorocyclohexane in Water and a Soil Slurry by a
3DQGRUDHD6SHFLHV´ Journal of Agricultural and Food Chemistry. 50:2548-2555.
Ortega-Martínez, A.; Juárez-López, K.; Solorza-Feria, O.; Ponce-Noyola, M.T.; Ríos-Leal, E.; RinderknechtSeijas, N.F.; Poggi-9DUDOGR+0³3DUDOOHO&RQQHFWLRQDQG6DQGZLFK(OHFWURGHVORZHUWKH,QWHUQDO
5HVLVWDQFHLQD0LFURELDO)XHO&HOO´Journal of New Materials for Electrochemical Systems. 15(3):187194.
Ortega-Martínez, A.C.; Juárez-López K.; Solorza-Feria, O.; Ponce-Noyola, M.T.; Galindez-Mayer, J.;
Rinderknecht-Seijas, N.; Poggi-9DUDOGR+0³Analysis of microbial diversity of inocula used in a
five-IDFH SDUDOOHOHSLSHG DQG VWDQGDUG PLFURELDO IXHO FHOOV´ International journal of hydrogen energy
38:12589-12599.
3DNQLNDU .0 1DJSDO 9 3HWKNDU $9 5DMZDGH -0 ³'HJUDGDWLRQ RI /LQGDQH IURP $TXHRXV
Solutions using Iron Sul¿de NanoSDUWLFOHV VWDELOL]HG E\ %LRSRO\PHUV´ Science and Technology of
Advanced Materials, 6:370±374.
3DQW'9DQ%RJDHUW*'LHOV/9DQEURHNKRYHQ.³$5HYLHZRIWKH6XEVWUDWHVXVHGLQ0LFURELDO
Fuel Cells (0)&VIRU6XVWDLQDEOH(QHUJ\3URGXFWLRQ´Bioresource Technology. 101(6): 1533±1543.
Poggi-Varaldo, H.M.; Carmona Martínez, A.; Vázquez-Larios A.L.; Solorza-)HULD2³(IIHFWRI,QRFXOXP
Type on the Performance of a Microbial Fuel Cell Fed with
Spent Organic Extracts from
Hydrogenogenic FerPHQWDWLRQRI2UJDQLF6ROLG:DVWHV´ Journal of New Materials for Electrochemical
Systems. 12:49-54.
Poggi-Varaldo, H.M.; Trejo-Espino, J.; Fernandez-Villagomez, G.; Esparza-Garcia, F.; S. Caffarel-Méndez,
S.; Rinderknecht-Seijas, N. 1999 ³Quality of Anaerobic Compost from Paper Mill and Municipal Solid
Wastes for Soil Amendment´. Water Science Technology. 40(11-12):179-186.
3UDNDVK 2 6XDU 0 5DLQD 9 'RJUD & 3DO 5 /DO 5 ³5HVLGXH RI +H[DFKORURF\FORKH[DQH
Isomers in Soil and Water samples from Delhi and adjoining Areas. Current Science. 87 (1):73±78.
4XLQWHUR-&0RUHLUD07)HLMRR*/HPD-0D³Effect of Surfactants on the Soil Desorption of
+H[DFKORURF\FORKH[DQH +&+ ,VRPHUV DQG WKHLU $QDHURELF %LRGHJUDGDWLRQ´ Journal of Chemical
Technology and Biotechnology. 80:1005±1015.
4XLQWHUR -& 0RUHLUD 07 )HLMRR * /HPD -0 E ³$QDHURELF 'HJUDGDWLRQ RI
+H[DFKORURF\FORKH[DQH,VRPHUVLQ/LTXLGDQG6RLO6OXUU\6\VWHPVĆhemosphere. 61:528-536.
Quintero, J.C.; Moreira, M.T.; Lema, J.M.; Feijoo, G. 2006. An anaerobic bioreactor allows the efficient
degradation of HCH isomers in soil slurry, Chemosphere, 63: 1005±1013
5H]DHL)5LFKDUG7//RJDQ%(³$QDO\VLVRI&KLWLQSDUWLFOHVL]HRQ Maximum Power Generation,
Power Longevity, and Coulombic Efficiency in Solid-substrate Microbial Fuel Cells. Journal of Power
Sources, 192:304±309.
Robles-González, I.V.; Fava, F.; Poggi-9DUDOGR +0 ³$ 5HYLHZ RQ 6OXUU\ %LRUHDFWRUV IRU
Bioremediation RI6RLOVDQG6HGLPHQWV´Microbial Cell Factories. 7: 5. doi:10.1186/1475-2859-7-5
Robles-González, I.V.; Ríos-Leal, E.; Sastre-Conde, I.; Fava, F.; Rinderknecht-Seijas, N.; Poggi-Varaldo,
+0³6OXUU\ELRUHDFWRUVZLWKVLPXOWDQHRXVHOHFWURQDFFHSWRUVIRUELRUHPHGLDWLRQRIDQDJULFXOWXUDO
VRLOSROOXWHGZLWKOLQGDQH´Process Biochemistry. 47(11):1640-1648.
Robles-González, I; Ríos-Leal, E.; Ferrera-Cerrato, R.; Esparza-García, F.; Rinderkenecht-Seijas, N.; Poggi9DUDOGR +0 ³%LRUHPHGLDWLRQ RI D PLQHUDO VRLO ZLWK KLJK FRQWHQWV RI FOD\ DQG RUJDQLF PDWWHU
contaminated with herbicide 2,4-dichorophenoxyacetic acid using slurry bioreactors: Effect of electron
DFFHSWRUDQGVXSSOHPHQWDWLRQZLWKDQRUJDQLFFDUERQVRXUFH´Procces Biochemistry. 41:1951-1960.
Sathish-Kumar, K.; Solorza-Feria, O.; Vázquez-Huerta, G.; Luna-Arias, J.P.; Poggi-Varaldo, H.M. 2012
³(OHFWULFDO 6WUHVV-directed Evolution of Biocatalysts Community Sampled from a Sodic-saline Soil for
0LFURELDO)XHO&HOOV´Journal of New Materials for Electrochemical Systems. 15 (3):181-186.
681
Sattler, C.; Kächele, H.; Gernot Verch, G ³$VVHVVLQJ WKH LQWHQVLW\ RI SHVWLFLGH XVH LQ DJULFXOWXUH´
Agriculture Ecosystems and Environment. 119: 299±304.
Shanon CE. (1948) A mathematical theory of communication. Bell System Technical Journal 27:379-423.
Valdez-Vázquez, I.; Ríos-Leal, E.; Muñoz-Páez, K.M.; Carmona-Martínez, A.; Poggi-Varaldo, H.M. 2006
³Effect of inhibition treatment, type of inocula, and incubation temperature on batch H 2 production from
organic solid waste´ Biotechnology and Bioengineering. 95:342-349.
Varo-Arguello, W.E.; Camacho-Pérez, B.; Ríos-Leal, E.; Vázquez-Landaverde, P.A.; Ponce-Noyola, M.T.;
Barrera-Cortés, J.; Sastre-Conde, I.; Rindernknecht-Seijas, N.F.; Poggi-Varaldo, H.M. 2012 Triphasic
slurry bioreactors for the bioremediation of lindane-impacted soil under aerobic and anaerobic
conditions. Environmental Engineering and Management Journal. 11(10):1811-1823.
Vázquez-Baeza, Y.; Pirrung, M.; Gonzalez, A.; Knight, R. 2013. EMPeror: a tool for visualizing highthroughput microbial community data. Gigascience, 2(1):16
Vázquez-Larios, A.L.; Solorza-Feria, O.; Vázquez-Huerta, G.; Esparza-García, F.; Ríos-Leal, E.;
Rinderknecht-Seijas, N.; Poggi-Varaldo, H.M. 2010 ³$1HZ'HVLJQ,PSURYHV3HUIRUPDQFHRID6LQJOH
&KDPEHU0LFURELDO)XHO&HOO´Journal of New Materials for Electrochemical Systems. 13: 219-226.
Vázquez±Larios, A.L.; Solorza-Feria, O.; Vázquez-Huerta, G.; Esparza-García, F.; Rinderknecht-Seijas, N.;
Poggi-9DUDOGR+0³(IIHFWVRI$UFKLWHFWXUDOFKDQJHVDQG,QRFXOXPW\SHRQ,QWHUQDO5HVLVWDQFHRI
a Microbial Fuel Cell designed for the Treatment of Leachates from the Dark Hydrogenogenic
)HUPHQWDWLRQRI2UJDQLF6ROLG:DVWHVÍnternational Journal of Hydrogen Energy. 36:6199-6209.
Venkata-Mohan, S., Sirisha, K.; Sreenivasa-5DR 5 6DUPD 31 ³%LRVOXUU\ SKDVH 5HPHGLDWLRQ RI
Chlorpyrifos Contaminated Soil: Process Evaluation and Optimization by Taguchi Design of
([SHULPHQWDO'2(0HWKRGRORJ\Écotoxicology and Environmental Safety. 68:252±262.
:DQJ ; &DL = =KRX 4 =KDQJ = &KHQ & ³%LRHOHFWURFKHPLFDO VWLPXODWLRQ RI 3HWUROHXP
Hydrocarbon Degradation in Saline Soil using u-WXEH 0LFURELDO )XHO &HOOV´ Biotechnology and
Bioengineering. 109(2):426-433.
:KLWHOH\$6-HQNLQV6:DLWH,.UHVRMH13D\QH+0XOODQ%$OOFRFN52¶'RQQHOO$
Microbial 16S rRNA Ion Tag and community metagenome sequencing using the Ion Torrent (PGM)
Platform. J Microbiol Methods, 91: 80±88.
:LOOHWW ./ 8OULFK (0 +LWHV 5$ ³'LIIHUHQWLDO 7R[LFLW\ DQG (QYLURQPHQWDO IDWHV RI
+H[DFKORURF\FORKH[DQH,VRPHUVÉnvironmental Science and Technolog., 32:2197±2207.
Yan, Z.; Song, N.; Cai, H.; Tay, J.H.; Jiang, H. 2012 ³Enhanced Degradation of Phenanthrene and
Pyrene in Freshwater Sediments by combined Employment of Sediment Microbial Fuel Cell
and Amorphous Ferric Hydroxide´. Journal of Hazardous Materials, 199±200: 217± 225.
Zaybak, Z.; Pisciotta, J.M.; Tokash, J.C.; Logan, B.E. (2013) Enhanced start-up of anaerobic facultatively
autotrophic biocathodesin bioelectrochemical systems. Journal of Biotechnology 168: 478± 485
682
Abbreviations and Acronyms
sCOD
EBCR
EEBCR
Êharvested
Êmixing
HCH
IEBQR
Max
MFC
NOM
PAn
Pave
PEBCR
PV
Rint
SMFC
SB
SR
soluble Chemical oxygen demand
Electrobiochemical slurry reactor
Voltage
Energy per tonne of soil associated to 30 days of treatment in an EBCR
Energy per tonne of soil required for mixing during 30 days of treatment
Hexachlorocyclohexane
Current intensity
Maximum
Microbial Fuel Cell
Natural organic matter
Surface area power density
average power
Power delivered
Volumetric power
Internal resistance
Soil microbial fuel cell technology
Slurry reactors
Sulphate reducing
Greek characters
Kcoul
Coulombic efficiency
KCOD
Removal efficiency of organic matter as chemical oxygen demand
KLindane
Removal efficiency of lindane
683
CHAPTER 6.12 POTENTIAL CRUDE OIL DEGRADATION BY A MICROBIAL
CONSORTIUM ISOLATED FROM THE GULF OF MEXICO
N. Ulises García-Cruz* (1); Rosa Canul-Puc (2); Juan Sanchez-Avila (1);
Gerardo Gold-Bouchot (1); M. Leopoldina Aguirre-Macedo (1)
(1) CINVESTAV- IPN, Yucatan, México; Dept. of Marine Resources.
(2) ITESCAM, Campeche, México;
Resumen
El petróleo crudo se encuentra entre los principales contaminantes de los sistemas
marinos. Este puede ser liberado al ambiente de manera accidental o deliberadamente
llevando a serios problemas de contaminación resultando en perturbaciones tanto en los
componentes biótico y abiótico de los ecosistemas. La bioaumentación emplea el
potencial de biodegradación de los microorganismos que pueden ser usados para
recuperar sitios contaminados. Por otra parte se a observado que un simple
microorganismos no tiene la capacidad de metabolizar los diferentes compuestos del
petróleo, sin embargo un consorcio el cual está compuesto de diferentes especies de
bacterias tiene la capacidad de degradar el petróleo significativamente. El objetivo de este
estudio fue aislar un consorcio microbiano a partir de sedimentos marinos capaz de
degradar los diferentes componentes del petróleo. Las muestras de los sedimentos
marinos fueron colectadas de diferentes localidades en el Golfo de México para
posteriormente aislar microorganismos que consumieran el petróleo.
Palabras Clave: Petróleo crudo, Consorcios microbianos, Golfo de México
Abstract
Crude oil can be deliberately or accidentally released into the environment leading to
serious pollution problems resulting on disturbances of biotic and abiotic components of
ecosystems. The bioaugmentation employs the biodegradative potentials of
microorganisms that can be used to recover contaminated sites. No single microorganisms
has the enzymatic ability to metabolize the different compounds of oil, nevertheless a
consortium composed of many different bacterial species can degrade oil significantly. The
aim of this study was to isolate a microbial consortium from marine sediments capable to
biodegrade oil constituents. Marine sediment samples collected from different localities in
the Gulf of Mexico were screened for microorganisms capability of oil degradation.
---------*Author for all correspondence: [email protected]
684
Introducción
El petróleo es una de las principales fuentes de energía en el mundo, por tal motivo en las
ultimas décadas se ha intensificado su extracción y transporte en los mares, lo que a
favorecido su dispersión, así como también un incremento en las concentraciones de
hidrocarburos, por otra parte también se ha incrementado el riesgo de contaminación por
derrames a gran escala en los océanos.
En el caso del Golfo de México, al ser una de las regiones marítimas que concentra
una gran actividad petrolera hace que esté propensa a derrames y a la contaminación por
hidrocarburos, por otra parte estos fenómenos se pueden ver favorecidos por el sistema
de corrientes marinas que circulan por la zona incrementando el impacto ambiental de
estos compuestos en el ambiente.
Los efectos de la contaminación ambiental por hidrocarburos son variados tanto por
la toxicidad sobre los individuos como por los daños ecológicos que acarrean, a su vez la
persistencia de estos contaminantes en el ambiente se ve influenciada por varios factores
como son: la naturaleza del contaminante, su concentración y su capacidad de interactuar
con las condiciones químicas, geológicas, físicas y biológicas (De Haro-Cristian, 2010).
Como se puede observar actualmente en las playas de Ciudad del Carmen (México) en
donde se puede encontrar hasta la fecha después de 30 años restos de petróleo que se
derramo durante el accidente del pozo Ixtoc en el Golfo de México (Schrope, 2010).
Actualmente se cuentan con tratamientos físico-químicos y biológicos para mitigar la
contaminación por petróleo.
Entre los primeros se encuentra el uso de dispersantes químicos, pero hasta la fecha
se encuentra en debate el uso de estos reactivos debido que por la magnitud de los
accidentes o derrames se requiere de una gran cantidad lo que se a reportado que puede
incluso incrementar la toxicidad de los hidrocarburos (Mascarelli, 2010), en contra parte se
a observado que los tratamientos biológicos presentan una alternativa para la remoción y
degradación de petróleo, al ser estos más compatibles con el ambiente, entre estos
tratamientos se encuentra la bioaumentación, que es la aplicación de microorganismo
previamente aislados y adaptados a los contaminantes, la cual se considera una de las
técnicas más efectivas, debido a que las bacterias se pueden aplicar para remediar
químicos específicos en condiciones ambientales especificas (Salleh et al., 2003), .
En este sentido se a observado que el uso de consorcios microbianos incrementan
la capacidad de degradación de contaminantes debido a las interacciones que se llevan a
cabo entre las bacterias presentes, como lo reportan Morris et al. (2013), debido a que
una sola especie de bacteria no cuenta con la maquinaria metabólica necesaria para
degradar los diferentes compuestos. En el caso del petróleo al ser una mezcla compleja
de hidrocarburos que van desde alifáticos lineales hasta hidrocarburos poliaromáticos, se
requiere de diferentes bacterias para degradar esa amplia variedad de compuestos. Por
otra parte se han reportado diferentes consorcios microbianos capaces de utilizar el
petróleo como donador de electrones y fuente de carbono, también cabe mencionar que
se han reportado una gran variedad de bacterias (incluyendo algunas cianobacterias),
diversos mohos, levaduras así como diversas algas verdes, que pueden degradar el
petróleo aerobicamente y que esto a su vez, va a depender de la velocidad de crecimiento
y de las condiciones ambientales como la temperatura, y nutrientes inorgánicos como son
el nitrógeno y fosforo (Madigan et al., 2013).
Por lo que el objetivo de este trabajo fue aislar un consorcio microbiano a partir de
sedimentos marinos de diferentes localidades del Golfo de México capaz de degradar los
685
principales componentes del petróleo (alifáticos, aromáticos y marcadores moleculares)
variando la relación C/N.
Material y Métodos
Sedimentos. Se obtuvieron diferentes sedimentos provenientes de la región suroeste del
Golfo de México (Figura 1). En cada estación el sedimento fue obtenido mediante una
draga tipo Smith-McIntyre de acero (0.3 x 0.3 x 0.2 m, 80 kg de peso), con una
penetración máxima en sedimento de 20cm y con un nucleador de caja Hessler-Sandia
MK-II de 40 x 40 cm. Los cuales fueron operados por gravedad en caída vertical.
FIGURA 1. Red de muestreo de la región suroeste del Golfo de México.
Inmediatamente de sacar la draga y el nucleador, se procedió a sacar la muestra de
la parte central de los dispositivos, antes de tomar la muestra, se retiró aproximadamente
5 cm de sedimento de la superficie, posteriormente se tomó la muestra usando bolsas
estériles las cuales se sellaron inmediatamente y se mantuvieron en refrigeración (4°C)
hasta su uso en el laboratorio.
686
Aislamiento y cuantificación de bacterias. Se tomó una alícuota de cada una de las
diluciones y se inocularon tres series de tubos con 5 ml cada uno con caldo de BushnellHass (Difco) más la fuente única de carbono (petróleo crudo) y Resazurina (100 µl) como
indicador de oxidación. Posteriormente se mezclaron las bacterias de diferentes puntos de
muestreos para incrementar la capacidad metabólica. Se obtuvieron 3 consorcios
microbianos R1, R2 y R3. De estos primero el R1 y R3 se probaron utilizando un medio
con agua dulce a una C/N=8, posteriormente se evaluaron los tres sedimentos a C/N=10,
salvo con la excepción de que el R2 se evaluó con el medio preparado en agua de mar
real.
Para la cuantificación de las bacterias hidrocarbonoclastas (BHC) se utilizó la
metodología antes mencionada, se diluyeron 5 mL de los reactores en tubos de ensaye
dejando incubar por 24 horas, posteriormente se cuantificaron por el numero más
probable. Para la cuantificación de las bacterias heterótrofas (BHT) se tomaron 5 mL en
una solución Ringer en un copro, se agitaron por 15 minutos en un termociclador MaxQ
2000 y se esperó otros 15 minutos de reposo, posteriormente del frasco se extrajeron
ȝO GH PXHVWUD \ VH VHPEUDURQ ODV EDFWHULDV HQ PHGLR 5$ PDUFD 'LIFR MR), por
duplicado dejando en incubación por 24 horas para luego realizar un conteo utilizando la
técnica en placa de unidades formadoras de colonias.
Identificación de bacterias. Para la identificación de las bacterias se utilizó el equipo
BBL Crystal, para tal motivo se prepararon las muestras de la siguiente manera: se usó
1ml de muestra de cada uno de los consorcios (B1 y B3), estas muestras se diluyeron en
9ml de Caldo Triptona Soja en incubación por 24 horas para poder sembrar por estría en
placa en cajas de petri con agar de cultivos Trisulfato citrato billis sacarosa (TCBS), MEnterococcus (ME), Cetrimide, MacConkey, Soja y Tripticaseina (TSA) (DifcoMR) y se
dejaron en incubación por 24 horas, posteriormente los cultivos que crecieron se
resembraron en cajas petri con agar TSA dejándolo en incubación otras 24 horas para la
purificación de la bacteria, y poder resembrar en suspensión para su identificación en el
equipo.
Operación de los reactores para la degradación de petróleo. Posteriormente de
obtener los consorcios, se inocularon con estos mismos, botellas (Kimax) de 1L, las
cuales contenían 0.9 L de medio mineral Visser (1995) preparado con agua destilada. Las
botellas fueron esterilizadas antes de inocularlas. Las relaciones carbono nitrógeno que
se evaluaron fueron de 8 y 10, utilizando una concentración de petróleo crudo ligero
(plataforma Akal-J, PEMEX) de 284.4 y 395.8 mg/L respectivamente el cual fue filtrado
(0.2 µm, CorningMR) y se utilizó como única fuente de carbono y energía. Los frascos
fueron operados por un periodo de 30 días a 24°C y agitados aprox. a 400 rpm con una
barra magnética.
Cuantificación de hidrocarburos. Para la cuantificación de los hidrocarburos se tomó
una alícuota de 20 mL en el t=0 y al t=30 de los reactores, posteriormente se
centrifugaron a 3000 rpm por 10 min, se retiró el sobrenadante y se hizo una extracción
con diclorometano (1:1) para posteriormente analizar los diferentes componentes del
petróleo.
La fracción de alifáticos se analizó por cromatografía de gases con un detector de
ionización de flama (GC-FID) de la marca AGILENT modelo 7890A.
687
La fracción de hidrocarburos policíclicos aromáticos y de biomarcadores se
analizaron por cromatografía de gases acoplada a un detector de espectrometría de
masas marca Perkin Elmer modelo Clarus 500.
Para la fracción de alifáticos, las concentraciones se calcularon usando una curva de
calibración de 7 puntos (1-20 µg/mL) de una mezcla de n-C8 a n-C40, pristano y fitano, en
un rango lineal (R2=0.9917 - 0.9932). La mezcla compleja no resuelta (UCM por sus
siglas en Inglés) se cuantificó midiendo el área del cromatograma entre la línea base.
Para los hidrocarburos policíclicos aromáticos (PAH´s) se empleó una curva de 7
puntos (0.01 - 2.5 µg/mL) de una mezcla de 33 estándares en un rango lineal (R2=0.9975
± 0.9993).
Para los biomarcadores, las concentraciones se calcularon usando una curva de
calibración de 7 puntos (0.0025- JP/ GH XQD PH]FOD GH ȕȕ +RSDQR
5 \&5ĮĮĮ(WLOFROHVWDQR5 Biosurfactantes. Los biosurfactantes se detectaron mediante la técnica de colapso de
gota reportada por Bodour y Miller-Maier (1998).
Resultados
Degradación de hidrocarburos. En la Figura 2 se muestran los resultados de los HTP
de los reactores a C/N de 8 y 10.
100
R1 (agua dulce)
80
C/N=8 60
C/N=10
40
HTP (%)
20
0
100
R3 (agua dulce)
80
C/N=8
40
C/N=10
60
20
0
0
8
24
32
Tiempo (d)
FIGURA 2. Porcentaje de degradación de HTP después de 30 días de operación
688
Como se puede observar en la Figura 2 a una C/N=8, se encontró un consumo muy
similar entre ambos reactores de 70 y 74% para los reactores R1 y R3 respectivamente,
mientras que para la relación C/N=10 se observó que el consumo se incrementó después
de 30 días de reacción alcanzando un 82% para el reactor R1 y 88% para el reactor R3,
estos resultados son similares a los reportado, en donde se a observado que a una
relación C:N cercana a 12:1 se observa una mayor degradación (Wang et al., 2011).
En base a estos resultados se corrió una cinética con el consorcio R2 a una relación
C/N=10 con agua de mar y se comparó con los otros reactores, los resultados se
muestran en la Figura 3.
100
C/N=10
R1 (agua ducle)
R3 (agua dulce)
R2 (agua de mar)
60
HTP (%)
80
40
20
0
0
8
24
32
Tiempo (d)
FIGURA 3. Comparación en el porcentaje de degradación de los tres reactores a una
C/N=10.
Como se puede observar en la Figura 3, el reactor R2 fue el que a una relación
C/N=10 tuvo el mejor porcentaje de consumo al alcanzar un 92% en comparación a los
reactores R1 y R3, a pesar de que se preparó el medio con agua de mar la cual tiene una
concentración de sal cercana al 3.6% de NaCl, estos resultados son diferentes a lo
reportado en la literatura en donde se ha observado que a concentraciones entre 2.8 y
3.3% NaCl se inhibe o disminuye el consumo de hidrocarburos (Leahy y Colwell, 1990).
Resultados similares se observaron en la degradación de HAP´s y los
biomarcadores (Figura 4)
689
100
C/N=10
R1
R3
R2
60
40
HAP´s (%)
80
20
0
80
60
Biomarcadores (%)
100
40
20
0
0
8
24
32
Tiempo (d)
FIGURA 4. Comparación entre los reactores en la degradación de HAP´s y
Biomarcadores.
Cuantificación e identificación de bacterias. Con respecto a la relación entre las
bacterias hidrocarbonoclastas y heterótrofas, no se observó que el suplementar petróleo
como única fuente de carbono se incrementara la concentración de BHC, solo se encontró
que el incremento en la relación C/N=10 favoreció que se incrementara en un 2% aprox.
La concentración de estas bacterias en los tres reactores.
Por otra parte el equipo BBL Crystal identifico 6 bacterias las cuales se muestran en
la Tabla 1
De las bacterias identificadas por el equipo se encuentraron 5 de las cuales están
reportadas que degradan algún componente del petróleo, a excepción de Serratia
odorifera la cual no está relacionada con la degradación de hidrocarburos, se ha
reportado en la corrosión de oleoductos.
Biosurfactantes. Se observó que en los tres reactores a las diferentes relaciones C/N
hubo presencia de biosurfactantes.
690
TABLA 7. Bacterias identificadas con el equipo BBL Crystal.
BACTERIAS
FACTOR DE
CONFIANZA EQUIPO:
"BBL CRYSTAL"
65%
Bhatia y Sharma,
2010
78%
Ojo, 2006
89%
Bachoon et al., 2001
99%
Wongsa et al., 2004
Klebsiella oxytoca
92%
Chamkha et al., 2011
Serratia odorifera
94%
Rajasekar et al.,
2007
Pantoea
agglomerans
Enterobacter
aerogenes
Enterobacter
cloacae
Serratia
marcescens
Conclusiones
Los consorcios degradaron cerca del 90% de petróleo en agua destilada y salada. Las
bacterias no perdieron la capacidad metabólica de degradar petróleo a pesar de que
fueron aisladas usando medios con agua destilada. Todos los consorcios fueron capaces
de producir biosurfactantes. A pesar de solo alimentar petróleo no se vio favorecida las
bacterias hidrocarbonoclastas.
Bibliografia
Bchoon, D., S., Hodson, R., E. y Araujo, R. (2001). Microbial community assessment in oil-impacted salt
marsh sediment microcosms by traditional and nucleic acid-based indices. Journal of Microbiological
Methods. 46(1): 37-49.
Bhatia, S. and Sharma, D. K., (2010). Biodesulfurization of dibenzothiophene, its alkylated derivatives and
crude oil by a newly isolated strain Pantoea agglomerans D23W3, Biochem. Eng. J. 50: 104-109.
Bodour, A. A., and R. M. Miller-Maier. 1998. Application of a modified drop-collapse technique for surfactant
quantitation and screening of biosurfactant-producing microorganisms. J. Microbiol. Methods 32:273-280.
Chamkha, M., Trabelsi, Y., Mnif, S. y Sayadi, S. (2011). Isolation and characterization of Klebsiella oxytoca
strain degrading crude oil from a Tunisian off-shore oil field. Journal of Basic Microbiology. 51(6): 580589.
Leahy, J., G. y Colwell, R., R. (1990). Microbial degradation of hydrocarbons in the environmental.
Microbiological Reviews. 54(3):305-315.
Madigan, T., M., Martinko, M., J. y Jack, P. Brock. Biología de los Microorganismos. Décima Edición. Pag.
666 ISBN: 84-205-3679-2, 2003.
Mascarelli, A. (2010). Debate grows over impact of dispersed oil. Nature. doi:10.1038.
Ojo, O.A. (2006). Petroleum-hydrocarbon utilization by native bacterial population from a wastewater canal
Southwest Nigeria. African Journal of Biotechnology. 5:333±337
Rajasekar, A., Ponmariappan, S., Maruthamuthu, S., y Palaniswamy, N. (2007). Bacterial Degradation and
Corrosion of Naphtha in Transporting Pipeline. Current Microbiology. 55(5): 374-381.
Schrope, M. (2010). The lost legacy of the last great oil spill. Nature. 466: 304-305.
691
Salleh, A., B., Ghazali, F.,M., Rahman, R., N., Z., A., y Basri., M., (2003). Bioremediation of petroleum
hydrocarbon pollution. Indian Journal of Biotechnology. 2: 411-425.
Wang, Q., Zhang, S., Li, Y y Klassen, W. (2011). Potential approaches to improving biodegradation of
hydrocarbons for bioremediation of crude oil pollution. Journal of Environmental Protection. 2:47-55.
Wongsa, P., Tanaka, M., Ueno, A., Hasanuzzaman, M., Yumoto, I. y Okuyama, H. (2004). Isolation and
Characterization of Novel Strains of Pseudomonas aeruginosa and Serratia marcescens Possessing High
Efficiency to Degrade Gasoline, Kerosene, Diesel Oil, and Lubricating Oil. Current Microbiology. 49(6):
415-422.
692
CHAPTER 6.13. EVALUATION OF HEXADECANE DEGRADATION USING DIFFERENT
MARINE SEDIMENTS FROM RIVERS AND LAGOONS OF THE GULF OF MEXICO
N. Ulises García-Cruz*(1); Juan Sanchez-Avila (1); David Valdés-Lozano (1);
Gerardo Gold-Bouchot (1); Leopoldina Aguirre-Macedo (1).
(1) CINVESTAV- IPN, Mérida, Yucatan.
Resumen
Se evaluó un total de diez sedimentos para calcular la velocidad de degradación aerobia
de hexadecano (HXD), seis provenientes de ríos (Gonzalez; RG, Coatzacoalcos; RC,
Grijalva; RGR, San Pedro-San Pablo; RSS, Tonalá; RT, y Papaloapan; RP.), tres de
lagunas (Mecoacan; LM, Ostion; LO, y Champoton; LC) y uno proveniente de un difusor
de aguas de una refinería (PCM), los resultados se correlacionaron con los parámetros
fisicoquímicos de cada zona. Los sedimentos que presentaron una mayor velocidad de
degradación fueron el RG (39.8 mgHXD/L·d) y LC (34.1 mgHXD/L·d) con una eficiencia
de 89 y 86% respectivamente en contraparte los sedimentos que tuvieron el menor
consumo fue el RC (5.0 mgHXD/L·d), RGR (8.6 mgHXD/L·d) y PCM (9.3 mgHXD/L·d) con
una eficiencia del 47, 81 y 49% respectivamente. Al correlacionar estos datos con los
parámetros físico-químicos, no se encontró una relación directa con un parámetro en
específico. Estos resultados mostraron que el consumo o inhibición del HXD va a
depender de la combinación de varios factores que van a favorecer o a inhibir el consumo
de HXD.
Palabras Clave: Hexadecano, Biodegradación, Sedimentos, Golfo de México
Abstrac
The goal of this research was to evaluated the effect of environmental factors in the
biodegradation of hexadecane (HXD). Ten marine sediments were used: six from rivers
(Gonzalez; RG, Coatzacoalcos; RC , Grijalva; RGR, San Pedro-San Pablo; RSS, Tonalá;
RT, and Papaloapan; RP), three from lagoons (Mecoacan; LM, Ostion; LO, and
Champoton; LC) and one marine outfall (PCM). All samples were taken from the river or
lagoon delta, except for marine outfall, the sample was taken at the end of the pipeline in
the sea. The results were correlated with physico-chemical parameters of the sites. The
highest rates of biodegradation were observed in sediments of RG (39.8 mgHXD/L·d) and
LC (34.1 mgHXD/L·d) with an efficiency of 89 and 86% respectively and the lowest
consumption rate was found in Coatzacoalcos river with 5 mg/L·d and an efficiency of
47%. In the case of Ostion lagoon and Grijalva river, a low consumption rate was observed
(12.5 and 8.6 mgHXD/L·d), nevertheless both sediments reached an efficiency around
80%. If not possible to find a correlation between consumption rate and the physicchemical parameters. These results showed that consumption rate variability depends of
the biotics and abiotics parameter of the sites.
-----------------*Author for correspondence: [email protected]
693
Introducción
En el sur del Golfo de México (Campeche, Tabasco y Veracruz) es donde se concentra la
mayor actividad de la industria petrolera para México en esta zona, debido a la extracción
y transporte (García-Cruz y Aguirre-Macedo, 2014) estas actividades aunadas al sistema
de corrientes marítimas que confluyen ahí hace que los ríos y lagunas de esta zona estén
propensos a ser contaminados por hidrocarburos los cuales pueden provenir de los
barcos que transitan la zona, por las actividades de exploración y descarga (MartínezLópez y Parés-Sierra, 1998) y en el peor de los casos, por un derrame accidental como se
observó durante el derrame de petróleo del pozo petrolero Ixtoc. El petróleo es uno de los
compuestos más contaminantes a nivel mundial en los medios marinos, ya que contiene
moléculas que representan un grave peligro para el ambiente al inhibir los procesos
metabólicos de los organismos interrumpiendo con esto las cadenas tróficas (Doyle et al.,
2008).
En los sistemas costeros este efecto se puede magnificar debido a que los
hidrocarburos se pueden adsorber en los sedimentos lo que provoca efectos a corto y
largo plazo (Seymour y Geyer, 1992). Por lo que se requiere de estudios de los
sedimentos de las áreas en riesgo para evaluar su capacidad de transformar los
hidrocarburos. Existen diferentes trabajos en donde se han estudiado las comunidades
microbianas para ver la degradación de hidrocarburos, presentes en sedimentos marinos
(Yanggou et al., 2014) y de agua dulce (Al-Wasafi et al., 2014), en donde se han podido
detectar bacterias que tienen la capacidad metabólica para degradar diferentes tipos de
hidrocarburos, entre los que se encuentra el HXD, este compuesto es uno de los
principales componentes del petróleo razón por la cual es utilizado como molécula para
estudiar la capacidad de metabolizar hidrocarburos (Chenier et al. 2003).
En el caso de la zona sur del Golfo de México son escasos los estudios que se han
llevado a cabo para conocer la capacidad de los microorganismos que tienen los
sedimentos para metabolizar los hidrocarburos, la mayoría de los estudios se han llevado
a cabo en la zona norte por lo que el objetivo de este trabajo fue evaluar diez sedimentos
provenientes de ríos y lagunas de la zona sur del Golfo de México en la degradación de
HXD.
Material y Métodos
Sedimentos. Los sedimentos (Figura 1) fueron tomados utilizando bolsas estériles las
cuales se sellaron, inmediatamente y almacenadas en refrigeración (4°C) hasta su uso en
laboratorio. Los sedimentos fueron muestreados 5 cm por debajo de la superficie, se tomó
un promedio de 500 mL de cada uno,
694
FIGURA 5. Mapa de los sitios de muestreo de la región sur del Golfo de México.
Parámetros físico-químicos. El oxígeno disuelto y el potencial-redox se determinaron
con un oximetro YSI5000, el pH se midió con un potenciómetro. El dióxido de carbono
total se determinó con el pH inicial y la alcalinidad total, la turbidez se midió con un
turbidimetro. Los resultados se muestran en la Tabla1.
Cuantificación de bacterias hidrocarbonoclastas (HCB). Primero se hizo una
propagación utilizando un medio general en placa, posteriormente se sembraron en tubo
de ensaye utilizando medio Bushnell-Haas (Difcotm) utilizando petróleo como única fuente
de carbono, posteriormente se cuantificaron siguiendo la metodología del número más
probable (Tabla 1).
Cuantificación de hidrocarburos policíclicos aromáticos y hexadecano. Se
analizaron por cromatografía de gases acoplado a un detector de espectrofotometría de
masas marca Perkin Elmer modelo Clarus 500. El hexadecano se cuantifico por
cromatografía de gases con detector de ionización de flama (CG-FID) utilizando nitrógeno
(1.5 mL/min.) como gas acarreador y una columna capilar (30m x 0.32mm x 25µm).
Cinéticas de degradación de hexadecano. Las cinéticas se llevaron a cabo en botellas
serológicas de 60 mL, utilizando 15 mL de sedimento y 35 mL de agua de mar real
esterilizada, para un volumen total de trabajo de 50 mL. A las cuales se les agrego
hexadecano a una concentración de 250 mg/L y tween 20 como surfactante (LizardiJiménez, 2011), posteriormente se sellaron con arillos de aluminio y septos de teflón
(PTFE). El experimento se llevó a cabo por duplicado con una duración de 28 días,
tomando muestra cada 7 días, cabe destacar que las botellas eran sacrificables. Los
controles abióticos se hicieron bajo las mismas condiciones antes mencionadas,
695
exceptuando que se sellaron y esterilizaron primero y posteriormente se les inyecto el
hexadecano y el tween 20. Se calculó la velocidad máxima de consumo utilizando el
modelo de Gompertz (Vigueras et al., 2008) y la eficiencia.
TABLA 8. Parámetros fisicoquímicos y microbiológicos de los puntos de muestreo.
Sitio
HCB
PAH
Carbón orgánico
DO
E°
Turbidity
pH
NMP/g
µg/g
%
mg/L
mV
NTU
RG
1700
0.63
0.49
4.44
125
112.9
8.2
RC
1800
0.98
2.12
4.13
63.5
14.4
9.1
RGR
935
0.05
1.09
5.17
48.1
29.9
9.1
RSS
1170
0.51
0.26
2.97
-114.3
70.6
9.0
RT
940
0.04
0.29
3.26
52.2
11.1
9.2
RP
1140
0.06
0.26
5.05
128.8
572
7.4
LM
635
0.02
0.47
5.88
174
25.6
8.3
LO
5340
0.13
0.62
4.81
60.0
31.2
8.2
LC
1660
0.12
0.43
3.42
63.5
105.4
7.47
PCM
605
0.27
0.53
ND
151
ND
ND
ND= no determinado, HCB= bacterias hidrocarbonoclastas, DO= oxígeno disuelto, PAH=hidrocarburos
policíclicos aromáticos.
Resultados
Las velocidades de consumo de los sedimentos y las eficiencias se muestran en la Tabla
2.
Como se puede observar en la Tabla 2, los sedimentos con una mayor qmax fueron el
RG (39.8 mgHXD/L·d) y LC (34.1 mgHXD/L·d) con una eficiencia de 89 y 86%
respectivamente en contraparte los sedimentos que tuvieron el menor consumo fue el RC
(5.0 mgHXD/L·d), RGR (8.6 mgHXD/L·d) y PCM (9.3 mgHXD/L·d) con una eficiencia del
47, 81 y 49% respectivamente. Se detectó una velocidad de consumo y una eficiencia en
promedio de 18.7±12 mgHXD/L y 73±16% respectivamente
En el caso del RC el bajo consumo y eficiencia se pudieron deber a la inhibición por
PAH o algún otro compuestos, debido que este sitio fue el que presento la mayor
concentración de estos compuestos así como de carbono orgánico, a pesar de que la
concentración de BHC fue de las más altas (Tabla 1), en el caso del sedimento RGR, el
bajo consumo se pudo deber a un efecto combinado entre la baja concentración de
bacterias hidrocarbonoclastas y el pH (Tabla 1), se ha reportado que la actividad en el
consumo del petróleo disminuye a partir de pH 9 (Das y Chandran, 2011). Esto se
evidencio en los sedimentos donde se encontró un pH de 9, los cuales tuvieron las
velocidades más bajas en el consumo de HXD.
696
TABLA 9.Velocidades de consumo y eficiencias de los diferentes sedimentos.
Sitio
q
max
Eficiencia
mg/L·d
%
RG
39.8
89
RC
5.0
RGR
Sitio
q
max
Eficiencia
mg/L·d
%
RP
18.9
80
47
LM
28.7
80
8.6
81
LO
12.5
85
RSS
16.1
75
LC
34.1
86
RT
13.6
56
PCM
9.3
49
De todos los sedimentos evaluados se encontró que el sedimento RSS presentaba
un potencial reductor negativo (-114.3 mV; Tabla 1), lo cual indica que en ese sedimento
la actividad anaerobia es la que predomina (Monroy, 1998), por tal motivo debido a la
metodología en que se llevaron a cabo los experimentos (aerobio) pudo influir en la tasa
de consumo.
En el caso de la eficiencia se pudo observar que a pesar de que algunos sedimentos
presentaron bajas velocidades de consumo (Tabla 2) lograron alcanzar eficiencias por
arriba del 75%, como fue el caso de los sedimentos RGR, RSS, LO, con una eficiencia del
81, 75 y 85% respectivamente, eficiencia muy similar a la de los sedimentos RG y LC que
tuvieron las velocidades más altas (Tabla 2), como se puede ver en la Figura 2, lo cual
pudiera indicar que las poblaciones microbianas no están adaptados a hidrocarburos, por
tal motivo tienen una lenta respuesta al consumo de HXD.
McGenty et al, (2012) reportan que dependiendo de la comunidad microbiana
presente en los ambientes marinos, se va a favorecer o no el consumo de hidrocarburos,
en el sentido que donde hay una población microbiana adaptada, se ven favorecidas las
interacciones de las bacterias, favoreciendo tanto la eficiencia como el consumo
697
100
40
Eficiencia
qmax 80
35
Eficincia (%)
60
25
40
20
15
qmax (mgHXD/Ld)
30
20
10
0
5
RG
LC
RGR
RSS
LO
Sitio
FIGURA 6. Velocidad de consumo y eficiencia de algunos sedimentos. En la figura se
puede apreciar como los sedimentos RGR, RSS y LO, a pesar de tener una qmax baja
alcanzaron eficiencias altas como los sedimentos con la mayor velocidad de consumo
(RG y LC).
Conclusiones
Todos los sedimentos evaluados degradaron el hexadecano, con una velocidad de
consumo promedio de 18.7±12 (mg/L·d) y una eficiencia del 73±16%.Ninguno de los
sedimentos degrado totalmente el hexadecano.
Se observó que uno de los parámetros que afecto el consume de hexadecano fue el
pH. Inhibiendo el consume de este.
A pesar de presentar una qmax baja algunos sedimentos lograron alcanzar eficiencias
de remoción altas.
Agradecimientos
Al M. en C. Arturo Centeno, por la cooperación y realización del mapa de estaciones.
698
Bibliografia
Al-Wasafi, R.S. y Hamed, S.R. (2014). Bacterial Biodegradation of Crude Oil Using Local Isolates.
International Journal of Bacteriology. Vol. 14. Article ID 863272. 8 paginas.
Das, N. y Chandran, P. (2011). Microbial Degradation of Petroleum Hydrocarbon Contaminants: An
Overview. Biotechnology Research international. Article ID: 941810.
Doyle, E., Muckian, L., Hckey, A.M., Clipson, N. (2008). Capitulo 2: Microbial PAH degradation. Advances in
Applied Microbiology vol. 6. pp 27-66.
García-Cruz N., U. y Aguirre-Macedo, M., L. (2014), en Golfo de México. Contaminación e Impacto
ambiental (En Prensa).
Lizardi-Jiménez, M. (2011). Contribución al studio de la hidrodinámica y transferencía simultánea de masa.
Tesis de doctorado. Universidad Autónoma Metropolitana, México
Martínez-Lopez B. y Parés-Sierra, A. (1998). Circulación del Golfo de México inducida por mareas, viento y
la corriente de Yucatán. Ciencias Marinas. 24(1): 65-93.
McGenety, T., J., Folwell, B., D., McKew, B., A., y Sanni, G., O. (2012). Marine crude-oil biodegradation: a
central role for interspecies interactions. Aquatic Biosystems. 8:10.
Monroy, O.A. (1998). Modelamiento y control de un sistema de digestión anaerobia en dos etapas. Tesis
doctoral. UNAM. México. pp: 6-9.
Seymour, R.J. y Geyer, R.A. (1992). Fates and effects of oil spills. Annual Review Energy Environment. 17:
261-283.
Vigueras, G., Shirai, K., Martins, D., Teixeira-Franco, T., Fleuri, L. y Revah, S. (2008). Toluene gas phase
biofiltration by Paecilomyces lilacinus and identification of a hydrophobin protein from produced
thereof. Appl. Microbiol. Biotechnol. 80(1):147-154.
Yanggou, Z., Min, C., Jie, B., Xinwei, L., Zulfiqar, F., Qianli, W. (2014). Response of Microbial Community to
Petroleum Stress and Phosphate Dosage in Sediments of Jiaozhou Bay, China. Journal Ocean
university of China. 13(2): 249-256.
699
Section 7.
Wastewater Treatment
700
Page
Chapter 7.1. Biodegradation of olive, corn and soy oil by free and immobilized
Yarrowia lipolytica YB-387
Henry Flores-Santiago; Aurora Martínez-Trujillo; Mayola García-Rivero
703
Chapter 7.2. Studies on bioremediation of phenolic groups in industrial oil fractions
degraded by ligninolytic basidiomicetes
Borja Rodríguez de Francisco; María C. Terrón; Aldo E. González-Becerra
707
Chapter 7.3. Adsorción de As (III) y As(V) en nanoferritas de manganeso
Elías E. Hernández±Beteta; Oscar F. Mijangos±Ricardez;
Jaime López±Luna; Sergio Martínez±Vargas
714
Chapter 7.4. Eliminacion de amonio y cresoles por un lodo nitrificante en un reactor SBR
Juan A. Salas-Cortés; Flor M. Cuervo-López; Anne-Claire Texier.
724
Chapter 7.5. Eliminacion de hormonas persistentes en aguas residuales
Ana F. Ruiz Esparza Ortiz; Adrián Rodríguez García.
733
Chapter 7.6. Estructura de instrumentos tecnológicos en el tratamiento de aguas
residuales en México
Lilian E. Domínguez-Montero; Blanca Jiménez-Cisneros;
Rosa O. Cañizares-Villanueva; Sergio Caffarel-Méndez;
Eugenio Frixione-Garduño; Miguel A. Pérez-Angón; Héctor M. Poggi-Varaldo
743
Chapter 7.7. Remoción de azul índigo por la biomasa seca de cuatro cepas
de Trichoderma sp.
Erika Sánchez-Torres; Oscar García-Barradas;
Remedios Mendoza-López; Rosalba Argumedo-Delira
753
Chapter 7.8. Decoloración del colorante azul directo 2 mediante peroxidasas
Gabriela E. Moeller-Chávez; María Y. Vargas- Espinoza
762
Chapter 7.9. Increased biodegradability of the fungicide imazalil after photo-fenton
treatment
Roberto Candal; Federico Ariganello; Gustavo Curutchet; Elsa López-Loveira
771
Chapter 7.10. Biotechnology process for the valorization of the liquid residual
effluents of the citric industry
Elizabeth León-Becerril; César A. Zárate-Cruz; José J. Díaz-Torres;
Alberto López-López
780
Chapter 7.11. Influence of organic loading rate on the hydrogen production from
agro-industrial residues as substrate using AFBR reactors
Samantha Christine Santos; Paula R. Ferreira Rosa; Edson Luiz Silva
791
Chapter 7.12. Study of the effect of sterilization and inoculation of lignocellulolytic
fungi for treatment of the azo dye black terasyl sr in peat packed biofilters for
organic matter and color removal in the textile industry
M. P. Mejía -López; G. E. Moeller ±Chávez
804
701
Page
Chapter 7.13. Effect of the cosolvent in the degradation of naphthalene by
catalytic ozonation
Claudia M. Aguilar-Melo; Jorge I. Chairez-Oria; Julia L. Rodríguez-Santillán;
Tatiana Poznyak
810
Chapter 7.14. Aplication of an electrooxidation process using PbO2 anodes
for the atrazine removal
Juan Napoles-Armenta; Celestino García-Gómez; Pablo Gortáres-Moroyoqui
820
Chapter 7.15. Congo red removal using a combined system: anaerobic treatment
coupled to electrochemical systems
Itzel Romero-Soto; Celestino García-Gómez; Luis Álvarez-Valencia;
Edna Meza-Escalante; Pablo Gortáres-Moroyoqui
828
Chapter 7.16. Optimization of phenol removal using TI/PBO2 anode with response
surface methodology
Celestino García-Gómez; Juan Nápoles-Armenta; Pablo Gortáres-Moroyoqui
838
Chapter 7.17. Remotion of color and cod by electro-oxidation for textile effluent
Vianey Ariadna Burboa-Charis; Petia Mijaylova-Nacheva;
José J. Treviño-Resendez; Celestino García-Gómez;
Itzel C. Romero-Soto; Pablo Gortáres-Moroyoqui
849
Chapter 7.18. Remoción de metales pesados en solución acuosa empleando
la biomasa de la persea americana
Lizbeth L. Díaz-Muñoz; Adrián Bonilla-Petriciolet
858
Chapter 7.19. Cadmium removal capacity from water and its effect on
Schoenoplectus pungens
Carmen Méndez-Hurtado, Edith Flores-Tavizón, Juan P. Flores-Marguez,
Sergio Saúl Solís, Laura Yáñez-Espinosa
864
Chapter 7.20. Swine wastewater treated by an anaerobic secuential system with
pulsed recirculation
Germán Giácoman-Vallejos; Ileana González-Canché;
María C. Ponce-.Caballero
871
Chapter 7.21. Contribution to the assessment of the health status of the coast of
sidi ifni: biomonitoring of the impact of wastewater
Mohamad Abbassi, Abderrazak Kaaya, Ali Banaoui, Asma El Khou, Maryem Nadir,
Latifa Lefrere
880
Chapter 7.22. Remoción de metales pesados utilizando sistemas inmovilizados
de Spirulina máxima
A.R. Domínguez-Bocanegra
892
Chapter 7.23. Marcadores fenólicos de estrés en plantas in vitro de
Vallisneria americana con enriquecimiento y desbalance ambiental n/c
Patricia Cruz-Cerino, Violeta Ruiz-Carrera; Miguel A.Salcedo-Meza;
Alberto J. Sánchez; Carlos A. Ledesma Escobar; Victor J. Robles-Olvera
901
702
CHAPTER 7.1. BIODEGRADATION OF OLIVE, CORN AND SOY OIL BY FREE AND
IMMOBILIZED Yarrowia lipolytica YB-387
Henry Flores-Santiago (1); Ma. Aurora Martínez-Trujillo (1); Mayola García-Rivero* (1)
(1) Laboratorio de Catálisis enzimática, Tecnológico de Estudios Superiores de Ecatepec, Ecatepec de
Morelos, Edo. de México, México, CP 55210
ABSTRACT
In order to know the capability of Yarrowia lipolytica to degrade vegetal oils some
experiments were carried out with corn, olive and soy oil in a free culture and immobilized
culture with different concentrations of oil, those concentrations were 27, 54, 108 and 216
gL-1. The biomass grew in the free cells was higher than in the immobilized cells, when
olive oil was used added in the free cells culture was produced 100 gL-1 of biomass with
olive oil in the order hand immobilized cells just could grow 33 gL-1. Oil degradation were
better in low concentrations using free cells in olive oil, but immobilized cells were able to
degrade 74% of any of the three oils. It could see that using free cells can get more
biomass although immobilized cells are better using the three oils with lower biomass
production.
Key words: biodegradation, Yarrowia lipolytica, yeast, oils
Introduction
The contamination from vegetable oil industry has many environmental problems because
their manufacturing process requires 21 liters of water per liter of oil, more over these
industries throw away the residual water containing important concentrations of oils. These
compounds have a high impact on the environment because just 1 liter of oil contaminates
1,000 liters of water, so the disposal of this wastewater into environment causes serious
problems because the oil has a detrimental effect on microbial activity (Sasarman et al.,
2007). The presence of oils in the water streams stops the exchange of oxygen demand
(Sanchez, 1995), and by serving as a substrate for algae causes an overgrowth of these
that gives what results the death of wildlife in rivers and lakes and when the contaminated
water flows into the sea causes the death of reefs (Sánchez, 1995).
There are several physical and chemical methods for treating water contaminated
with oils, but these have one disadvantage that is the operating costs because of this they
have been sought alternative low-cost and environmentally friendly. In this sense a
biological treatment is the best alternative due to it is less expensive and friendly with the
environment and if we use immobilized cells we can reuse this cells (Papanikolaou et al
2001) and solve some of the typical problems that occur when free cells cultures were
used.
Because of this is important improve biodegradation of vegetal oil, for example by
using immobilized cells, and evaluate their capacity to degrade some vegetal oils. In this
study, we evaluated the ability of Y. lipolytica YB-387 immobilized on polyurethane foam to
degrade different vegetal oils concentrations, in addition the capacity of immobilized cells
was compared with free cells.
703
Inoculum preparation of Y. lipolytica YB-387 was performed in 300 mL of LPG medium
composed (%): yeast extract, 1; meat peptone, 2; glucose, 2; the medium pH was adjusted
to 7 (Lan et al., 2008). The culture flask was incubated in a shaker at 200 rpm for 48 hours
(Lan et al., 2008). For the immobilization of Y. Lipolytica previous mycelium suspension
was incubated, in the medium described, with polyurethane foam cubes of 1 cm 3 to 200
rpm and 30 ° C for 48 hours.
Trials oil removal. Removal tests were carried olive oil in a 125 ml Erlenmeyer flasks
containing 75 mL medium LPG, 10 cubes of polyurethane foam with immobilized cells and
different oil concentrations (27, 54, 108 and 216 gL -1). The flasks were incubated at 200
rpm and 30 ° C for 3 days. Samples were taken every 12 hours to determine the dry
weight biomass both in the foam and in the liquid phase, and the residual oil
(Papanikolaou et al., 2008).
Separation of the polyurethane foam. The contents of the flasks were removed through
a metal mesh separating the liquid phase of the foam. The foam retained in the metal
mesh was pressed in order to remove the excess of medium, with a pressure of 0.5
kg/cm2 for 10 min, subsequently was washed with hexane to drag the largest oil content
and the foam was left in foil to dry in the oven for 24 hours at 60 °C.
Treatment of samples. The liquid phase was removed from the flasks and it was mixed
with hexane into a separating funnel, letting the sample stand for 10 minutes. The organic
phase was separated and was filtered through filter paper using a pump, the biomass
retained on the paper was dried in an oven at 60 °C for 24 hours. (Papanikolaou et al
2008).
Quantification of residual oil. Olive oil was quantified by gravimetric, previously
separated the inorganic phase was deposited in aluminum containers to pass to the oven
at 60 º C for 24 hours.
The results in the figure 1 shows the growth of the biomass in the free culture, it can see
that culture that contained olive oil had the most biomass production, with 108 gL-1 and
216 gL-1 of oil it was produced 100 gL-1 of biomass. In the corn oil the Y. lipolytica just
could produce 39 gL-1 of biomass in the oil concentration of 108 gL-1. In the other hand the
Figure 2 shows the immobilized biomass, in this case one more time the olive oil have the
better biomass production reaching in the medium 30 gL-1of the biomass, that was higher
than those obtained with the corn oil and soy oil. These results demonstrated that
effectively the immobilized cells cultures can reduce biomass production that is a typical
problem in free cells culture.
704
FIGURE 1. Formation of free biomass
using olive oil (Ŷ), corn (Ŷ), and soy
(Ŷ) at a concentration of 27, 54, 108
and 216 gL-1
FIGURE 2. Formation of free
biomass using olive oil (Ŷ), corn (Ŷ),
and soy (Ŷ) at a concentration of 27,
54, 108 and 216 gL-1
The results in the figure 3 shows the degradation of the three oils in the free cell culture, it
can see that just oil olive oil can be degrade over 80% in the concentrations of 27, 54 and
108gL-1. The only one which was degrade as olive oil was the soy oil but just with 108 gL-1
of oil, the worst oil degraded by Y. lipolytica was corn oil with 55% in concentration of 108
gL-1. On the contrary, the figure 4 shows the degradation by immobilized cell culture, in
this case the degradation is more stable than free culture owing to the average
degradation that is 74%, 50%, 55%, and 35% with the concentrations of 27, 54, 108 and
216 gL-1, this can indicate that immobilized cell culture is better than free cells culture.
FIGURE 3. Degradation of vegetable
oils in a free cell culture. Olive oil (Ŷ),
cornoil (Ŷ), and soy oil (Ŷ) in
concentration of 27, 54, 108 and 216
gL-1.
FIGURE 4. Degradation of vegetable
oils in immobilized cell culture. Olive
oil (Ŷ), corn oil (Ŷ), and soy oil (Ŷ) in
concentrations of 27, 54, 108 and 216
gL-1.
705
Conclusion
The free cell cultures showed a lower degradation and higher biomass production in
comparison with immobilized cell culture, so the better option is to use immobilized cells
cultures moreover the previous results suggest that immobilized Y. lipolytica in
polyurethane foam might be applicable to remove oils from a wastewater.
References
/DQ: *H * \ :DQ - ³%LRGHJUDGDWLRQ RI RLO ZDVWHZDWHU E\ IUHH DQG LPPRELOL]HG Yarrowia lipolytica
:´Journal of Environmental Sciences, 237±242, 2009.
3DSDQLNRODRX60DULD36W\OLDQRV)0LFKDHO.\*HRUJH$³&LWULFDFLGSURGXFWLRQE\Yarrowia lipolytica
cultivated on olive-mill wastewater-EDVHGPHGLD´Bioresource Technology, 737-744, 2008.
Papanikolaou, S., Muniglia I., Chevalot1 G. y Marc I., ³Yarrowia lipolytica as a potential producer of citric acid
IURPUDZJO\FHURO´Journal of Applied Microbiology,419±2428, 2001.
6iQFKH]/(³&RQWUROGHODFRQWDPLQDFLyQGHODVDJXDV´(Q$VSHFWRV*HROyJLFRVGH3URWHFFLyQ$PELHQWDO
Unesco volume 1, 1995.
6DVDUPDQ('LFXWD&-XUFRDQH6/XSHVFX,*URSRVLOD'\7FDFHQFR/³,QIOXHQFHRIVRPHQXWULWLRQDO
factors on lipase production by <DUURZLD OLSRO\WLFD´ 5RPDQLDQ %LRWHFKQRORJLFDO /HWWHUV, 322-336,
2007.
706
CHAPTER 7.2. STUDIES ON BIOREMEDIATION OF PHENOLIC GROUPS IN
INDUSTRIAL OIL FRACTIONS DEGRADED BY LIGNINOLYTIC BASIDIOMICETES
Borja Rodríguez de Francisco (1); María C. Terrón (2); Aldo E. González Becerra (1)*
(1) CBMSO, Microbiology and Virology Department, Madrid, Spain.
(2) Centro Nacional de Microbiología, Instituto de Salud Carlos III Ctra. Majadahonda a Pozuelo Km 2 E28220 Majadahonda Madrid, Spain.
ABSTRACT
The ability of 34 strains of ligninolytic fungi to degrade a fraction of oil was tested. Their
enzymatic activities (enzymes Laccase, Manganese-peroxidase, Versatile-peroxidase and
Lignin-peroxidase) were quantified and those strains showing greater bioremediation
ability, both in solid and liquid medium were selected. These chosen strains were grown in
Kirk medium with three inductors (Poly-R 478, ABTS and Remazol blue) in order to detect
the presence of the indicated enzymes. Later, the cytotoxicity of different concentrations of
oil (1%, 2% and 4% v/v, equivalent to concentrations of 10.000, 20.000 and 40.000 ppm,
respectively) was tested on Kirk solid medium. These high concentrations are
unprecedented in these types of studies. Six strains were able to grow at every oil
concentration tested: Ganoderma resinaceum, Pycnoporus sanguineus, Trametes I-62,
Laccaria amethystina, Fomes fomentarius and Daedalea quercina. The production of
degrading enzymes was tested in Kirk liquid medium with 2% v/v of oil. Culture media with
/ without glucose and with / without oil were used as controls. Ganoderma resinaceum
showed Laccase activity in the media with oil; Laccaria amethystina showed Laccase and
MnP activity in all media assayed but at low concentrations. The strain Trametes I-62
showed activity for all the assayed enzymes, except Lignin-peroxidase, at high
concentrations in all tested media arising as the best candidate with the most
bioremediation potential.
Key words: biodegradation, enzymes, ligninolytic fungi, petroleum.
Introduction
Due to the exponential growth of the world population, (currently estimated at about 7.200
billion which is expected to reach 9.600 billion in 2050, according to the UN) it is expected
that the demand for oil as feedstock will rise considerably, and thus, the probability for
environmental disasters to increase. Likewise, during the last three decades, white-rot
fungi have been emerging in the area of biotechnology and bioremediation due to its high
potential to degrade recalcitrant phenolic compounds such as petroleum (Pointing, 2001).
707
Saprophytic fungi are responsible for mineralizing the decaying organic matter and,
for that, they use a number of enzymes that are secreted outside the cells. These
enzymes have very low substrate specificity, a feature that can be used in biotechnology.
The enzymes studied were: Laccase, Lignin peroxidase (LiP), manganese peroxidase
(MnP) and versatile peroxidase (VP). Laccase is a phenol oxidase that catalyzes the oneelectron oxidation of diphenols and aromatic amines by removing an electron and a proton
generating phenoxy and amino radicals which disrupt the molecular structures of the target
compounds (Thurston, 1994; Shah & Nerud, 2002). The LiP enzyme is a peroxidase that
oxidizes aromatic rings via redox reactions with hydrogen peroxide (Tien & Kirk, 1984).
MnP enzyme has a Mn2+ cation in the catalytic center that is oxidized to Mn3+ and acts as
a highly reactive and diffusible agent capable of oxidizing non phenolic and mainly
phenolic substrates (Wariishi et al., 1988; Wariishi et al, 1990). The VP has the same
activities as MnP enzyme but does not require a Mn 2+ cation for it. It uses H2O2 as an
electron donor (Mester & Field, 1998).
Petroleum is a homogeneous mixture consisting mainly of several hydrocarbon
compounds such as paraffins, asphaltenes, resins and aromatic compounds. Among these
petroleum compounds stand out those with aromatic rings that can be substrates for
enzymes of ligninolytic fungi for their structural resemblance to the wood lignin monomers.
They are asphaltenes and aromatic compounds, with pure aromatic rings or alkyl
substituents type.
In this paper the ability of 34 different ligninolytic basidiomycetous fungi to degrade a
fraction of petroleum compounds in solid and liquid media was tested, and Laccase, LiP,
MnP and VP enzymes activities were quantified under different culture conditions.
The petroleum used for our study was provided by the Spanish petroleum refining company REPSOL and corresponds to a light oil fraction. On this subject, 34 ligninolytic
fungi strains were tested. They were kept in a cold chamber (4ºC) and inoculated on Petri
dishes containing malt agar and Benomyl 3ppm to ensure the growth of basidiomycetes
and prevent external contaminations.
Confirming the presence of ligninolytic enzymes: Pieces of agar medium (MEA) of 0.5
x 0.5 cm containing mycelia of the corresponding strain, were inoculated in Petri dishes
containing 10 mL of fresh agar medium. Every 48 h the radial growth (in mm) and
coloration/discoloration of the medium by the influence of different inductors were
observed. Three inducers were used: Poly R 478, indicating the presence of Manganese
Peroxidase enzyme (MnP) when the medium is decolorized; Remazol blue which
discoloration indicates the presence of the enzyme Laccase (Levin et al., 2004); and
ABTS, colorless, but in the presence of Laccase enzyme changed to a more or less
intense green color depending on the enzyme concentration. Veratryl alcohol was added
to these media as inductor of these enzymes because it acts as a substrate.
Cytotoxicity test: Petri dishes were prepared with 10 mL of culture medium and
concentrations of 1%, 2% and 4% v/v of petroleum in these three media: i) agar and
petroleum; ii) agar, petroleum and Kirk medium and iii) agar, petroleum and modified Kirk
708
medium without glucose. Likewise, in order to check the cytotoxicity produced by control
phenolic-compounds, the strains were grown in media with Kraft-lignin or
lignosulphonates, wastes from the manufacture of paper bleaching with high
concentrations of phenols in solution.
Liquid culture preparation: Standardized Inocula were prepared from fungi grown in
Petri dishes with Malta agar medium. The agar was divided into small pieces that were
poured into a 250 mL Erlenmeyer flask containing 50 mL of Kirk medium. Later, 100 mL of
growth medium were poured in 500mL Erlenmeyer flasks with 2% v/v of the petroleum
fraction and 10% v/v of inoculum. The cultures were kept in a rotary shaker at 100 rpm at
25 °C for 240 hours. Thus, each strain grew in four different media: Two control-media,
one with standard Kirk medium and the other with the modified Kirk medium without
glucose; and two petroleum-media (with 2% v/v petroleum concentration): one with
standard Kirk medium and the other with kirk medium modified without glucose. Enzyme
assays were performed by extracting, every 48 hours, 2 mL of the culture broth of the
Erlenmeyer. The activity of the Laccase, MnP and VP was measured according to Santoyo
et al.(2008 (cf. Wariishi et al. 1992)) at a wavelength of 468 nm using plastic cuvettes and
2,6-dimethoxyphenol (DMP) as the substrate; LiP enzyme activity was measured in a
quartz cuvettes at a wavelength of 310 nm with 100 µL of 0.5 M sodium tartrate, 100 µL of
10 mM veratryl alcohol, 50 µL of 4 mM H2O2 and 300 µL of sample (Tien and Kirk, 1984).
For measuring the enzymes activities a Shimadzu (UV-1603) spectrophotometer was
used.
Results on the degradation of petroleum in solid medium: Only 15 of the 34 strains
showed high levels of enzyme production in media with ABTS, Poly-R 478 and Remazol
Blue. Thereafter, only 6 species of our CBMSO Collection showed to successfully
developed in the cytotoxicity test with petroleum: Ganoderma resinaceum (BR4),
Pycnoporus sanguineus (B42), Trametes I-62 (B24), Laccaria amethystina (5111), Fomes
fomentarius (BR5) y Daedalea quercina (Exp 31). These 6 strains were able to grow at
concentrations of 4% v/v, although at these concentrations the growth was much slower
and not all strains developed correctly. These concentrations are equivalent to 10.000,
20.000 and 40.000 ppm. Only one reference has been found using high concentrations of
oil degradation, but the maximum concentration assayed was 15.000 ppm (Hadibarata &
Tachibana, 2009). In short, the results suggest that these 6 strains selected are very good
candidates as phenol-degraders and with a promising biotechnological potential. Similarly,
the growth in Kraft lignin and lignosulfonate indicates that these species are able to grow
in media containing high concentrations of phenols, target groups in petroleum
bioremediation processes.
Figure 1 shows the average of the radial mycelial growth on agar containing petroleum.
The lowest growth in millimeters was obtained in Agar without Kirk medium and the
highest growth rate was observed in Kirk Agar medium with glucose, indicating the need to
supplement the medium to enhance the growth of microorganisms.
709
FIGURE 1. Mycelial growth on agar containing petroleum.
Results on the degradation of petroleum in liquid medium and enzymatic
quantification: Enzymatic activities of the 6 selected strains were analyzed. Of these,
Fomes fomentarius did not growth under the conditions assayed and the fungus Daedalea
quercina grew significantly in controls but the enzymatic activity were not detected.
The species Pycnoporus sanguineus (B42) presented laccase activity only in glucose
control. The MnP activity was measured late; over 144 h in both controls, highlighting the
control without glucose, suggesting that this species secretes this enzyme in adverse
conditions because of survival (see Figure 2.). It also produced VP activity but only in
controls and at very low concentrations. This species has been thoroughly studied and
their Laccase activity properties have a great biotechnological potential (Vikineswary et al,
1997; Lu et al, 2007.). Our results indicate that it is not a species with interest for
bioremediation of petroleum due to the absence of growth of the strains in the presence of
this raw material at least at the conditions assayed.
FIGURE 2. Activity of MnP (U/L) for 240 h in Picnoporus sanguineus (B42)
710
The fungus Ganoderma resinaceum (BR4) produced a notable Laccase activity with levels
up to 0.2 U/L in the petroleum medium with glucose. The species Laccaria amethystina
(5111) presents Laccase and MnP activity in the range of 0.18 U/L, as shown in figures 3
and 4. This genus contains only mycorrhizal species so that the expression of these
amounts of ligninolytic enzymes contrasts with the idea of plant-fungus symbiosis. This
phenomenon has already been observed, but it is still an area to investigate (Baldrian
2009).
FIGURE 3. Laccase activity (U/L) for 240h in Laccaria amethystine (5111).
FIGURE 4. MnP activity (U / L) during 240h in Laccaria amethystine (5111).
Finally, the strain of Trametes I-62 (B42) showed the highest concentrations of enzymes
laccase, MnP and VP in the media with petroleum. Especially noteworthy is the Laccase
activity concentrations up to 0.6 U/L in the petroleum media supplemented with glucose as
shown in Figure 5 and 6. This strain has been studied before because of its Laccase
activity and its relationship with the degradation of lignin (Mansur et al., 1997). No
bibliographic citations have been found regarding with its use with petroleum.
711
FIGURE 5. Laccase activity (U/L) during 240h in Trametes I-62 (B24).
FIGURE 6. MnP activity (U/L) during 240h in Trametes I-62 (B24).
Conclusion
Growth of certain strains on solid medium with ABTS, Poly-R 478 and Remazol Blue
inductors permit to select fungal species with the ability to degrade xenobiotic compounds
structurally similar to these molecules. Furthermore, these strains can grow at high
concentrations of kraft lignin and lignosulphonate providing a biotechnological alternative
in the area of the paper industry and paper bleaching process. On the other hand, the
survival of these strains at concentrations of petroleum as higher as 4% v/v makes this to
be an experiment without precedent and opens the door of petroleum biodegradation in
areas with high pollution risk. Finally, of the 34 basidiomycetous strains tested, Trametes I62 (B24) was chosen as the best candidate for petroleum bioremediation because it
produced Laccase, MnP and VP activities in all media with petroleum assayed and at the
highest concentrations.
712
References
Baldrian, P. (2009). Ectomycorrhizal fungi and their enzymes in soils: is there enough evidence for their role
as facultative soil saprotrophs? Oecologia, 161, 657±660.
Hadibarata,T., Tachibana, S. (2009). Microbial degradation of crude oil by fungi pre-grown on wood meal. En
Obayashi, Y., Isobe, T., Subramanian, A., Suzuki, S., y Tanabe, S. (Eds.) Interdisciplinary Studies on
Environmental Chemistry- Environmental Research in Asia. Terrapur, 317-322.
Levin, L., Papinutti, L. y Forchiassin, F. (2004). Evaluation of Argentinean white rot fungi for their ability to
produce lignin-modifying enzymes and decolorize industrial dyes. Bioresource Technology, 94, 169±
176.
Lu, L., Zhao, M., Zhang, B., Yu, S., Bian, X. J., Wang, W. y. Wang, Y. (2007). Purifi- cation and
characterization of laccase from Pycnoporus sanguineus and decolorization of an anthraquinone dye
by the enzyme. Applied Microbiology and Biotechnology, 74, 1232-1239.
Mansur, M., T. Suárez, J. B. Fernández-Larrea, M. A. Brizuela, and A. E. González. (1997). Identification of
a laccase gene family in the new lignin degrading basiodiomycete CECT 20197. Appl. Environ.
Microbiol., 63, 2637±2646.
Mester, T., Field, J.A. (1998). Characterization of a novel manganese peroxidase-lignin peroxidase hybrid
isozyme produced by Bjerkandera species strain BOS55 in the absence of manganese. J. Biol.
Chem., 273, 15412-15417.
Pointing, S. B. (2001). Feasibility of bioremediation by white-rot fungi. Appl. Microbiol. Biotechnol., 57, 20±
33.
Santoyo, F., A. E. González, M. C. Terrón, et al. 2008. Quantitative linkage mapping of lignin-degrading
enzymatic activities in P. ostreatus. Enzyme. Microb Technol 43:137-143.
Shah, V., Nerud, F. (2002) Lignin degrading system of white-rot fungi and its exploitation for dye
decolorization. Can. J. Microbiol., 48, 857-870.
Tien, M., Kirk, T.K. (1984). Lignin-degrading enzyme from Phanerochaete chrysosporium: purification,
characterization, and catalytic properties of a unique H 2O2-requiring oxygenase. Proc. Natl. Acad. Sci.,
81, 2280±2284.
Thurston, C.F. (1994) The structure and function of fungal laccases. Microbiology, 140, 19±26.
Vikineswary, S., N. Abdullah, M. Renuvathani, M. Sekaran, A. Pandey y Jones E. B. G. (2006). Productivity
of laccase in solid substrate fermentation of selected agro- residues by Pycnoporus sanguineus.
Bioresource Technology, 97, 171-177.
Wariishi, H., Akileswaran, L., y Gold, M. H. (1988). Manganese peroxidase from the basidiomycete
Phanerochaete chrysosporium: spectral characterization of the oxidized states and the catalytic cycle.
Biochemistry, 27:5365-5370.
Wariishi, H. y Gold, M. H. (1990). Lignin peroxidase compound III. Mechanism of formation and
decomposition. J. Biol. Chem., 265, 2070-2077.
Wariishi, H., Valli, K., Gold, M. H. (1992). Manganese (II) oxidation by manganese peroxidase from the
basidiomycete Phanerochaete chrysosporium. J. Bio. Chemistry, 267, 23688-23695.
713
CHAPTER 7.3. ADSORCIÓN DE As(III) Y As(V) EN NANOFERRITAS DE
MANGANESO
Elías E. Hernández±Beteta (1); Oscar F. Mijangos±Ricardez (1);
Jaime López±Luna* (1);Sergio Martínez±Vargas (2)
(1) Instituto de Estudios Ambientales, Universidad de la Sierra Juárez, Oaxaca., México.
(2) Facultad de Ingeniería, Universidad Autónoma del Carmen, Ciudad del Carmen, Campeche, México.
RESUMEN
Las nanoferritas de manganeso se sintetizaron por el método de co±precipitación
empleando una solución mixta de Fe (II), Mn (II) y Fe (III). Se obtuvieron partículas con
características bien definidas, tamaño promedio de 2.5 nm, densidad 5.22 g mL í1, área
superficial 81.09 m2 gí1 y punto de carga cero (pzc) 7.6. Los experimentos de adsorción
se realizaron con As(III) y As(V) a pH 2 y 8, observándose diferencias significativas
(p<0.05, prueba LSD). Los resultados de la adsorción de ambas especies del metaloide
demostraron que son dependientes del pH, siendo más evidente para As(V) ya que la
adsorción fue significativamente menor a pH 8. El porcentaje de remoción disminuyó con
el incremento de las concentraciones iniciales, 99%±28% para As(III) y 92%±27 para
As(V). El Modelo Cinético de Pseudo±Segundo±Orden (PSO) mostró valores de la
velocidad inicial h (g mgí1 miní1) mayores para As(V) que para As(III), lo cual demuestra
que la nanoferrita adsorbe más rápido As(V) que As(III), aunque la adsorción en el
equilibrio qe (mg g-1) fue mayor para As(III), en las condiciones de pH establecidas. La
adsorción (qe) para la máxima concentración inicial (40 mg Lí1) y pH 2 fue 49.53 y 27.18
mg gí1, para As(III) y As(V), respectivamente. En tanto que a pH 8 se obtuvieron valores
de qe de 33.96 y 18.46 mg gí1, lo cual indica que la adsorción no fue dependiente de la
velocidad inicial (h). La sustitución de Fe(II) por Mn(II) incrementa significativamente los
grupos hidroxilo de la nanoferrita, favoreciendo la capacidad adsortiva de la partícula y
ofreciendo mayor resistencia a la oxidación, en comparación con nanopartículas de
magnetita.
Palabras clave: adsorción, arsénico, modelo PSO, nanoferrita.
Introducción
La contaminación del recurso hídrico por la incorporación progresiva de metales pesados
y metaloides ha incrementado la preocupación mundial en las últimas décadas. A
diferencia de los contaminantes orgánicos, los cuales en su mayoría son biodegradables,
los metales pesados no se degradan y son altamente tóxicos para la vida acuática,
incluso en concentraciones relativamente bajas (Nadeem et al., 2006).
714
El arsénico (As), metaloide incluido en el Listado de Contaminantes Prioritarios de la
US±EPA, se encuentra de manera natural en rocas, suelo, agua, aire, plantas y animales.
Es más común encontrar a este elemento en formas inorgánicas. El arsenato As(V) es
más dominante en ambientes aeróbicos, mientras que el arsenito As(III) lo es en
ambientes anaeróbicos (Shipley et al., 2009). De estas dos especies químicas, el arsenito
es más tóxico, soluble y móvil, del cual se han detectado concentraciones de 100 a 2000
g L-1 en aguas subterráneas de muchos países (Zhang et al., 2010).
La adsorción es uno de los métodos fisicoquímicos más empleados para eliminar estos
contaminantes, método atractivo debido a sus ventajas de eficiencia, bajo costo y
operación simple (Crini, 2005). Recientemente, diversos estudios se han enfocado en
nano±adsorbentes como una técnica prometedora para el saneamiento de agua (Andrew,
2008). Las nanopartículas de óxidos de hierro, especialmente magnetita, hematita,
maghemita y ferritas han mostrado ser buenos adsorbentes de metales pesados y
metaloides (Karami, 2013).
Numerosos publicaciones demuestran que los óxidos de hierro tienen alta afinidad
para la adsorción de arsenito y arsenato (Yean et al., 2005; Hoang et al., 2009; Shipley et
al., 2009; Zhang et al., 2010; Pérez±Caballero, 2012). Las nanoferritas que proceden de la
sustitución parcial o total de Fe(II) por Mn(II) en la estructura de la magnetita, ofrecen
mayor resistencia a la oxidación sin afectar las propiedades adsorbentes del material. De
esta manera, el objetivo de este trabajo fue evaluar la capacidad adsortiva de nanoferritas
de manganeso para la eliminación de As(III) y As(V) a escala de laboratorio, mediante el
modelo cinético de Pseudo±Segundo±Orden (PSO).
Síntesis y caracterización de la partícula. La nanoferrita de manganeso se sintetizó por
la técnica de co±precipitación (Massart, 1981) empleando una solución mixta de Fe (II),
Mn (II) y Fe (III), compuesta de 0.5 mmol de sulfato ferroso (FeSO4+2O), 1.5 mmol de
sulfato manganoso (MnSO4+2O) y 2 mmol de nitrato férrico (Fe(NO3)3+2O ), a la que se
agregaron gota a gota 50 mL de NaOH 1.5 M. La solución se mantuvo en agitación
constante a 80°C 1.5 h, obteniéndose un ferrofluido completamente negro. La partícula se
lavó tres veces con agua desionizada para eliminar el exceso de hidróxido de sodio y se
secó a 30°C.Se determinó el tamaño promedio, el punto de carga cero (PZC), la densidad
y el área superficial de las nanopartículas como se describe brevemente.
En el último lavado de la partícula se tomaron 50 µL del ferrofluido y se
suspendieron en 10 mL de acetona grado reactivo. Sobre esta suspensión se hizo incidir
un láser obteniendo el tamaño promedio en un equipo Microtrac Nanotrac 252. Para la
determinación del PZC se prepararon soluciones de distintos valores de pH en NaCl
0.01M, ajustando con ácido clorhídrico e hidróxido de sodio. Se agregó 1 g L í1 de
nanoferrita y después de medir el pH inicial las soluciones se agitaron 48 h a 250 rpm,
después de lo cual se determinó el pH final. El PZC se obtuvo al punto de intersección
entre la curva del pHinicial contra el pHfinal y la recta del pHinicial = pHfinal, en base a la técnica
descrita por Bajpai (2008). La densidad se determinó por el método del picnómetro.
El área superficial (m2 gí1) se calculó con el inverso del producto de la densidad y el
diámetro promedio de la partícula que se obtuvo con el equipo Nanotrac.
715
Cinéticas de adsorción. Las cinéticas de adsorción se realizaron con concentraciones
iniciales de 2, 5, 10, 20 y 40 mg Lí1 de As(III) y As(V), a pH 2 y 8, con una proporción de
partícula de 0.4 g Lí1. Los tiempos de contacto fueron 5, 10, 15, 30, 60, 120, 180 y 240
min con agitación oscilatoria de 250 rpm. La concentración inicial y final del metaloide se
determinó por ICP±OES. La adsorción se calculó con la fórmula:
୲ ൌ ሺ୭ െ ୲ ሻ ‫כ‬

Donde qt (mg g±1) es la adsorción del metaloide por la ferrita de manganeso en el tiempo,
Co y Ct (mg L±1) son las concentraciones iniciales y las concentraciones de arsénico en el
tiempo, respectivamente, V es el volumen de la solución (mL), y m es la masa de la
nanoferrita de manganeso (Boparai et al., 2011).
Modelo de Pseudo±Segundo±Orden. El modelo de PSO se aplicó para analizar la
cinética de quimiosorción en soluciones líquidas.

ͳ
ͳ
ൌ
൅ ‫ݐ‬
ଶ
୲ ݇ଶ ‫ݍ‬௘ ‫ݍ‬௘
Donde k2 (g mgí1 miní1) es la constante de velocidad para la adsorción de PSO y K2qe2 o
h (g mgí1 miní1) es la velocidad inicial de la adsorción.
Este modelo asume que un ion de arsénico puede ser adsorbido en dos sitios de sorción
en la superficie de la nanoferrita (Boparai et al., 2011).
Síntesis y caracterización de la partícula. Se obtuvieron partículas con características
bien definidas de magneticidad y el color negro característico de la magnetita. Las
nanopartículas presentaron un tamaño promedio de 2.35 nm (Figura 1). El PZC de la
partícula se localizó a pH 7.6 (Figura 2), en concordancia con los datos reportados por
Zhang et al. (2010). Esto significa que en estas condiciones de pH la partícula presenta el
mismo número de cargas positivas que negativas, con lo que se ve desfavorecida la
adsorción de especies iónicas. La ferrita presentó una densidad 5.22 g mL±1 y un área
superficial de 81.09 m2 gí1 superior a la partícula de magnetita (10 nm tamaño promedio)
reportada por Phu et al. (2009).
716
FIGURA 1. Diagrama de dispersión de tamaño de partícula de la nanoferrita de
manganeso.
FIGURA 2. Determinación del Punto de Carga Cero (PZC) de la nanoferrita de
manganeso.
Cinéticas de adsorción. El porcentaje de remoción disminuyó con el aumento de las
concentraciones iniciales, 99%±28% para As(III) y 92%±27 para As(V). La adsorción de
ambas especies del metaloide mostró diferencias significativas (p<0.05, prueba LSD)
entre pH 2 y 8 en concordancia con lo reportado por Shipley et al. (2009).
La adsorción del metaloide en concentraciones iniciales relativamente bajas (2±10 mg
í1
L ) fue rápida en los primeros 50 min, alcanzando el equilibrio a los 240 min (Figura 3 y
4). Resultados contrarios a los reportados por Zhang et al. (2010), quienes obtuvieron la
mayor adsorción a los 120 minutos y condiciones de equilibrio a las 12 horas empleando
una ferrita de manganeso para la adsorción de las mismas especies del metaloide.
Conforme se fueron incrementando las concentraciones iniciales se alcanzaron más
717
lentamente las condiciones de equilibrio, que puede deberse a la menor velocidad del
gradiente de concentración. Esto fue más notorio para As(V) a pH 8 (Figura 4B), en donde
también influyó el PZC (pH 7.6) de la nanoferrifa, ya que a un pH por debajo de este punto
los grupos hidroxilo en la superficie se protonan formando OH 2+, lo que facilita el
intercambio de enlace con el anión arsenato. Con el aumento del pH se da una
desprotonación progresiva de los grupos hidroxilo, cargándose negativamente, lo cual es
desfavorable para la adsorción de As(V) porque la partícula lo repele, llegando incluso a
desorberlo a un pH mayor de 9 (Yean et al., 2005), lo cual a su vez explica el
comportamiento errático de la adsorción de esta especie (Figura 4b). Respecto al As(III),
en el rango de pH 2±8 su carga es neutra, lo cual no modificaría sustancialmente su
adsorción en la nanoferrita. No obstante, la adsorción fue mayor a pH 2 que a pH 8
(Figuras 3 y 4 a).
FIGURA 3. Cinética de adsorción de As(III) y As(V) en ferritas de manganeso a pH 2.
A y B muestran la remoción del metaloide (mg L -1); a y b muestran la adsorción en la
nanopartícula (mg g-1).
718
FIGURA 4. Cinética de adsorción de As(III) y As(V) en ferritas de manganeso a pH 8.
A y B muestran la remoción del metaloide (mg L -1); a y b muestran la adsorción en la
nanopartícula (mg g-1).
En las figuras 5 y 6 se observa la grafica de t/tq en función del tiempo. En las tablas 1 y 2
se muestran los parámetros de velocidad obtenidos con el modelo de PSO para ambas
especies del metoloide. Los resultados indican que la cinética de la adsorción de As(III) y
As(V) en nanoferritas de manganeso se ajustó bien al modelo cinético de PSO (R 2>0.99).
719
FIGURA 5. Ajuste al Modelo de Pseudo±Segundo±Orden de la adsorción de
Arsénico en nanoferritas de manganeso a pH 2.
TABLA 1. Parámetros cinéticos de la adsorción de As(III) y As(V) en nanoferritas de
manganeso a pH 2
As(III)
(mg L±1)
2
5
10
20
40
As(V)
(mg L±1)
2
5
10
20
40
k2
(g mgí1
miní1)
0.1159
0.0134
0.0096
0.0057
0.0060
k2
(g mgí1
miní1)
0.2880
0.0181
0.0090
0.0069
0.0062
h
(mg gí1
miní1)
3.407
2.309
3.838
5.695
14.786
h
(mg gí1
miní1)
6.143
2.658
2.748
3.341
4.938
R2
qe,exp
(mg gí1)
qe,cal
(mg gí1)
X²
0.99999
0.99987
0.99635
0.99806
0.99961
5.39
12.85
20.54
32.05
49.53
5.422
13.151
19.968
31.546
49.751
0.00019
0.00689
0.01639
0.00805
0.00098
R2
qe,exp
(mg gí1)
qe,cal
(mg gí1)
X²
0.99998
0.99976
0.99967
0.99873
0.99926
4.59
12.04
17.22
22.03
27.98
4.62
12.13
17.49
21.96
28.18
0.00019
0.00067
0.00417
0.00022
0.00142
720
FIGURA 6. Ajuste al Modelo de Pseudo±Segundo±Orden de la adsorción de As(III) y
As(V) en nanoferritas de manganeso a pH 8.
TABLA 2. Parámetros cinéticos de la adsorción de As(III) y As(V) en nanoferritas de
manganeso a pH 8
As(III)
(mg L±1)
2
5
10
20
40
As(V)
(mg L±1)
2
5
10
20
40
k2
(g mgí1
miní1)
0.0255
0.0137
0.0078
0.0067
0.0034
k2
(g mgí1
miní1)
0.0524
0.0069
0.0119
0.0070
0.0134
h
(mg gí1
miní1)
0.417
0.862
1.180
3.185
3.938
h
(mg gí1
miní1)
0.510
0.197
0.641
0.990
4.553
R2
qe,exp
(mg gí1)
qe,cal
(mg gí1)
X²
0.99420
0.99897
0.99653
0.99980
0.99927
3.93
7.80
12.03
21.36
32.98
4.04
7.94
12.28
21.80
33.97
0.00300
0.00247
0.00509
0.00888
0.02885
R2
qe,exp
(mg gí1)
qe,cal
(mg gí1)
X²
0.99825
0.97173
0.99724
0.98905
0.99777
3.12
5.34
7.34
11.94
18.45
2.95
5.24
7.28
12.45
18.73
0.00980
0.00191
0.00049
0.02089
0.00419
El modelo de PSO mostró que a pH 2 la velocidad inicial de adsorción h de As(III) se
incrementó (3.407±14.786 mg gí1 miní1) mientras que la constante de velocidad k2
disminuyó (0.1159±0.0060 g mgí1 miní1) con el incremento de las concentraciones
iniciales. De la misma forma, a partir de la concentración inicial de 5 mg L -1, h (2.658±
4.938 mg gí1 miní1) se incrementó y k2 (0.018±0.0062 g mgí1 miní1) disminuyó para As(V).
721
A pH 8 se observaron tendencias similares pero con velocidades menores. Así, para
As(III) se obtuvo h (0.417±3.938 mg gí1 miní1) y k2 (0.0255±0.0034 g mgí1 miní1).
Mientras que para As(V) se obtuvo h (0.197±4.553 mg gí1 miní1) y k2 (0.0134±0.0069 g
mgí1 miní1). La velocidad inicial de la adsorción (h) es empleada como un indicador para
evaluar el inicio del proceso de adsorción. Es claro que h aumenta en ambas especies
siendo mayores los valores para As(V) en concentraciones bajas (2 y 5 mg L -1), lo cual
indica que la nanoferrita de manganeso adsorbe más rápido As(V) que As(III) con estas
concentraciones. Sin embargo como se mencionó anteriormente, la adsorción del
metaloide fue mayor para As(III), es decir, la adsorción de ambas especies no dependió
de la velocidad inicial de adsorción (h), lo cual concuerda con lo reportado por Tang et al.
(2011).
La adsorción en equilibrio (qe) para la máxima concentración inicial (40 mg L-1) y pH
2 fue de 49.53 y 27.18 mg gí1, para As(III) y As(V), respectivamente. En tanto que a pH 8
se obtuvieron valores de qe de 33.96 y 18.46 mg g-1, lo cual indica que la adsorción del
metaloide depende en gran medida del pH, incluso empleando partículas distintas (Yean
et al., 2005; Hoang et al., 2009; Shipley et al., 2009; Cabe recalcar que la sustitución de
Fe(II) por Mn(II) incrementa significativamente los grupos hidroxilo de la nanoferrita
(Zhang et al., 2010), obteniendo una mayor adsorción en comparación con la magnetita.
Esta sustitución no solo incrementa los sitios de sorción en la partícula, también le otorga
mayor resistencia a la oxidación (Phu et al., 2009). Nuestros resultados también indican
que la partícula adsorbe en mayor medida al As(III), el cual es más tóxico y móvil en el
ambiente que el As(V).
Conclusiones
Las nanoferritas de manganeso presentaron mayor adsorción de As(III) y As(V) a pH2, sin
embargo puede ser un adsorbente ideal en condiciones ambientales de pH. Al adsorber
en mayor medida As(III), estas partículas pueden ser consideradas como una alternativa
prometedora de bajo costo para sanear aguas contaminadas con esta especie, la cual ha
ocasionado diversos problemas de salud debido a las altas concentraciones detectadas
en todo el mundo. El modelo de PSO describió con gran precisión el comportamiento
adsortivo en las condiciones experimentales, lo que permitió establecer la adsorción en el
equilibrio.
Agradecimientos
Este trabajo fue financiado por los proyectos SEP±CONACYT CB±2010±157232, SEP±
CONACYT CB±2012±01 181592 y UNSIJ±CA±3.
722
Referencias
%RSDUDL + -RVHSK 0 2¶&DUUROO ' 0 .LQHWLFV DQG WKHUPRG\QDPLFV RI FDGPLXP LRQ UHPRYDO E\
adsorption onto nano zerovalent iron particles, Journal of Hazardous Materials 186:458±465.
Crini G., (2005), Recent developments in polysaccharide±based materials used as adsorbents in wastewater
treatment, Progress in Polymer Science. 30:38±70.
Karami H. (2013), Heavy metal removal from water by magnetite nanorods, Chemical Engineering Journal
219:209±216.
Massart R. (1981). Preparation of aqueous magnetic liquids in alkaline and acidic media. IEEE Transactions
Magnetics. 17:1247±1248.
Nadeem M, Mahmood A, Shahid S A, Shah S S, Khalid A M, McKay G, 2006. Sorption of lead from aqueous
solution by chemically modified carbon adsorbents. Journal of Hazardous Materials B, 138:604±613.
Pérez-Caballero I., (2012). Síntesis de nanopartículas de magnetita para la remoción de arsénico en agua.
Tesis de Licenciatura. Universidad de la Sierra Juárez. Oaxaca.
Phu N.D., Phong P.C., Chau N., Luong N.H., Hoang L.H., Hai N.H. (2009). Arsenic removal from water by
magnetic Fe1±xCoxFe2O4 and Fe1±yNiyFe2O4 nanoparticles. Journal of Experimental Nanoscience Vol.
4, No. 3:253±258.
Shipley J., Yean S., Amy T. y Mason B., (2009), Adsorption of arsenic to magnetite nanoparticles: effect of
particle concentration, ph, ionic strength, and temperature, Environmental Toxicology and Chemistry,
Vol. 28, No. 3, pp. 509±515.
Tang W., Qi Li, Shian G., Shang J., (2011), Arsenic (III,V) removal from aqueous solution by ultrafine Į±
Fe2O3 nanoparticles synthesized from solvent thermal method, Journal of Hazardous Materials
192:131±138.
Yean S., Cong L., Yavuz C., Mayo J., Yu W., Kan A., Colvin V., Tomson M., (2005), Effect of magnetite
particle size on adsorption and desorption of arsenite and arsenate, Journal of Materials Research,
20(12):3255±3264.
Zhang S., Niu H., Cai Y., Zhao X., Y. Shi, (2010), Arsenite and arsenate adsorption on coprecipitated bimetal
oxide magnetic nanomaterials: MnFe2O4 and CoFe2O4, Chemical Engineering Journal 158:599±607.
723
CHAPTER 7.4. ELIMINACION DE AMONIO Y CRESOLES POR UN LODO
NITRIFICANTE EN UN REACTOR SBR
Juan Antonio Salas-Cortés (1), Flor M. Cuervo-López (1),
Anne-Claire Texier*(1)
(1) Universidad Autónoma Metropolitana-Iztapalapa, Departamento de Biotecnología, División CBS, Av. San
Rafael Atlixco 186, Col. Vicentina, C.P. 09340, México D.F., México.
RESUMEN
Los procesos biológicos acoplados de nitrificación y desnitrificación permiten la
eliminación de amonio de las aguas residuales en forma de N 2. Recientemente, el uso de
consorcios nitrificantes en reactores de lotes secuenciados (SBR) ha sido propuesto como
una nueva alternativa para la oxidación simultánea de amonio y compuestos fenólicos de
efluentes industriales. Sin embargo, todavía no se tiene información de tipo fisiológica y
cinética de la oxidación simultánea del amonio y de los isómeros del cresol por consorcios
nitrificantes en reactores biológicos. El objetivo del trabajo fue evaluar la capacidad
metabólica de un consorcio nitrificante para oxidar simultáneamente amonio, o-cresol (ocr), m-cresol (m-cr) y p-cresol (p-cr) en un reactor SBR. Se operó un SBR de 1l con ciclos
de 12 h (llenado: 0.07h; reacción aerobia: 11.26h; sedimentación: 0.25h; drenado:0.42h).
Se operó el SBR sin cresoles (fase I: ciclos 1-172) y posteriormente se alimentó con
cresoles: fase II: m-cr (10, 20 y 30 mg C/l) (ciclos 173-240); fase III: m-cr (30 mg C/l) y o-cr
(10, 20 y 30 mg C/l) (ciclos 241-354); fase IV: m-cr (30 mg C/l) con o-cr (30 mg C/l) y p-cr
(30 mg C/l) (ciclos 355-413). En la fase I, se estabilizó un proceso respiratorio nitrificante
con altos valores de eliminación de amonio (ENH4 = 100%). El producto mayoritario fue el
nitrato (rendimiento YNO3 = 0.94 r 0.10 g N-NO3-/g N-NH4+ consumido) y no hubo
acumulación de nitrito. Al adicionar los cresoles (fase II a IV), se observó que la actividad
nitrificante no fue alterada ya que todo el amonio fue oxidado a nitrato (E NH4 = 100% y
YNO3 = 0.91 r 0.06), indicando que los tres isómeros en mezcla no afectaron la
nitrificación. En todos los casos, los diferentes cresoles fueron totalmente consumidos con
eficiencias de 100%. Se observó un proceso de adaptación metabólica del lodo para
oxidar los cresoles ya que las velocidades específicas (q) aumentaron a través de los
ciclos de operación (qm-cr aumentó de 0.0003 a 0.450 mg C/mg proteína.h en 54 ciclos; qocr pasó de 0.024 a 0.061 en 98 ciclos). El o-cresol fue el más recalcitrante de los tres
isómeros. El consorcio presentó la capacidad metabólica para oxidar simultánea y
totalmente el amonio y los tres cresoles. El uso del SBR permitió una adaptación
metabólica del lodo en consumir los cresoles. Este tipo de reactor puede constituir una
buena alternativa para tratar efluentes industriales en un solo reactor.
Palabras claves: amonio, cinética, cresoles, nitrificación, reactor SBR.
724
ABSTRACT
Ammonium removal from wastewaters can be performed by two coupled biological
processes: nitrification and denitrification. Recently, the use of nitrifying consortia in
sequencing batch reactors (SBR) has been proposed as a novel alternative for the
simultaneous removal of ammonium and phenolic compounds from industrial effluents.
However, there is still not physiological and kinetic information on the simultaneous
oxidation of ammonium and isomers of cresol by nitrifying consortia in biological reactors.
The aim of this study was to evaluate the metabolic capacity of a nitrifying consortium to
simultaneously oxidize ammonium, o-cresol (o-cr), m-cresol (m-cr) and p-cresol (p-cr) in a
SBR. A 1l SBR was operated with cycles of 12 h (fill: 0.07h; aerobic reaction: 11.26h;
settle: 0.25h; draw: 0.42h). First, the reactor was operated without cresols addition (phase
I: cycles 1-172) and then, it was fed with cresols: phase II: m-cr (10, 20, and 30 mg C/l)
(cycles 173-240); phase III: m-cr (30 mg C/l) and o-cr (10, 20, and 30 mg C/l) (cycles 241354); phase IV: m-cr (30 mg C/l) with o-cr (30 mg C/l) and p-cr (30 mg C/l) (cycles 355413). During phase I, the nitrifying respiratory process was stabilized with high values of
ammonium consumption (ENH4 = 100%). Nitrate was the main product (yield YNO3 = 0.94 r
0.10 g NO3--N/g NH4+-N consumed) and there was no accumulation of nitrite. When
cresols were added to the reactor (phase II to IV), it was observed that the nitrifying activity
of the sludge was not altered because all the ammonium was oxidized to nitrate (ENH4 =
100% and YNO3 = 0.91 r 0.06), showing that the three isomers in mixture not altered the
nitrification process. In all cases, the different cresols were totally consumed with
efficiencies of 100%. A metabolic adaptation of the consortium to oxidize the cresols was
observed as the specific rates of consumption (q) increased throughout the operation
cycles (qm-cr increased from 0.0003 to 0.450 mg C/mg protein.h in 54 cycles; qo-cr
increased from 0.024 to 0.061 in 98 cycles). o-Cresol was the most recalcitrant of the three
isomers. The sludge showed the metabolic ability to oxidize simultaneously and totally the
ammonium and the three cresols. The use of SBR allowed a metabolic adaptation of the
consortium to consume cresols along the operation cycles. This type of reactor can be a
good alternative for treating industrial effluents in a unique reactor.
Key words: ammonium, cresols, kinetic data, nitrification, sequencing batch reactor
Introducción
Los efluentes provenientes de la industria química, petrolera, de manufactura del acero,
de producción de resinas, entre otras, pueden contener altas concentraciones de amonio
y compuestos fenólicos (Olmos y col., 2004). Los compuestos fenólicos (fenol, cresoles,
clorofenoles, etc.) pueden causar efectos nocivos sobre la salud del ser humano y causar
efectos de toxicidad y bioacumulación en organismos vegetales y animales (Davi y Gnudi,
1999). La contaminación del agua por amonio también representa un riesgo para los
ecosistemas y la salud del hombre generando problemas de toxicidad, bioacumulación y
eutroficación de cuerpos de agua (Cervantes, 2009).
El sistema nitrificante-desnitrificante constituye una forma efectiva, sustentable y
económicamente factible de eliminar biológicamente el nitrógeno de aguas residuales en
725
forma de nitrógeno molecular. La nitrificación se define como un proceso respiratorio
aerobio y litoautotrófico mediante el cual se lleva a cabo la oxidación del amonio hasta
nitrato vía nitrito, por bacterias amonio oxidantes y nitrito oxidantes (Bock y col., 1991). La
desnitrificación es un proceso metabólico anóxico donde el nitrato es utilizado como
aceptor terminal de electrones y es reducido a nitrógeno molecular (Cuervo-López y col.,
2009).
En la literatura, se puede encontrar estudios sobre el efecto inhibitorio de
compuestos fenólicos en la nitrificación y en conjunto sobre la capacidad metabólica de
los lodos nitrificantes en oxidar dichos compuestos: fenol (Yamagishi y col., 2001; Amor y
col., 2005; Silva y col., 2011), p-cresol (Texier y Gómez, 2007; Beristain-Cardoso y col.,
2011), 2-clorofenol (Martínez-Hernández y col., 2011; Silva y col., 2011; Pérez-Alfaro y
col., 2013), p-hidroxibenzaldehido (Téllez-Pérez y col., 2013). Recientemente, se ha
propuesto que el uso de consorcios microbianos con actividad nitrificante previamente
estabilizada en reactores de lotes secuenciados representa una alternativa de tratamiento
efectiva para llevar a cabo la oxidación simultánea de amonio a nitrato y compuestos
fenólicos a dióxido de carbono (Téllez-Pérez y col., 2013; Zepeda y col., 2013; Silva y col.,
2014). Al realizarse todo el proceso en un solo reactor, el proceso es más económico.
Cabe señalar que se sabe muy poco todavía de la oxidación del p-cresol y m-cresol por
lodos nitrificantes en reactores biológicos y aún menos información sobre la eliminación
del o-cresol. Además, hacen falta estudios sobre la oxidación de los compuestos fenólicos
cuando están presentes en mezclas. Por consiguiente, el objetivo de este estudio fue evaluar la capacidad de un
consorcio nitrificante para oxidar el amonio, m-cresol, o-cresol y p-cresol en un reactor de
lotes secuenciados.
Inóculo y medio de cultivo. El inóculo se obtuvo de un reactor nitrificante en continuo de
tanque agitado en estado estacionario alimentado con un medio de cultivo litoautotrófico
para la nitrificación. El reactor continuo presentó los siguientes valores promedio de
eficiencia de consumo de amonio y rendimiento de nitrato: 99 ± 5% y 0.95 ± 0.05 g N-NO3/g N-NH4+ consumido, respectivamente. De acuerdo con los resultados obtenidos, el
consorcio del reactor en continuo presentó una actividad fisiológica nitrificante estable con
altos valores de eficiencia de consumo de amonio y rendimiento de formación de nitrato y
se pudo utilizar como inóculo para el reactor de lotes secuenciados. El reactor SBR se
inoculó inicialmente con una concentración de 200 mg de proteína microbiana por litro.
La alimentación del reactor SBR se realizó mediante dos medios de cultivo utilizados
como fuentes de nitrógeno y carbono (Tabla 1). La concentración inicial en amonio en
cada ciclo de operación fue de 72.4 ± 16.5 mg N/l.
726
TABLA 1. Composición química del medio de cultivo en el reactor SBR.
Fuente de nitrógeno (g/l)
(NH4)2SO4
NH4Cl
KH2PO4
MgSO4
NaCl
0.585
0.474
0.459
0.353
0.541
Fuente de carbono (g/l)
NaHCO3
4.573
CaCl2
0.032
Reactor de lotes secuenciados. Se utilizó un reactor de volumen nominal de 1l
(Microferm Fermentor, New Brunswick Scientific, mod. MF-114). Todos los ensayos
fueron realizados a temperatura ambiente (25 r Û& /D DLUHDFLyQ HQ ORV FXOWLYRV VH
mantuvo constante y la agitación de 250 rpm. La concentración promedio de oxígeno
disuelto se mantuvo a 5.8 ± 0.2 mg/l. El valor promedio de pH se mantuvo en un intervalo
de 7-8 utilizando bicarbonato de sodio como amortiguador. Cada ciclo de operación de 12
h consistió de las siguientes fases: llenado (0.07 h), reacción biológica (11.26 h),
sedimentación (0.25 h) y drenado (0.42 h).
Primeramente, se efectuaron ensayos abióticos control en el reactor. Esto se realizó
con el fin de descartar posibles interferencias dentro del reactor como pueden ser
reacción química con los componentes del medio de cultivo, adsorción en la unidad
experimental y/o volatilización de los cresoles y garantizar la reproducibilidad de los
estudios.
Posteriormente, se inoculó el reactor y se operó bajo condiciones nitrificantes sin la
adición de cresol para obtener un proceso respiratorio nitrificante estable con altos valores
de eficiencia de eliminación de amonio y de rendimiento de formación de nitrato.
Finalmente, se procedió a la adición de los diferentes isómeros de acuerdo a la Tabla 2.
TABLA 2. Concentraciones iniciales de cresoles adicionados en el SBR.
Fase
I
Compuesto(s)
sin cresol
II
m-cresol
III
m-cresol (30 mg C/l) + ocresol
m-cresol + o-cresol
+ pcresol
IV
Concentración (mg C/l)
0
10
20
30
10
20
30
30 cada cresol
Ciclos
1-172
173-198
199-226
227-240
241-282
283-310
311-354
355-413
Variables de respuesta. Para evaluar el proceso nitrificante en el SBR sin y con la
adición de cresoles, se determinaron las siguientes variables de respuesta: la eficiencia
de consumo en amonio (ENH4, (g N-NH4+ consumido/g N-NH4+ inicial) x 100) y los
rendimientos de formación de nitrato (YNO3, g N-NO3- formado/g N-NH4+ consumido) y
biomasa (YBM, g N-biomasa formada/g N-NH4+ consumido).
Para evaluar la capacidad del lodo en oxidar los diferentes cresoles, se utilizaron las
siguientes variables de respuesta: eficiencia de consumo (Ecresol, (g cresol consumido/g
727
cresol inicial) x 100) y velocidades específicas de consumo (q, mg C-cresol/mg proteína
microbiana.h). Se determinaron las velocidades volumétricas (V) a partir de las cinéticas
realizadas a través de una regresión lineal y se calcularon las velocidades específicas (q =
V/X donde X es la concentración en biomasa).
Métodos analíticos. El amonio se midió con un electrodo selectivo de amonio. El nitrito,
el nitrato y los cresoles se cuantificaron por HPLC de acuerdo a la metodología descrita
por Silva y col. (2014). La proteína microbiana se midió por el método colorimétrico de
Lowry (Lowry y col., 1951). Estadísticamente los métodos analíticos presentaron un
coeficiente de variación en la pendiente menor que 6% y los valores de R 2 fueron
cercanos a 1, indicando una alta reproducibilidad y linealidad de los métodos.
Estabilización del proceso nitrificante sin adición de cresoles. En la fase I de
operación del reactor sin adición de cresoles (ciclos 1-172), se obtuvieron altos valores de
consumo de amonio (ENH4 = 100%). El producto mayoritario fue el nitrato (rendimiento
YNO3 = 0.94 r 0.10 g N-NO3-/g N-NH4+ consumido) y no hubo acumulación de nitrito. Estos
resultados muestran que se obtuvo la estabilización del proceso nitrificante dentro del
reactor. El rendimiento en biomasa fue solamente de 0.004 ± 0.001 g N-proteína/g N-NH4+
consumido, indicando un proceso respiratorio principalmente desasimilativo.
Nitrificación en el SBR alimentado con cresoles. El perfil nitrificante del SBR
alimentado con cresoles se presenta en la Figura 1. Se observó una completa oxidación
de amonio, la formación mayoritaria de nitrato sin acumulación de nitrito en el efluente,
obteniéndose una ENH4 de 100% y un YNO3 de 0.91 r 0.06 g N-NO3- formado/g N-NH4+
consumido. Estos resultados muestran que los tres isómeros del cresol en mezcla no
afectaron la nitrificación en el SBR.
Concentración (mg N/l)
90
80
70
60
50
40
30
20
10
0
Fase II
Fase III
Fase IV
165 185 205 225 245 265 285 305 325 345 365 385 405 425
Ciclos
FIGURA 1. Perfil nitrificante del SBR alimentado con cresoles. Ÿ $PRQLR HQ HO
influente; ( QLWUDWRHQHOHIOXHQWHżQLWULWRHQHOHIOXHQWH[DPRQLRHQHOHIOXHQWHFase
II: m-cr (10, 20 y 30 mg C/l) (ciclos 173-240); fase III: m-cr (30 mg C/l) y o-cr (10, 20 y 30
mg C/l) (ciclos 241-354); fase IV: m-cr (30 mg C/l) con o-cr (30 mg C/l) y p-cr (30 mg C/l)
(ciclos 355-413).
728
Oxidación de los cresoles por el lodo nitrificante. Primero, se realizaron ensayos
abióticos dentro del reactor SBR. Las concentraciones de o-, m- y p-cresol se mantuvieron
constantes a lo largo del tiempo, comprobando que no hubo pérdida de los cresoles bajo
las condiciones experimentales utilizadas en el sistema.
a
12
10
8
6
4
2
0
0
1
2
3
4
5
6
7
b
Concentración (mg-C/l)
25
20
15
10
5
0
0
1
2
3
4
5
6
7
c
35
30
25
20
15
10
5
0
0
1
2
3
4
Tiempo (h)
5
6
7
FIGURA 2. Cinéticas de oxidación de cresoles a diferentes ciclos de operación del
SBR. (a) (') Ciclo 173: m-FUHVROPJ&O\ŶFLFOR m-cresol (10 mg C/l), (b) Ciclo
199: m-FUHVRO PJ &O F &LFOR Ŷ m-FUHVRO PJ &O \ ż o-cresol (10 mg
C/l).
En el primer ciclo de adición de m-cresol al SBR (ciclo 173), no hubo consumo
significativo de m-cresol a lo largo de 6 h de cultivo (Figura 2a). Posteriormente, en el ciclo
729
185, todo el m-cresol (10 mg C/l) fue consumido en tan solo 1 h, indicando que el lodo
adquirió una mayor capacidad metabólica para oxidar al m-cresol. Al ciclo 199, el lodo fue
capaz de consumir 20 mg C/l de m-cresol en menos de 10 min (Figura 2b). La velocidad
específica de oxidación del m-cresol aumentó de 77 veces su valor entre el ciclo 173 y el
ciclo 185 y de 12 veces entre el ciclo 185 y el 199, alcanzando un valor máximo de 0.450
mg C/mg proteína microbiana.h al ciclo 227 (Tabla 3). Estos resultados muestran que el
lodo adquirió una mayor capacidad metabólica para oxidar al m-cresol a través de los
ciclos, la cual podría ser relacionada con un mayor crecimiento de las cepas
consumidoras de cresoles del consorcio y/o la inducción de enzimas involucradas en la
degradación de cresoles (Silva y col., 2014). Estudios previos han reportado que las
condiciones estables y repetitivas del sistema SBR pueden propiciar un incremento en las
velocidades específicas de consumo (Zhuang y col., 2005; Texier y Gómez, 2007).
Posteriormente en el ciclo 241, se añadió o-cresol a 10 mg C/l junto con el m-cresol a
30 mg C/l y se observó que los dos isómeros fueron completamente oxidados en 1 h
(Figura 2c). Es interesante observar que la velocidad de oxidación del m-cresol disminuyó
de 70% cuando se añadió o-cresol (10 mg C/l) por primera vez al reactor, mostrando el
efecto inhibitorio del o-cresol sobre la capacidad metabólica del lodo para consumir el mcresol (Tabla 3). Estos resultados están de acuerdo con lo reportado por Zepeda y col.
(2007) en su estudio con mezclas de compuestos BTX (benceno, tolueno y xileno). Estos
autores enfatizan en la importancia de considerar las posibles interacciones (sin
LQWHUDFFLyQ LQKLELFLyQ LQGXFFLyQ FRPHWDEROLVPR« TXH SXHGHQ RFXUULU HQ OD
biotransformación de mezclas de compuestos aromáticos en sistemas de nitrificación.
Del ciclo 241 al ciclo 257, la velocidad de consumo de o-cresol aumentó su valor de 7
veces, mostrando una mayor capacidad metabólica del lodo para oxidar también el ocresol (Tabla 3). Al aumentar la concentración inicial de o-cresol a 20 mg C/l (ciclo 283),
las velocidades de oxidación tanto del m-cresol como del o-cresol disminuyeron,
probablemente por el efecto inhibitorio del o-cresol y posteriormente aumentaron de
nuevo. Es destacable que a pesar de la adición de 30 mg C-o-cresol/l al ciclo 311, las
velocidades siguieron en aumento. Cabe hacer notar también que en la mayoría de los
ciclos, el o-cresol fue más lentamente consumido que el m-cresol, indicando que el ocresol fue el más recalcitrante de los dos isómeros. Estos resultados concuerdan con los
citados por Lee y col. (2011) en su estudio realizado con gránulos aerobios donde los
autores identifican al o-cresol como el compuesto más recalcitrante con la siguiente
secuencia de eliminación de los isómeros del cresol: m-cresol > p-cresol > o-cresol. Silva
y col. (2014) reportaron previamente una velocidad de consumo de p-cresol de 0.91 mg
C/mg proteína microbiana.h en un reactor SBR con un lodo nitrificante. Este valor es
mayor que las velocidades obtenidas en el presente estudio para el o- y el m-cresol. De
acuerdo a Ho y col. (2010), el p-cresol fue el isómero de los tres más rápidamente
consumido en lodos activados. Estos resultados muestran que los isómeros presentan
una recalcitrancia distinta en función de la posición del grupo metilo en la molécula pero
también de las condiciones de estudio.
Por último, se añadieron los tres isómeros del cresol al reactor nitrificante a una
concentración de 30 mg C/l cada uno (Tabla 3). En los ciclos 355 y 369, se obtuvo la
completa oxidación de los tres isómeros del cresol en menos de 2 h, obteniéndose
eficiencias (Em-cresol, Eo-cresol y Ep-cresol) del 100%. No se detectó por HPLC la presencia de
intermediarios de la oxidación de los cresoles en el efluente.
730
TABLA 3. Velocidades específicas de consumo de m-cresol, o-cresol y p-cresol en
el SBR.
Ciclo
Concentración (mg C/l)
173
185
199
213
227
241
257
283
297
311
339
355
369
m-cresol (10)
m-cresol (10)
m-cresol (20)
m-cresol (20)
m-cresol (30)
m-cresol (30) + o-cresol (10)
m-, o- y p-cresol (30 cada uno)
m-, o- y p-cresol (30 cada uno)
q (mg C/mg proteina microbiana.h)
m-cresol
o-cresol
p-cresol
0.0003±0.0001
--0.023±0.001
--0.277±0.011
--0.289±0.010
0.450±0.009
--0.137±0.003
0.024±0.001
-0.140±0.003
0.158±0.003
-0.026±0.003
0.019±0.002
-0.062±0.001
0.024±0.001
-0.189±0.004
0.096±0.002
-0.312±0.016
0.061±0.003
--0.094±0.012
0.193±0.025*
-0.073±0.007
0.139±0.014*
Notes: *Velocidad de consumo de (p-cr + m-cr) ya que los dos isómeros no se pudieron
separar por HPLC.
Conclusión
El lodo nitrificante presentó la capacidad metabólica para oxidar simultáneamente amonio,
m-cresol, o-cresol y p-cresol en un reactor SBR, obteniéndose eficiencias de consumo del
100% y rendimientos en nitrato cercanos a 1. De manera general, se observó que las
velocidades específicas de oxidación de los isómeros aumentaron con el número de ciclos
de operación, mostrando una adaptación metabólica del lodo en consumir los tres
cresoles. Estos resultados indican que el reactor de lotes secuenciados nitrificante puede
ser una buena opción para eliminar simultáneamente amonio y compuestos inhibitorios o
recalcitrantes del agua, tales como los derivados del fenol.
Agradecimientos
Este trabajo fue financiado por el proyecto SEP-Conacyt-CB-2011-01-165174.
731
Referencias
Amor, L.; Eiroa, M.; Kennes, C.; Veiga, M.C. (2005). Phenol biodegradation and its effect on the nitrification
process. Water Research. 39: 2915-2920.
Beristain-Cardoso, R.; Pérez-González, D.N.; González-Blanco, G.; Gómez, J. (2011). Simultaneous
oxidation of ammonium, p-cresol and sulfide using a nitrifying sludge in a multipurpose biorreactor: A
novel alternative. Bioresource Technology. 102: 3623-3625.
Bock, E.; Koops, H.P.; Harms, H.; Ahler, B. (1991). The biochemistry of nitrifying organisms. Variations in
autotrophic life. Editorial: Academic Press, San Diego. Pp. 171-200.
Cervantes, F.J. (2009). Anthropogenic sources of N pollutants and their impact on the environment and on
public health. In: Cervantes, F.J. (Ed.). Environmental Technologies to Treat Nitrogen Pollution. pp. 117, IWA Publishing, London.
Cuervo-López, F.; Martínez-Hernández, S.; Texier, A.-C.; Gómez, J. (2009). Principles of denitrifying
processes. In: Cervantes, F.J. (Ed.). Environmental Technologies to Treat Nitrogen Pollution. pp. 4154, IWA Publishing, London.
Davi, M.L.; Gnudi, F. (1999). Phenolic compounds in surface water. Water Research. 33: 3213-3219.
Ho, K.L.; Chen, Y.Y.; Lee, D.J. (2010). Functional consortia for cresol-degrading activated sludges: Toxicityto-extinction approach. Bioresource Technology. 101: 9000-9005.
Lee, D.J.; Ho, K.L; Chen, Y.Y. (2011). Degradation of cresols by phenol-acclimated aerobic granules.
Applied Microbiology and Biotechnology. 89:209±215.
Lowry, O.H.; Rosebrough, N.J.; Farr, A.L.; Randall, R.J. (1951). Protein measurement with the folin phenol
reagent. The Journal of Biological Chemistry. 193: 265-275.
Martínez-Hernández, S.; Texier, A.-C.; Cuervo-López, F.M.; Gómez, J. (2011). 2-Clorophenol consumption
and its effect on the nitrifying sludge. Journal of Hazardous Materials. 185: 1592-1595.
Olmos, A.; Olguin, P.; Fajardo, C.; Razo, E.; Monroy, O. (2004). Physicochemical characterization of spent
caustic from the OXIMER process and sour waters from Mexican oil refineries. Energy & Fuels. 18:
302-304.
Pérez-Alfaro, J.E.; Buitrón, G.; Gomez, J.; Texier, A.-C.; Cuervo-López, F.M. (2013). Kinetic and
Physiological Evaluation of Ammonium and Nitrite Oxidation Processes in Presence of 2Chlorophenol. Applied Biochemistry and Biotechnology. 169: 990-1000.
Silva, C.D.; Gómez, J.; Beristain-Cardoso, R. (2011). Simultaneous removal of 2-clorophenol, phenol, pcresol and p-hydroxybenzaldehyde under nitrifying conditions: Kinetic study. Bioresource Technology.
102: 6464-6468.
Silva, C.D.; Beristain-Montiel, L.; Cuervo-López F.M.; Texier, A.-C. (2014). p-Cresol mineralization and
bacterial population dynamics in a nitrifying sequential batch reactor. Journal of Environmental
Sciences. In press. DOI 10.1016/j.jes.2014.06.033.
Téllez-Pérez, S.K.; Silva, C.D.; Texier, A.-C. (2013). Oxidación simultánea de amonio y phidroxibenzaldehido en un reactor de lotes secuenciados. Revista Mexicana de Ingeniería Química.
12: 97-104.
Texier, A.-C.; Gómez, J. (2007). Simultaneous nitrification and p-cresol oxidation in a nitrifying sequencing
batch reactor. Water Research. 41: 315-322.
Yamagishi, T.; Leite, J.; Ueda, S.; Yamaguchi, F.; Suwa, Y. (2001). Simultaneous removal of phenol and
ammonia by an activated sludge process with cross-flow filtration. Water Research. 35: 3089±3096.
Zepeda, A.; Texier, A.-C.; Gómez, J. (2007). Batch nitrifying cultures in presence of mixtures of benzene,
toluene, and m-xylene. Environmental Technology. 28: 355-360.
Zepeda, A.; Ben-Youssef, C.; Rincón, S.; Cuervo-López F.M.; Gómez, J. (2013). Complete and
simultaneous removal of ammonium and m-cresol in a nitrifying sequencing batch reactor.
Biodegradation. 24: 377-385.
Zhuang, W.Q.; Tay, J.H. ; Yi, S.; Tay, S.T.L. (2005). Microbial adaptation to biodegradation of tert-butyl
alcohol in a sequencing batch reactor. Journal of Biotechnology. 118: 45-53.
732
CHAPTER 7.5. ELIMINACION DE HORMONAS PERSISTENTES EN AGUAS
RESIDUALES
Ana Fátima Ruiz Esparza Ortiz (1), Adrián Rodríguez García (1)
(1) Centro de Investigación y Desarrollo Tecnológico en Electroquímica, Parque Tecnológico Querétaro s/n,
Sanfandila Pedro Escobedo, Querétaro, C.P. 76703 México.
[email protected]; [email protected]
RESUMEN
El objetivo de este estudio es la investigación de la combinación de procesos biológicos
con procesos de oxidación avanzada para la eliminación de compuestos recalcitrantes
como hormonas en aguas residuales de granjas lecheras. Se inició con un proceso de
Digestión Anaerobia en un reactor UASB como pretratamiento obteniendo resultados de
remoción de DQO de hasta 85%, posteriormente se compararon procesos de oxidación
avanzada como Fenton con diferentes combinaciones de reactivos obteniendo los
mejores resultados de remoción de DQO de hasta 83% con las concentraciones más altas
tanto de FeSO4 como de H2O2 con una relación de 2.5mg/L: 25 mL/L respectivamente. El
proceso Fenton se comparó con el proceso Electrofenton tomando como parámetros
principales de este proceso, la densidad de corriente y la concentración de FeSO 4,
obteniendo también los mejores resultados con la mayor densidad de corriente y la mayor
concentración de FeSO4 con un porcentaje de remoción de DQO de 61%. Para la
GHWHUPLQDFLyQ GH OD UHPRFLyQ GH ODV KRUPRQDV VH WUDEDMy FRQ XQ HVWiQGDU GH ȕ
Estradiol siendo esta hormona una de las más comunes en Aguas Residuales. Se
observó su comportamiento mediante HPLC para determinar la posible degradación de
esta en cada una de las etapas del proceso, la combinación de estos procesos nos
permite lograr un tratamiento viable, eficaz y económico permitiendo obtener bajas
concentraciones de carga orgánica y altas remociones de hormonas en aguas residuales.
Palabras Clave: ȕ HVWUDGLRO DJXDV UHVLGXDOHV, hormonas, oxidación avanzada,
procesos Biológicos, UASB.
Introducción
En la actualidad una de las problemáticas más graves que se afronta en el mundo es la
falta de suministros de agua, por lo que las operaciones de alimentación concentrada de
animales (CAFO) es una estrategia de reciclado de agua factible y económico en los
campos agrícolas, brindándoles nutrientes y materia orgánica para el crecimiento vegetal,
aun así estas suelen contener altas concentraciones de hormonas consideradas como
disruptores endocrinos, con riesgos potenciales en ecosistemas.
Un disruptor endocrino ha sido recientemente descrito como "una sustancia química
exógena o de la mezcla que altera la estructura o función del sistema endocrino y causa
efectos adversos a nivel del organismo, su progenie, las poblaciones o subpoblaciones de
los organismos, con base en principios científicos principios, los datos, la ponderación de
ODV SUXHEDV \ ORV SULQFLSLRV GH SUHFDXFLyQ´ %LJVE\ 5REHUW \ FRO /RV HVWXGLRV LQ
vitro han demostrado que las concentraciones bajas de hormonas estrogénicas
733
esteroideas, incluso a niveles tan bajos como ng/L, pueden afectar negativamente a la
biología reproductiva de los recursos acuáticos (Jobling y col. 1998, Thrope y col. 2003).
La exposición a ciertos disruptores endocrinos ha contribuido a la aparición de efectos
adversos sobre poblaciones y especies silvestres. Estos efectos varían desde cambios
sutiles en la fisiología y comportamiento sexual de las especies hasta alteraciones
permanentes de la diferenciación sexual (Jason. W. y col. 2003). Los efectos en los
humanos para los que vincula la exposición de disruptores endocrinos incluyen lo
siguiente: reducción temporal en el conteo y calidad de esperma, incidencia en el
incremento de cancer testicular y prostata, incidencia en el incremento de cáncer de
mama en mujeres, efectos neurológicos.
Es por eso que el objetivo de este proyecto es la eliminación de estos compuestos
recalcitrantes por medio del tratamiento de aguas residuales combinado usando un
proceso de digestión anaerobia como pretratamiento para la remoción de materia
orgánica, sólidos, grasas y aceites y un proceso de oxidación como post tratamiento
para la degradación de disruptores endocrinos, obteniendo así un proceso viable, eficiente
y económico.
Proceso biológico UASB.Los procesos biológicos son métodos destructivos y
constituyen una serie de importantes procesos de tratamiento que tienen en común la
utilización de microorganismos (entre las que destacan las bacterias) para llevar a cabo la
eliminación de componentes indeseables del agua, aprovechando la actividad metabólica
de los mismos sobre esos componentes (Rodríguez y col. 2006) Se distinguen tres tipos
de sistemas: sistemas aerobios, sistemas anaerobios y sistemas anóxicos. Se trabajó con
un proceso anaerobio en el cual las muestras iniciales antes del pretratamiento biológico
fueron tomadas en una granja lechera, manteniéndolas en refrigeración a -4°C en
garrafas de 20 litros, posteriormente de estas garrafas la muestra fue tomada como
influente para ser pre-tratado en un reactor UASB a nivel laboratorio con un volumen de 7
L. El tratamiento en el reactor UASB se mantuvo con un tiempo de retención de 18 horas
y una carga orgánica de 2.5 kgDQO/L*d. Las muestras en el efluente del reactor UASB
fueron recolectadas para su posterior tratamiento con procesos de oxidación avanzada.
Procesos de Oxidación Avanzada. Los procedimientos avanzados de oxidación también
VRQ PpWRGRV GHVWUXFWLYRV TXH VH GHILQHQ FRPR ³DTXHOORV SURFHVRV GH R[LGDFLyQ¶¶ TXH
implican la generación del radical hidroxilo para la destrucción de contaminantes. Estos
radicales tienen un alto potencial redox (2,9 eV) y son capaces de destruir e incluso
mineralizar cualquier contaminante orgánico (Comninellis, 2008; Sirtori, 2009).
734
FIGURA 1. Reactor Anaerobio de Flujo Ascendente
Fenton. Dentro de los procesos de oxidación avanzada seleccionamos el proceso Fenton
químico ya que es el más común y con el que más se ha trabajado. Este sistema se basa
en la generación de los radicales hidroxilo (·OH) a partir de los iones de hierro que actúan
como catalizadores homogéneos, preferentemente a temperatura ambiente y pH ácidos
(Walling, 1975; Bautista, 2008).
En cuanto a las condiciones de operación óptimas el reactivo Fenton está
condicionado por el pH, la concentración de peróxido y por la concentración del
catalizador (Gogate, 2004; Bautista, 2008), entre otros factores. Entre las principales
ventajas del reactivo Fenton se encuentran su facilidad de operación, que la reacción
discurre a temperatura y presión ambiental y que no son necesarias fuentes adicionales
de energía.
El tratamiento Fenton se llevó a cabo en lotes usando vasos de precipitado de 250
mL tomando diferentes dosis de reactivos Fenton. Se aplicaron 2 concentraciones de 2.5
y 1 g/L de FeSO4 mientras que el H2O2 (solución acuosa al 30%, la densidad de 1.11 kg /
L) se aplicó a 10 y 25 mL / L con un pH de 3 a temperatura ambiente; se usó un diseño de
experimentos 23 con réplica, siendo un total de 8 combinaciones diferentes. Todos los
experimentos se llevaron a cabo a pH 3, el cual fue ajustado usando H2SO4 1 N o NaOH 1
N. Después de la adición de la concentración requerida de FeSO 4 y H2O2 a las muestras
de aguas residuales, el agua se somete a un mezclado a 100 rpm durante 3 horas,
tomando muestras a 30, 60, 90,150 y180 min para determinar la eficiencia del proceso
dependiendo del tiempo.
Después del proceso se subió el pH 8 para detener la reacción y se esperó 15 hrs.
para aclaramiento de la muestra. El sobrenadante fue filtrado y las muestras se analizaron
por medio de DQO y COT.
735
Electrofenton. Este proceso se seleccionó por las altas eficiencias de remoción en
compuestos recalcitrantes y por qué según la literatura este tipo de procesos son muy
efectivos en la remoción de disruptores endocrinos como hormonas (Naimi y Col 2012).
El electro-Fenton es un proceso que se incluye en el grupo de los métodos de
electro-oxidación indirecta con oxidantes fuertes. Esta técnica ha despertado un interés
importante por su compatibilidad medioambiental y su efectividad.
Debido a que el peróxido por si solo carece de un poder oxidante elevado, no se
suele usar en estos casos por sí solo, pero si mantenemos una generación constante y
añadimos un ión de ferroso (Fe2+), resulta de Ia combinación de ambos un agente de gran
poder oxidante, el radical hidroxilo (*OH) que es el principal agente de este proceso de
oxidación avanzada (Iniesta, 1999).
Las pruebas de Electrofenton se llevaron a cabo en una celda electroquímica tipo
filtro prensa con una capacidad de 3 L. Como cátodo y ánodo se usaron placas de
Diamante dopado con Boro y se trabajó con densidades de corriente de 7.81 mA/cm 2,
15.62 mA/cm2 y 31.25 mA/cm2 mientras que las concentraciones de FeSO4 fueron de 0.5
y 0.7 mM, haciendo todas las combinaciones posibles entre estos dos parámetros. El
tiempo del proceso fue de 3 hrs tomando muestras en 30, 60, 90, 120, 150 y 180 min,
para su posterior análisis de DQO y COT.
Antes de iniciar cada prueba fue necesario agregar oxígeno a nuestra muestra para
que al estar en contacto con corriente aplicada empezara la producción de H2O2 en el
medio, también fue necesario bajar el pH a 3 para que la eficiencia de la reacción fuera
mayor.
Fuente de poder
Cátodo
Oxigeno
Ánodo
Fuente de poder FIGURA 2. Funcionamiento de proceso Electro-Fenton.
UASB pre-tratamiento. El pretratamiento anaerobio en el reactor UASB degradó gran
parte de DQO y COT como se muestra en las Gráficas 1 y 2. La mayor degradación se
dio en el día 16 con un DQO en el influente de 1713 mg/L y después del tratamiento la
concentración de DQO en el influente fue de 288 mg/L por lo que la eficiencia de
degradación fue de 85%, teniendo un promedio de degradación de DQO en el proceso de
68%.
En COT la mayor degradación se registra en el día 11 con una concentración de carbón
orgánico total en el influente de 178.5 ppm y en el efluente una concentración de 70.7
736
ppm teniendo un porcentaje de eficiencia de remoción del 60% con un promedio total de
degradación de COT de 42%. La eficiencia de proceso biológico en gran medida
dependió de la composición de las aguas residuales.
A
B
GRAFICA 1. A) Monitoreo de influente y efluente de DQO en UASB B) Monitoreo de
influente y efluente de TOC en UASB.
Post-tratamiento Procesos de oxidación avanzada (POA´s)
Fenton. El proceso de Fenton se llevó a cabo en aguas del efluente del reactor UASB,
utilizando diferentes combinaciones de concentraciones de los reactivos Fenton (FeSO 4
g/L y H2O2 mL/L) tomando muestras en diferentes tiempos, estas combinaciones se
muestran en la Tabla 1.
TABLA 1. Concentraciones de FeSO4 y H2O2 usadas en el proceso Fenton
FeSO4 (g/L) H2O2 (mL/L)
1
2.5
2.5
1
10
25
10
25
El tratamiento Fenton reduce el color del agua residual. Los mejores resultados de
combinaciones fueron (2.5 mg de FeSO4/L + 25 mL de H2O2/L) a los 180 min, causado
completa decoloración del efluente, con una degradación de DQO del 85% y de COT de
un 78 %. La combinación de concentraciones con la cual también se obtuvieron buenos
resultados de eficiencia fue (2.5 mg de FeSO4/L + 10 mL de H2O2/L), también con 180
min, con resultados de degradación de DQO y COT de 70% y 68% respectivamente. La
combinación de reactivos con menor eficiencia tanto en remoción de DQO como de COT
fue la de menores concentraciones (1 mg de FeSO4/L + 10 mL de H2O2/L) teniendo
737
resultados de remoción de 60% en DQO y 76% en COT. Aún así estos resultados son
buenos ya que la diferencia entre las dos combinaciones no es muy grande.
Aunque se esperaban mayores porcentajes de degradación tanto en DQO como en
COT se puede especular que la complejidad de las aguas residuales podría afectar la
oxidación de los compuestos por el reactivo de Fenton. Aun se trabaja en la optimización
de la reacción Fenton ya que la cantidad de reactivos usados para cada prueba se puede
considerar como altos y esto provoca que los costos en el proceso aumenten.
La muestra para las pruebas de Fenton presentaba una DQO de 800 mg/L logrando
una degradación de 300 mg/L a 150 mg/L dependiendo de las combinaciones de
reactivos.
GRAFICA 2. Degradación de DQO según la combinación y tiempo de la reacción
Fenton.
GRAFICA 3. Porcentaje de Remoción de las combinaciones de reactivos Fenton
dependiendo del tiempo
Electrofenton. El tratamiento Electrofenton fue aplicado al efluente del reactor UASB
para hacer una comparación de tratamientos junto con la oxidación Fenton. En la Tabla 2
se muestran las combinaciones utilizadas para este tratamiento.
738
TABLA 2. Combinaciones de parámetros usados en Electrofenton.
Tratamiento Concentración FeSO4 Densidad de Corriente Electrofenton Electrofenton Electrofenton Electrofenton 5 mM 5mM 5mM 7mM 7.81 mA/cm2 15.62 mA/cm2 31.25 mA/cm2 31.25 mA/cm2 Se determinó la eficiencia del tratamiento por DQO y COT. De las combinaciones
usadas para el tratamiento la que tuvo una mayor eficiencia de remoción en DQO como
en COT fue la que tenía mayor densidad de corriente con la mayor concentración de
FeSO4 (31.25 mA/cm2 + 7mM FeSO4) a un tiempo de 180 min con resultados de remoción
de 61% en DQO y 64% en COT, aunque en la Gráfica 4 se puede observar que en los
primeros minutos existe una gran remoción de materia orgánica y los siguientes minutos
existe remoción pero en menor cantidad por lo que se puede ver una tendencia lineal y
continua. La combinación que presentó eficiencias de remoción más bajas fue en la que
se usó la una densidad de corriente más pequeña con la concentración de FeSO 4 (7.81
mA/cm2 + 5mM) a los 180 min, con eficiencias de remoción de 52% y 45% de DQO y COT
respectivamente, en esta combinación también se puede observar una remoción alta en
los primeros minutos, pero a comparación de la combinación (31.25 mA/cm2 + 7mM
FeSO4), la remoción no es continua en los primeros minutos si no que empieza a mostrar
el comportamiento de remoción más estable a partir del minuto 120.
GRAFICA 4. Porcentaje de remoción según las combinaciones de parámetros
usados en Electrofenton.
Determinación de Hormonas en Aguas Residuales. La determinación de hormonas en
las muestras tanto del influente como del efluente del reactor UASB y de los tratamientos
de Fenton y Electrofenton se llevó a cabo mediante HPLC (High Performance Liquid
Chromatography). El trazador hormonal seleccionado para darle seguimiento a la
HOLPLQDFLyQGHKRUPRQDVHQDJXDVUHVLGXDOIXHȕ(VWUDGLROHQEDVHDODOLWHUDWXUDTXH
739
menciona que esta es una hormona natural producida en el organismo femenino con
grandes impactos tanto en la fauna como en el organismo humano, es una de las
hormonas naturales de las cuales se encuentran en mayor cantidad en aguas residuales.
Los análisis se realizaron a la muestra de agua en el influente del reactor UASB para
determinar si esta hormona se encontraba en el agua, por el detector usado en este
cromatógrafo no fue posible detectar concentraciones menores a 0.1 mg/L, por lo que se
tuvo la necesidad de aplicar la hormona a la muestra en una concentración conocida y
observar el comportamiento del reactor UASB y darle seguimiento a la degradación de
esta mediante los tratamientos aplicados.
Medidas analíticas. Se usó un HPLC (Modelo Agilent) equipado con un detector de red
de diodos (DAD) y Detector Uv-vis, y la columna de análisis. Se utilizó una mezcla de
metanol, agua, Acetonitrilo (23:53:24) como fase móvil, a un caudal 1mL/min. La
WHPSHUDWXUDGHODFROXPQDVHPDQWXYRDƕ&(OWLHPSRGHUHWHQFLyQIXHGHPLQ/D
PXHVWUDWRPDGDSDUDODLQ\HFFLyQIXHGHȝO
Se inició con la curva de calibración y la identificación de nuestro trazador hormonal
diluido en metanol, tomando como concentraciones iniciales 100, 50, 30, 15, 10, 5, 1, y
0.5 mg/L. Estas soluciones fueron preparadas en matraces volumétricos de 10 ml
tomando como solución madre la concentración de 100 mg/L y haciendo diluciones para
otras concentraciones.
*5$),&$&XUYDGHFDOLEUDFLyQFRQGLIHUHQWHVFRQFHQWUDFLRQHVȕ(VWUDGLRO
Extracción de las muestras. Los análisis de las muestras del influente y efluente se
sometieron antes a una extracción usando 100 mL para el efluente y 250 mL para el
influente, en la extracción se usaron cartuchos C18 que se acondicionaron antes de la
etapa de extracción usando metanol, acetonitrilo y agua. Se acondicionó pasando 8 mL
de acetonitrilo a través de la fase, a continuación, 7 mL de metanol y, finalmente, 5 mL de
agua. La extracción se llevó a cabo en un colector que estaba conectado a una bomba de
vacío con una presión y controlador de vacío. Para la hormona el caudal fue de 3-5
mL/min y una presión de vacío de 3 mm Hg. Antes de la etapa de elución, los cartuchos
se secaron al vacío (5 mm de Hg) durante aproximadamente veinte minutos para eliminar
todo el agua debido a que estas moléculas podrían producir bajas recuperaciones de
740
compuesto de la fase estacionaria. La elución de las hormonas fue hecha con dos partes
alícuotas (5 mL) de ACN y se combinaron en una botella ámbar (Vallejo y col., 2011).
Las muestras de los tratamientos de Fenton y Electrofenton fueron introducidas al HPLC
directamente ya que la carga orgánica era mucho menor y tenía menos interferencias en
la señal, los resultados de los análisis de HPLC se muestran en la gráfica 6.
Estradiol
GRAFICA 6. 6HxDOHVGHGHJUDGDFLyQGHȕ(VWUDGLROVHJ~QORVWUDWDPLHQWRV
Conclusiones
De acuerdo a la literatura, los ácidos de cadena corta pueden afectar el proceso de
degradación Fenton y electrofenton. Esto se puede ver reflejado en nuestros resultados
ya que la eficiencia de degradación tanto de Fenton como Electrofenton es constante
después de cierto tiempo, debido a que en el proceso biológico en el reactor UASB
aumenta la producción de ácidos grasos volátiles que son ácidos de cadena corta; por lo
que puede ser necesario usar algún otro proceso para romper los enlaces de esos
compuestos y de esta forma incrementar la eficiencia de nuestro proceso de depuración.
A pesar de que el proceso Fenton tiene buenos resultados de degradación es
recomendable hacer experimentación con otras concentraciones de reactivos más bajas
para optimizar el proceso ya que al usar concentraciones altas los costos aumentan y esto
puede provocar que nuestro proceso ya no sea económico ni viable.
La combinación de procesos utilizada en este trabajo para la remoción de hormonas en
DJXDVUHVLGXDOHVHVHILFLHQWHGHPRVWUDQGRODWRWDOUHPRFLyQGHODKRUPRQDȕ(VWUDGLRO
y una disminución considerable de materia orgánica en el medio.
741
Referencias
Bett, J. W., Lester, J. N. (2003) Endocrine disrupters in wastewater and sludge treatment processes pp 4- 8
Bigsby, R., Chapin, R., Daston, G. P., Davis, B., Gorski, J., Gray, L. E., Howdeshell, K. L., R. Zoeller,T., and
Vom Saal, F. S. (1999). Evaluating the Effects of Endocrine Disruptors on Endocrine Function during
Development, Environmental Health Perspectives pp. 613-618.
Comninellis C., Kapalka A., Malato S., Parsons S.A., Poulios I., Mantzavinos D.(2008).'Advanced oxidation
processes for water treatment: advances and trends for R&D', Journal of Chemical Technology and
Biotechnology, 83, 769-776
Gogate P.R., Pandit A.B., (2004) ³$UHYLHZRILPSHUDWLYHWHFKQRORJLHVIRUZDVWHZDWHUWUHDWPHQW,R[LGDWLRQ
WHFKQRORJLHVDWDPELHQWFRQGLWLRQV´, Advances in Environmental Research pp.525-530
Jobling, S., Nolan, M. Tyler, C. R., Brighty, G., Sumpter, J. P. (1998) Wildespread sexual disruption in wild
fish. Environ. Sci. Technol.pp. 2498í2506.
Naimi, I., Bellakhal N. (2012). Removal of 17ȕ-Estradiol by Electro-Fenton Process. Materials Sciences and
Applications, 3, 880-886
Rodríguez Fernández-Alba, A., Letón García, P., Rosal García, R., Dorado Valiño, M., Villar Fernández, S.
y Sanz García., J. M. (2006). Tratamientos avanzados de aguas residuales industriales. Fundación
para el conocimiento Madrid CEIM. pp.30-31.
6LUWRUL & =DSDWD $ 2OOHU , *HUQMDN : $JHUD $ 0DODWR 6 µ'HFRQWDPLQDWLRQ RI LQGXVWULDO
pharmaceutical wastewater by combining solar photo-)HQWRQ DQG ELRORJLFDO WUHDWPHQW¶ :DWHU
Research, 43, 661-668.
Thrope, K. L; Cummingsm, R. L.; Tutchinson, T. H.; Scholz, M.; Brighty, G.; Sumpter, J. P.; Tyler, C. R.
(2003) Relative potencies and combination effects of steroidal estrogens in fish. Environ. Sci. Technol,
37, 1142í1149.
Vallejo-Rodríguez., R., Lopez-Lopez, A., Saldarriaga-Noreña, H., Murillo-Tovar, M., Hernández-Mena.,(
2011) L. Optimization of Analytical Conditions to Determine Steroids and Pharmaceuticals Drugs in
Water Samples Using Solid Phase-Extraction and HPLC. American Journal of Analytical Chemistry
pp.864-865
:DOOLQJ&µ)HQWRQ¶VUHDJHQWUHYLVLWHG¶$FFChem. Res.pp.125±131.
Lista de abreviaturas
UASB
HPLC
COT
DQO
ACN
Up flow anaerobic sludge blanket (Reactor anaerobio de flujo ascendente)
High performance liquid chromatography (Cromatografía de líquidos de alta
resolución)
Carbono Orgánico Total
Demanda química de Oxigeno
Acetonitrilo
742
CHAPTER 7.6. ESTRUCTURA DE INSTRUMENTOS TECNOLÓGICOS EN EL
TRATAMIENTO DE AGUAS RESIDUALES EN MÉXICO
Lilian E. Domínguez-Montero (1); Blanca Jiménez-Cisneros (2);
Rosa O. Cañizares-Villanueva (1,3); Sergio Caffarel-Méndez (4);
Eugenio Frixione-Garduño (1); Miguel A. Pérez-Angón (1);
Héctor M. Poggi-Varaldo*(1,3)
(1) CINVESTAV del IPN, Programa Transdisciplinario en Desarrollo Científico y Tecnológico para la
Sociedad
(2) II-UNAM, México DF, México;
(3) CINVESTAV del IPN, Dept. de Biotecnología y Bioingeniería, México DF, México;
(4) TESE, Div. Ing. Qca. y Bqca., Ecatepec, Edo. de Méx, México
RESUMEN
El desarrollo y adaptación de tecnología es de vital importancia para que el tratamiento de
aguas residuales (TAR) sea eficiente, fácil de implementar y con bajos costos de
construcción y operación. Las patentes pueden ser un indicador útil para medir y analizar
dichos avances tecnológicos. El propósito de este trabajo fue evaluar el desarrollo
tecnológico existente en México relacionado con el TAR. En particular, analizamos la
tecnología patentada mexicana de origen internacional y autóctona. Se llevó a cabo la
revisión de bases de datos de patentes en los últimos 18 años relacionados con el
tratamiento de aguas residuales. Hemos examinado las bases de datos del Instituto
Mexicano de la Propiedad Industrial (IMPI) y la Organización Mundial de la Propiedad
Intelectual (OMPI). El período analizado se definió teniendo en cuenta todas las normas
de aguas residuales desde 1996. Los resultados muestran que la tecnología patentada en
México para TAR es escasa en comparación con el volumen y la variedad de las patentes
correspondientes a otros países del mundo. México sólo ha contribuido con 34 patentes
de cada 5,041 (0.67%) en el período de 1996 a 2013 en función de la base de datos de la
OMPI. Además, se determinó que la tecnología en patentes mexicanas es una proporción
menor de todas las patentes en TAR. De 1996 a 2013, en la base de datos del IMPI se
registraron 137 patentes, de las cuales 42 pertenecen a empresas mexicanas,
universidades, centros de investigación e inventores particulares (30.65%). Esto sugiere
que el desarrollo tecnológico en el sector de TAR en México, puede ser considerado débil
y depende principalmente de la tecnología extranjera y paquetes tecnológicos de origen
externo.
Palabras Clave: México, patentes,tratamiento de aguas residuales.
-----------*Author for all correspondence: [email protected]
743
ABSTRACT
The development and adaptation of technology is vital for the wastewater treatment
(WWT) be efficient, easy to implement, and with low costs of construction and operation.
Patents can be a useful indicator for measuring and analyzing these technological issues.
The purpose of this study was to evaluate the existing technological development in
Mexico related to WWT. Particularly, we analyzed Mexican patented technology of
international and autochthonous origin. We examined the databases of the Mexican
Institute of Industrial Property (IMPI) and World Intellectual Property Organization (WIPO).
The analyzed period was defined taking into account all standards wastewater since 1996.
The results show that the patented technology in Mexico for WWT is scarce compared to
the volume and variety of patents for other countries in the world. Mexico has contributed
only with 34 patents out of 5,041 (0.67%) in the period 1996 to 2013 according to the
database of WIPO. Furthermore, it was determined that the Mexican technology in
Mexican patents is a minor proportion of all the Mexican patents in WWT. Since 1996 to
2013, in the database of IMPI were registered 137patents, of which 42 belong to Mexican
companies, institutions, research centers and inventors (31.82%). This suggests that the
technological development in the WWT sector in Mexico can be considered weak and is
mostly dependent on foreign technology and technological packages originated abroad.
Key words: Mexico, patents, wastewater treatment.
Introducción
En las últimas décadas, la contaminación del agua se ha convertido en un problema que
la humanidad trata de controlar cada día. De las aguas superficiales de ríos y lagos, la
mitad se encuentran contaminadas, y solo una tercera parte de estas aguas son de buena
calidad. Por ello, de 122 países evaluados, México ocupa el lugar 106 en cuanto a la
calidad del agua (Ortiz Rendón, 2008), por ello se han desarrollo nuevas tecnologías para
reducir dicha contaminación y su impacto ambiental.
Los Tratamientos de Aguas Residuales (TAR) tienen como propósito remover el
material contaminante, orgánico e inorgánico que tiene el agua residual, el cual puede
estar en forma de partículas en suspensión y/o disueltas, con el objetivo de alcanzar una
calidad de agua que cumpla con la normativa de descarga o por el tipo de reutilización a
la que se destinará (Noyola, et al., 2013).
En los últimos años, los TAR han recibido un gran interés por parte de especialistas
en el tema de saneamiento del agua, así como de las autoridades encargadas del manejo
del agua (De la Peña, et al., 2013), y hoy en día se cuenta con diversas tecnologías como
son las convencionales altamente mecanizadas hasta tecnología ecológica de bajo costo
(Zurita-Martínez, et al., 2011).
Un estudio estratégico del desarrollo de las invenciones en un campo dado, se
realiza a través de la información de patentes, las cuales contienen información jurídica,
económica y técnica de la invención. Diversos autores como Lanjouw y Mody, Jaffe y
Palmer y Popp han estudiado el desarrollo de invenciones en medio ambiente, a través
del análisis de patentes (Ramírez Guardado, 2013).
744
8QDSDWHQWHGHDFXHUGRDOD2UJDQL]DFLyQ0XQGLDOGHOD3URSLHGDG,QWHOHFWXDOHV³un
derecho exclusivo concedido a una invención, es decir, un producto o procedimiento que
aporta, en general, una nueva manera de hacer algo o una nueva solución técnica a un
problema´
Actualmente, la infraestructura y tecnología empleada en los TAR en México, en su
mayoría no es nacional, y además, existe un rezago tecnológico, por lo el estudio de
patentes permite conocer el desarrollo de nuevas tecnologías, y saber sobre los avances
que hay en el tema (Ramos , et al., 2005).
Materiales y Método
La fuente de información que conformo este estudio, se consultó en dos bases de datos
de documentos de patentes de libre acceso en Internet, la primera fue la base de la
Organización Mundial de la Propiedad Intelectual (OMPI), y la base del Instituto Mexicano
de la Propiedad Industrial (IMPI). El método utilizado se centró en la estrategia de
búsqueda trazada, que tuvo como premisa la definición del objeto, objetivo y la estrategia
de búsqueda.
Objeto de búsqueda: Tratamiento de Aguas Residuales (TAR).
Objetivo de la búsqueda: Evaluar el desarrollo las invenciones y tecnologías para el
TAR existentes en México. En particular, analizar la tecnología patentada mexicana de
origen internacional y autóctona, en el periodo 1996-2013.
Estrategia de Búsqueda: Se centró en la búsqueda por palabra clave en el campo de
título y de resumen. El período analizado se definió teniendo en cuenta el conjunto de
normas relevantes a aguas residuales que se actualizó significativamente en 1996 para
México.
Por otro lado, los datos que se extrajeron de los documentos de patentes son: la
fecha de publicación, el número de solicitud, el solicitante de la patente, el país del
solicitante de la patente y el título de la patente.
A partir de las bases de datos obtenidas, se realizaron dos categorías independientes a
las proporcionadas por OMPI e IMPI. La primera categoría es el tipo de solicitante, donde
se permite analizar quienes se encargan de realizar las investigaciones e invenciones de
las patentes; en esta categoría se incluyeron 4 agrupaciones: universidad, centro de
investigación, empresa e inventor particular. La segunda categoría pertenece al objeto de
invención de la patente, la cual determina el alcance de protección de las invenciones, así
como permite ver en qué trabaja la comunidad vinculada al tratamiento de aguas
residuales; se incluyeron 5 agrupaciones en dicha categoría: equipo, equipo y
procedimiento, equipo y uso, procedimiento y procedimiento y uso.
Los resultados de las patentes aceptadas para TAR durante el periodo 1996-2013, en la
base de datos OMPI, son 5,041. Se observa que la evolución de dichas patentes se ha
incrementado (Figura 1 a., ya que durante el año de 1996 se contaba con 36 invenciones,
745
mientras que para el año 2012 se obtuvieron 634 patentes. Durante el 2013, se han
registrado 261 patentes, sin embargo, es importante señalar que el número puede
incrementarse, ya que existen invenciones solicitadas que se encuentran en estado de
aplicación.
FIGURA 1. Evolución de las patentes mundiales para TAR, 1996-2013. Fuente:
Elaboración propia con datos de OMPI
El país con mayor número de patentes otorgadas, de acuerdo a OMPI, es la República de
Corea con 1,524 invenciones, lo que representa el 30.23% de la producción mundial en
patentes para TAR; mientras que México cuenta con solo 34 patentes otorgadas, por lo
que su contribución es del 0.67%.
La evolución de las 34 patentes de México en el mismo periodo ha sido aleatoria, y
además logra verse que el patentamiento es nulo en algunos años (Figura 2).
FIGURA 2. Evolución de las patentes otorgadas a México para TAR, 1996-2013.
Fuente: Elaboración propia con datos de OMPI
746
A pesar de que México tiene 34 patentes concedidas para TAR, el principal país que
patenta en México es Estados Unidos con 20 patentes otorgadas. Mientras que 7
patentes corresponden a solicitantes de nacionalidad mexicana (residentes); por lo que la
aportación real de invenciones de México es solo de 0.14%.
Al realizar la comparación de la evolución de las patentes de México (34), con las
patentes de residentes mexicanos (7), se observa que en el periodo 2008-2010 los
mexicanos realizaron la mayor aportación de patentes (Figura 3).
FIGURA 3. Comparación de las patentes otorgadas a México y a los residentes
mexicanos. Fuente: Elaboración propia con datos de OMPI
En las patentes otorgadas a residentes mexicanos se decidió establecer dos categorías,
el tipo de solicitante y el objeto de protección (Tabla 1). En la categoría de tipo de
solicitante, se puede observar que los inventores particulares son los que principalmente
realizan dicho trámite, los cuales representan el 57.14%. Al analizar los objetos de
protección que los mexicanos trabajan más en cuanto a TAR, se observa que la
WHQGHQFLD HV SURWHJHU WDQWR HO ³SURFHGLPLHQWR \ XVR´ DVt FRPR HO GHVDUUROOR GH QXHYRs
³HTXLSRV´\DTXHFDGDDJUXSDFLyQWLHQHSDWHQWHVUHJLVWUDGDV
TABLA 1. Tendencia por tipo de solicitante y objeto de protección de las patentes
para TAR.
747
Tipo de
Objeto de
Solicitante
Protección
Centro
de Equipo
Investigación
Empresa
Equipo y Uso
Procedimiento
Inventor Particular
Equipo
Equipo
Procedimiento
Procedimiento
Uso
No.
Patentes
1
y
1
1
1
1
y
2
En materia de patentes otorgadas, la oficina de patentes de México, IMPI, ha otorgado
137 patentes para TAR, en el mismo periodo (1996-2013). La Figura 4, muestra que la
evolución de patentamiento en dicho periodo ha sido aleatoria, aunque ha existido un leve
incremento en los últimos 7 años.
FIGURA 4. Evolución de las patentes para TAR de la oficina de México, IMPI.
Fuente: Elaboración propia con datos de IMPI
De manera muy similar a los resultados de la base de datos OMPI, México tiene un papel
activo al otorgar el 69.34% de la patentes a inventores, instituciones o empresas
extranjeras, mientras que el 30.66% (42 patentes) pertenecen a residentes de
nacionalidad mexicana (Figura 5).
748
FIGURA 5. Países con patentes otorgadas para TAR en México, 1996-2013. Fuente:
Elaboración propia con datos de IMPI
Al analizar las 42 patentes otorgadas a residentes mexicanos, se observa que el inventor
particular es el principal solicitante de patentes, ya que tiene concedidas 22 de éstas.
Además los resultados evidenciaron que sus principales objetos de protección son el
³SURFHGLPLHQWR´DOLJXDOTXHHO³HTXLSR\SURFHGLPLHQWR´HVLPSRUWDQWHVHxDODUTXHHVWD
última agrupación es importante ya que le permite al dueño de la patente tener una mejor
transferencia tecnológica (Figura 6).
FIGURA 6. Patentes otorgadas para TAR a residentes mexicanos por tipo de
solicitante y objeto de protección, 1996-2013. Fuente: Elaboración propia con datos
de IMPI
749
Cabe destacar que al analizar las patentes otorgadas para TAR de las bases de datos
OMPI e IMPI, se muestra que la cantidad de patentes que tiene México sobre el tema es
menor a las patentes internacionales registradas en OMPI, ya que solo las 137 patentes
de IMPI representarían un 2.72% de las patentes registradas en OMPI (5,041).
FIGURA 7. Comparación de los países con patentes otorgadas para TAR en México,
1996-2013. Fuente: Elaboración propia con datos de IMPI y OMPI.
Como se observa en la Figura 7, en la concesión de patentes en México, OMPI como
IMPI tienen un papel similar, ya que en ambas bases de datos se otorga un mayor
número de patentes a otros países, mientras que los residentes mexicanos ocupan el
segundo lugar en otorgamiento de patentes, y el país que predomina con patentes para
TAR es Estados Unidos.
FIGURA 8. Tendencia por tipo de
solicitante de las patentes otorgadas de
residentes
mexicanos.
Fuente:
Elaboración propia con datos de IMPI y
OMPI.
FIGURA 9.Tendencia por objeto de
protección de las patentes otorgadas de
residentes
mexicanos.
Fuente:
Elaboración propia con datos de IMPI y
OMPI.
750
Los resultados mostrados en la Figura 8 evidencian que solo 3 de las patentes de
residentes mexicanos estan registradas tanto en OMPI como en IMPI; dos de ellas
pertenecen a inventores particulares, y una patente a empresa. Mientras que la Figura 9
se observa que el objeto de protección no es homogéneo, ya que de las tres patentes
se estan protegiendo diferentes aspectos: procedimiento y uso; equipo y uso; y equipo.
Conclusiones
El análisis de este trabajo, nos permite observar como se han ido desarrollando las
patentes respecto a las invenciones y tecnologías sobre el tratamiento de aguas
residuales en nuestro país durante el periodo de 1996-2013. Se ha demostrado que la
tecnología patentada en México para TAR es escasa en comparación con el volumen y
la variedad de las patentes correspondientes a otros países del mundo. La mayor parte
de las patentes que tiene México resgistradas pertenecen a solicitantes extranjeros, y
se determinó que el componente de tecnología patentada mexicana solo es un 30% de
origen autóctono. Por ello, el desarrollo tecnológico en el sector de TAR en México
puede ser considerado débil y dependiente principalmente de la tecnología extranjera y
paquetes tecnológicos de origen externo. Esto, a su vez, es un fuerte indicador para las
autoridades gubernamentales. De hecho, puede ser que deseen implementar
programas de estímulos a niveles empresariales y educativos, en México con el fin de
fomentar la generación y aplicación de la tecnología nacional propia en TAR.
Referencias
De la Peña, M. E.; Ducci, J.; Zamora Plascencia, V. (2013). Tratamiento de aguas residuales en México.
Banco Interamericano de Desarrollo.
García, B.; Corvo, F. (2007) Patent Information in Concrete Corrosion Research. Revista CENIC Ciencias
Químicas. 38: 410-414.
Meyer, M.; Timm Utecht, J.; Goloibeva, T. (2003). Free patent information as a resource for policy
analysis. World Patent Information. 25: 223-231.
Noyola, A.; Morgan-Sagastume, J. M.; Güereca, L. P. (2013). Sección de tecnologías para el tratamiento
de aguas residuales municipales: Guía de apoyo para ciudades pequeñas y medianas. 1ª ed.
Universidad Autónoma de México-Instituto de Ingeniería. MX.
Organización
Mundial
de
la
Propiedad
Intelectual
(2014).
http://www.wipo.int/patentscope/es/patents_faq.html#patent; consulta 12 junio 2014.
Ortiz Rendón, G. A. (2008). Evolución y perspectiva del marco jurídico del agua en México: Nuevos retos
y oportunidades para la gestión integrada del recurso hídrico. En: Arriaga García, C. B.; Rabasa, E.
O. Agua: Aspectos Constitucionales. 1ª ed., Universidad Autónoma de México. 17-52.
Ramírez Guardado, P. (2013). Política hídrica y cambio tecnológico en tecnologías aplicadas al
tratamiento de las aguas residuales. Journal of technology management &innovation. 8: 64-75.
Ramos, G.; García, B. (2005). Tendencias en el desarrollo de nuevas tecnologías para el tratamiento de
aguas, residuos y lodos. Revista CENIC Ciencias Biológicas. 36.
Ramos Muñiz, G.; García Delgado, B.; Tundidor Camba, A.; Bataller, M. (2005). Vigilancia Tecnologica e
Información de Patentes: Su papel en el enfrentamiento a la sequía. Revista CENIC Ciencia
Químicas. 36.
Zurita-Martínez, F.; Castellanos-Hernández, O.; Rodríguez-Sahagun, A. (2011). El tratamiento de aguas
residuales municipales en las comunidades rurales de México. Revista Mexicana de Ciencias
Agrícolas. 1: 139-150.
751
Bases de Datos de Patentes
Base
de datos del Instituto Mexicano de la Propiedad Intelectual,
http://siga.impi.gob.mx/#busqueda; consulta marzo 2014.
Base de datos de la Organización Mundial de la Propiedad Intelectual,
http://patentscope.wipo.int/search/en/result.jsf; consulta julio 2014.
IMPI
(2014).
OMPI
(2014).
Notación
IMPI
OMPI
TAR
WIPO
WWT
Instituto Mexicano de la Propiedad Industrial
Organización Mundial de la Propiedad Intelectual
Tratamiento de Aguas Residuales
World Intellectual Property Organization
Wastewater Treatment
752
CHAPTER 7.7. REMOCIÓN DE AZUL ÍNDIGO POR LA BIOMASA SECA DE
CUATRO CEPAS DE Trichoderma sp.
Erika Sánchez Torres (1); Oscar García Barradas (1);
Ma. Remedios Mendoza López (1); Rosalba Argumedo-Delira*(1)
(1) Universidad Veracruzana, Unidad de Servicios de Apoyo en Resolución Analítica (SARA), Xalapa,
Veracruz, México.
RESUMEN
El colorante azul índigo es ampliamente utilizado en la industria textil, ya que es usado
para teñir mezclilla, la cual es empleada para la confección de pantalones, siendo
México uno de los principales productores de mezclilla. No obstante, el uso de estas
prendas por influencia de la moda deja una huella de contaminación que afecta a los
ecosistemas acuáticos, debido a la liberación de efluentes coloreados con azul índigo a
los ríos mexicanos, primordialmente por textileras informales que no cuentan con
plantas tratadoras de aguas residuales. Con el fin de generar información básica de la
interacción Trichoderma-azul índigo que pueda ayudar a generar biotecnologías de
bajos costos, útiles para el tratamiento de efluentes contaminados con azul índigo, la
presente investigación tuvo como objetivo determinar la capacidad de biosorción de la
biomasa seca de cuatro cepas de Trichoderma sp. (Trichoderma sp. H-1, Trichoderma
sp. H-2, Trichoderma sp. H-11 y Trichoderma sp. H-12.) ante cinco concentraciones del
colorante azul índigo (9, 19, 28, 33 y 39 mg L-1) durante 24 h, 25 °C y pH=1.5. Los
resultados muestran que la biomasa seca de Trichoderma sp. H-11 tuvo una remoción
del 100% en todas las concentraciones probadas, mientras que la biomasa seca de
Trichoderma sp. H-1 (95%), Trichoderma sp. H-2 (83%) y Trichoderma sp. H-12 (79%)
presentaron una remoción promedio menor. Los datos obtenidos tuvieron una buena
correlación con las isotermas de adsorción de Langmuir y Freundlich, encontrando que
la capacidad máxima de adsorción (qm) para la biomasa seca de Trichoderma sp. H-2,
Trichoderma sp. H-12 y Trichoderma sp. H-1 fue de 57.3 mg g-1, 27.1 mg g-1 y 3.9 mg
g-1 respectivamente. En el caso de la cinética de adsorción los datos se describieron
mejor con el modelo cinético de pseudo segundo orden, la constante de velocidad de
pseudo segundo orden (K2) indicó que la biomasa seca de Trichoderma sp. H-1
adsorbe en menor tiempo al colorante azul índigo bajo las condiciones probadas.
Finalmente la biomasa seca de la cepa Trichoderma sp. H-11 demostró ser un eficiente
biosorbente fúngico ante el colorante azul índigo, lo que indica el amplio potencial de
las especies fúngicas del género Trichoderma para el tratamiento de efluentes
contaminados con colorantes.
Palabras clave: azul índigo, biomasa fúngica, biosorción, Trichoderma.
-----------------*Author for correspondence
753
Introducción
El azul índigo es uno de los colorantes más antiguos utilizados por el hombre, su
enorme consumo hoy en día es debida a la popularidad de los pantalones vaqueros
azules, que se tiñen con este colorante, ya que el colorante es capaz de desvanecerse
para dar tonos azules cada vez más pálidos, tendencia que sigue la moda actual
(Vuorema, 2008). A lo largo de la historia el colorante azul índigo se ha obtenido de
varias plantas como Indigofera tinctoria (especie nativa de la India), Indigosfera
suffruticosa Mill (especie abundante en el Caribe, Centro América y América del Sur),
pero en el siglo XIX llegaron los colorantes sintéticos y actualmente el colorante azul
índigo se sintetiza principalmente a partir de subproductos de combustibles fósiles
(Clark et al., 1993).
El consumo del colorante azul índigo y otros colorantes indigoides por la industria
textil alcanza unos 33 millones de kilogramos por año, lo cual ha causado preocupación
ambiental, debido a que el colorante azul índigo se considera una sustancia
recalcitrante que afecta a los ecosistemas acuáticos (Balan y Monteiro, 2001; Roessler
y Jin, 2003). En México se ha reportado la contaminación de ríos localizados en
Tlaxcala y Puebla por descargas con azul índigo que han causado la muerte de la flora
y fauna acuática (García y Solís-Fuentes, 2008; Solís et al., 2013). Para tratar de dar
solución a esta problemática se han reportado técnicas electroquímicas, tal es el caso
de Sanromán et al. (2005) quienes reportan que el colorante azul índigo (5,000 mg L-1)
es decolorado en un 76% en 45 minutos, no obstante el gasto energético es
considerable (16 W h). Otro método electroquímico empleado para la decoloración de
índigo es la electroincineración, el cual es capaz de decolorar completamente altas
concentraciones (226,270 mg L-1) de azul índigo en 3 h y sin formación de lodos (SolísOba et al., 2009).
Con respecto a los métodos químicos Manu (2007) reporta altas remociones del
colorante azul índigo (500 mg L-1) empleando 2.0 g L-1 de alumbre (97%), 2.5 g L-1 de
limo (97%) y 2.5 g L-1 de sulfato ferroso (95%) como coagulantes en un tanque que
contenía efluentes de este colorante. Sin embargo, Unlu et al. (2009) mencionan que la
coagulación con sulfato de aluminio y cloruro férrico no es efectiva como un pretratamiento para efluentes con azul índigo provenientes de la industria textil, debido a
las altas dosis requeridas del coagulante.
Dentro de los procesos biológicos reportados para la decoloración de azul índigo
con hongos filamentosos, ha sido reportada la decoloración de azul índigo por
Phanerochaete chrysosporium, en dicho reporte se expone al hongo a 100, 200 y 400
mg L-1 del colorante, obteniendo una decoloración del 96.5, 97.5 y 95%
respectivamente (Cing et al., 2003). Mientras tanto Balan y Monteiro (2001) informan
sobre la decoloración de azul índigo (200 mg L-1) por los hongos Phellinus gilvus
(100%), Pleurotus sajor-caju (94%), Pycnoporus sanguineus (91%) y P. chrysosporium
(75%), la cual es llevada a cabo después de 4 días de incubación. Por otra parte, en lo
referente a la utilización de biomasa fúngica para la remoción de azul índigo, Khelifi et
al. (2013) reportan la biosorción de azul índigo (100 mg L -1) por la biomasa fresca
muerta de Aspergillus alliaceus, mostrando una capacidad máxima de adsorción (qm)
de 50 mg g-1 a pH 4 y 30 °C.
Considerando lo anterior, el presente estudio tuvo como objetivo determinar la
capacidad de biosorción de la biomasa seca de cuatro cepas de Trichoderma sp.
754
(Trichoderma sp. H-1, Trichoderma sp. H-2, Trichoderma sp. H-11 y Trichoderma sp. H12) ante cinco concentraciones del colorante azul índigo (9, 19, 28, 33 y 39 mg L-1)
durante 24 h, 25 °C y pH=1.5, con el fin de generar información básica que pueda
ayudar a generar biotecnologías de bajos costos, útiles para el tratamiento de efluentes
contaminados con azul índigo.
Microorganismos. Trichoderma sp. H-1 se aisló de suelo rizosférico de helecho de la
USBI, Xalapa-Veracruz México, mientras que Trichoderma sp. H-2 se aisló de suelo
(con residuos de pilas, computadoras y placas de circuito impreso) de los alrededores
del relleno sanitario el Tronconal, Xalapa-Veracruz México y Trichoderma sp. H-12 se
aisló de una placa de circuito impreso proveniente del mismo suelo. Entre tanto,
Trichoderma sp. H-11 se aisló de suelo la rizosféra de liquidámbar sp. del Parque
Natura, Xalapa-Veracruz México.
Remoción de azul indigo por la biomasa seca de cuatro cepas de Trichoderma.
Las cepas fúngicas Trichoderma sp. H-1, Trichoderma sp. H-2, Trichoderma sp. H-11 y
Trichoderma sp. H-12 se crecieron en cajas de Petri con PDA (Bioxon®) a 25 ± 2 °C por
5 días. Después se cortaron discos individuales de PDA con micelio de cada hongo (7
mm de diámetro). Posteriormente a biorreactores de plástico de 2.5 L de capacidad, se
le adicionaron 750 mL del medio mineral estéril (g L -1) 0.25 CaCl2; 0.25 MnCl2; 0.15
MgCl2; 1.5 (NH4)2SO4; 0.5 NaH2PO4; 7.5 Glucosa; 1.5 extracto de levadura; pH 5.
Después a cada respectivo reactor se le agregaron 20 discos de cada respectiva cepa
de Trichoderma, los cultivos se incubaron a temperatura ambiente durante 7 días. Al
término de la incubación, el micelio fúngico se separó del medio de cultivo mediante
filtración al vació, luego se secó a temperatura ambiente por 96 h para determinar su
peso seco y después el micelio fúngico se molió en un mortero.
Los experimentos de remoción se llevaron a cabo por la agitación de 0.2 g del
respectivo adsorbente (biomasa seca Trichoderma sp. H-1, Trichoderma sp. H-2,
Trichoderma sp. H-11 y Trichoderma sp. H-12) con 10 mL de azul indigo a las
siguientes concentraciones: 0, 9, 19, 28, 33 y 39 mg L -1 respectivamente, a 25°C y 500
rpm en un agitador para tubos (Multi-tube Vortexer VWR®). La concentración de cada
colorante se estimó espectrofotométricamente mediante el monitoreo de la absorbancia
del azul índigo utilizando un espectrofotómetro (Milton Roy Company, Spectronic 20D®)
y la capacidad de adsorción se determinó mediante la ecuación 1.
ൌ
୚ሺେ౟ ିେ౛ ሻ
୑
(1)
Donde, Q es la cantidad de colorante adsorbido (mg L -1), V representa el volumen
de la solución del colorante (L), Ci y Ce son la concentración inicial y en equilibrio del
colorante (mg L-1) y M es el peso seco de biomasa fúngica (g).
Por otra parte, para evaluar la capacidad de adsorción del adsorbente se
emplearon las isotermas de Langmuir (1916) y Freundlich (1906) con las siguientes
ecuaciones:
755
Modelo Langmuir
୯ ୏ େ
(2)
ౣ ై ౛
ୣ ൌ ଵା୏
େ
ై ౛
Modelo Freundlich
భ
(3)
ୣ ൌ ୊ ୣ౤
Donde qm es la capacidad máxima (mg g-1) y KL es el coeficiente de afinidad (L
mg-1), mientras que qe (mg g-1) representa la capacidad de adsorción en el equilibrio.
KF y n representan los coeficientes de Freundlich (n, a dimensional; KF mg1-1/n g-1 11/n).
En lo referente a la cinética de adsorción, se realizaron cuatro muestreos cada tres
horas para una concentración de 39 mg L-1 de azul índigo y los datos se analizaron con
los modelos de pseudo-primer y pseudo-segundo orden. El modelo de pseudo-primer
orden en su forma lineal descrito por Lagergren (1898):
୏
(4)
భ
ሺ ୣ െ ୲ ሻ ൌ ሺ ୣ ሻ െ ଶǤଷ଴ଷ

Donde qt es la cantidad de adsorbato adsorbido por tiempo t (mg g-1) y K1 es la
constante pseudo-primer orden (min-1). Mientras que la forma lineal del modelo cinético
pseudo-segundo orden descrito por Ho (1999):
୲
୯ ൌ ୏
౪
ଵ
మ
మ ୯౛
ଵ
൅୯
౛
(5)
Donde K2 es la constante pseudo-segundo orden (g mg-1 min-1).
Remoción de azul indigo por la biomasa seca de cuatro cepas de Trichoderma. En
cuanto a la remoción de azul índigo a las 24 h por la biomasa seca de Trichoderma sp.
H-11 (Figura 1) se observó una remoción del 100% del colorante ante las cinco
concentraciones probadas (9, 19, 28, 33 y 39 mg L-1), lo cual no permitió realizar las
isotermas de adsorción de Langmuir y Freundlich para esta cepa en particular. Mientras
que Trichoderma sp. H-1 tuvo una remoción promedio del 95%, Trichoderma sp. H-2 del
83% y Trichoderma sp. H-12 del 79%. Los datos obtenidos para la remoción de azul
índigo por la biomasa seca de las cepas de Trichoderma sp. H-1, H-2 y H-12, se
analizaron con las isotermas de adsorción de Langmuir (Ecuación 2) y Freundlich
(Ecuación 3).
756
FIGURA 1. Remoción de azul índigo por la biomasa seca de la cepa Trichoderma
sp. H-11 a 25 °C y pH 1.5.
En ambos modelos matemáticos los datos experimentales se describen
correctamente bajo las condiciones empleadas, es importante mencionar que la cepa
de Trichoderma sp. H-2 mostro mejor correlación con el modelo de Langmuir tipo II
(Figura 2). Los resultados de estos gráficos se muestran en la Tabla 1. Para el caso de
la biomasa seca de Trichoderma sp. H-1 la constante de Freundlich KF es mayor
(0.969) en comparación con Trichoderma sp. H-2 (0.061) y Trichoderma sp. H-12
(0.003), mientras que el valor del coeficiente de Freundlich también es mayor para la
biomasa seca de Trichoderma sp. H-1 ante azul índigo. En el caso de la isoterma de
Langmuir, la capacidad de saturación de la monocapa para la biomasa seca de
Trichoderma sp. H-2 (57.271) mostró ser mejor ante azul índigo, en comparación con
las otras biomasas probadas.
TABLA 1. Constantes de las Isotermas de Freundlich y Langmuir para la biomasa
seca de tres cepas de Trichoderma sp. a 25 °C y pH 1.5.
Freundlich
Langmuir
Adsorbente
KF (mg1-1/ng111/n)
n
R2
qm (mg g-1)
KL (L mg-1)
R2
Trichoderma sp. H-1
0.969
1.092
0.966
3.930
0.396
0.994
Trichoderma sp. H-2
0.061
0.095
0.968
57.271
0.019
1.000
Trichoderma sp. H-12
0.003
0.532
0.963
27.144
0.007
0.985
757
FIGURA 2. Isotermas de adsorción para las tres cepas fúngicas de Trichoderma, a
25 °C y pH 1.5. (a) Langmuir tipo I, (b) Langmuir tipo II, (c) Freundlich.
En el caso de los ensayos para determinar la cinética de adsorción se encontró
que la remoción de azul índigo (39 mg L-1) por la biomasa seca de Trichoderma sp. H-2
(98%) y Trichoderma sp. H-12 (92%) a las 12 h mostraron ser mayores a las
remociones obtenidas a las 24 h (Figura 3). Lo cual se atribuyó a que la biomasa
obtenida para las isotermas de adsorción, presentó algunos cambios de color
provocada por la variabilidad en la esporulación de las cepas fúngicas.
Por otra parte, la cinética de adsorción del colorante azul índigo se analizó
empleando el modelo de pseudo primer orden y pseudo segundo orden, y se observó
que los datos experimentales se ajustaron mejor al segundo modelo. Los valores para
las constantes cinética pseudo-segundo orden K2 y la capacidad de adsorción en el
equilibrio qe se calcularon del grafico t/qt contra t (Figura 4). Los valores de qe y K2
demuestran que Trichoderma sp. H-1 es el hongo que adsorbe al azul índigo en menor
tiempo bajo las condiciones probadas (Tabla 2).
En la literatura se ha demostrado la remoción de efluentes con azul índigo
empleando diferentes biosólidos activados como carbón de hueso de vaca, carbón de
alga (182 mg g-1), carbón de lechuga de agua (235 mg g-1), carbón de pasto elefante
(167 mg g-1) y carbón de caparazón de cangrejo (111 mg g-1). Encontrando que el
carbón de hueso de vaca (571 mg g-1) tiene una saturación de la monocapa mayor que
la de los otros adsorbentes, sin embargo el carbón activado grado comercial (1000 mg
g-1) es mucho mejor que los biosólidos mencionados (Adetuyi y Jabar, 2011).
758
FIGURA 3. Remoción de azul índigo (39 mg L-1) a diferentes tiempos por la
biomasa seca de cuatro cepas de Trichoderma sp., a 25 °C y pH 1.5.
.
FIGURA 4. Cinética de adsorción de pseudo segundo orden para la biomasa seca
de cuatro cepas de Trichoderma sp., ante 39 mg L-1 de azul índigo a 25°C y pH 1.5.
TABLA 2. Constantes del modelo cinético de pseudo segundo orden para la
biomasa seca de cuatro cepas de Trichoderma sp a 25°C y pH 1.5.
Concentración
Colorante
Índigo
(mg
L-1)
qe
Hongo
(mg
g-1)
K2
(g mg-1 h-1)
R2
39
Trichoderma sp. H-1
8.957
0.112
0.964
39
Trichoderma sp. H-2
2.108
0.198
0.985
39
1.291
1.224
0.979
39
1.271
43.361
0.959
Dichos biosólidos presentan mayor capacidad de saturación de la monocapa que las
biomasas secas de las tres cepas de Trichoderma probadas. Aunque es importante
759
mencionar que los estudios se realizaron para el índigo leuco (dihidroindigo), que no fue
empleado en la presente investigación.
Khelifi et al. (2013) reportan la biosorción de azul índigo (100 mg L -1) por la
biomasa fresca muerta de Aspergillus alliaceus, mostrando una capacidad de
saturación de la monocapa de 50 mg g-1 a pH 4 y 30 °C. Mientras que la capacidad de
saturación de la monocapa para la biomasa seca de la cepa Trichoderma sp. H-2 es
ligeramente mayor (57. 271 mg g-1) a la de Aspergillus, sin embargo, estos estudios se
realizaron a pH 1.5, temperatura ambiente y a concentraciones más bajas. Por otro
lado, con lo que respecta a la capacidad de la biomasa seca de algunas especies de
Trichoderma ante otros colorantes, se ha reportado una qm de 10.39 mg g-1 para T.
viride en el caso de cristal violeta y de 21.69 mg g-1 para T. virens ante azul de toluidina,
las cuales son más bajas que en el caso de la biomasa de Trichoderma sp. H-11 ante
azul índigo (Argumedo-Delira et al., 2014).
Conclusiones
La biomasa seca de la cepa Trichoderma sp. H-11 demostró ser un eficiente
biosorbente fúngico ante el colorante azul índigo, por lo que sería pertinente probarse a
más altas concentraciones de este colorante y bajo otras condiciones de temperatura y
pH, que incrementen su remoción. Adicionalmente la remoción de las otras cepas de
Trichoderma podría mejorarse probando otras condiciones de temperatura y pH.
Finalmente un factor a considerar para tener respuestas más homogéneas en la
remoción de azul índigo por la biomasa seca de cepas de Trichoderma, es tener la
misma esporulación.
Agradecimientos
Los autores le agradecen al CONACyT por haber financiado esta investigación (55784)
y al Dr. Jesús Samuel Cruz Sánchez por su inmenso apoyo.
Referencias
Adetuyi, A.O.; Jabar, J.M. (2011). Kinetic and thermodynamic studies of indigo adsorption on some
activated bio-solids. Journal of the Chemical Society of Pakistan. 33: 158-165.
Argumedo-Delira, R.; Alarcón, A.; Ferrera-Cerrato, R.; Mendoza-López, Ma. R.; García-Barradas, O.;
Cruz-Sánchez, J.S. (2014). Estudio comparativo de la adsorción de azul de toluidina y cristal violeta
por la biomasa de dos especies del género Trichoderma y carbón activado. Revista Química Hoy.
4, 17-22.
Balan, D.S.L.; Monteiro, R.T.R. (2001). Decolorization of textile indigo dye by ligninolytic fungi. Journal of
Biotechnology. 89: 141-145.
Cing, S.; Asma, D.; Apoltan, E.; Ye ilada, O. (2003). Decolorization of textile dyeing wastewater by
Phanerochete chrysosporium. Folia Microbiologica. 47: 639-642.
Clark, R.J.H.; Cooksey, C.J.; Daniels, M.A.M.; Withnall, R. (1993). Indigo, woad, and Tyrian Purple:
important vat dyes from antiquity to present. Endeavour, 17: 191-199.
Freundilich, H.M.F. (1906). Over the adsorption in solution. Journal of the American Chemical Society. 57:
385-479.
García, B.E.; Solís-Fuentes, J.A. (2008). ¿Contaminando con sus jeans?. La ciencia y el Hombre de la
Universidad Veracruzana. 21: 1-6.
Ho, Y.S.; Mckay, G. (1999). Pseudo-second order model for sorption processes. Process Biochemistry.
34: 451-465.
760
Khelifi, E.; Touhami, Y.; Bouallagui, H.; Hamdi, M. (2013). Biosorption of indigo from aqueous solution by
dead fungal biomass Aspergillus alliaceus. Desalination and Water Treatment.
DOI:10.1080/19443994.2013.850745.
Langmuir, I. (1916). The constitution and fundamental properties of solids and liquids. Journal of the
American Chemical Society. 38: 2231-2245.
Largergren, S. (1898). About Theory of So-Called Adsorption of Soluble Substances. Kungliga Svenska
Vetenskapsakademiens Handlingar. 24: 1-39.
Manu, B. (2007). Physico-chemical treatment of indigo dye wastewater. Coloration Technology. 123: 197202.
Roessler, A.; Jin, X. (2003). State of the art technologies and new electrochemical methods for
Sanromán, M.A.; Pazos, M.; Ricart, M.T.; Cameselle, C. (2005). Decolourisation of textile indigo dye by
DC electric current. Engineering Geology. 77: 253-261.
Solís, M.; Gil, J.L.; Solís, A.; Pérez, H.I.; Manjarrez, N.; Perdomo, M. (2013). The sedimentation process a
simple method to diminish contaminants in textile effluents. Revista Mexicana de Ingeniería
Química. 12: 585-594.
Solís-Oba, M.; Eloy-Juárez, M.; Teutli, M.; Nava, J.L.; González, I. (2009). Comparison of advanced
techniques for the treatment of an indigo model solution: electro incineration, chemical coagulation
and enzymatic. Revista Mexicana de Ingeniería Química. 8: 275-282.
the reduction of vat dyes. Dyes and Pigments. 59: 223-235.
Unlu, M.; Yukseler, H.; Yetis, U. (2009). Indigo dyeing wastewater reclamation by membrane-based
filtration and coagulation processes. Desalination. 240: 178-185.
Vuorema, A. (2008). Reduction and analysis Methods of indigo. Annales Universitatis Turkuensis. Turku,
Finland. 388: 1-72.
.
Notación
Q
V
Ci
Ce
M
qe
qm
KL
KF
n
qt
t
K1
K2
Cantidad de colorante adsorbido
Volumen de la solución del colorante
Concentración inicial del colorante
Concentración en equilibrio del colorante
Peso seco de la biomasa fúngica
Capacidad de adsorción en el equilibrio
Capacidad máxima de adsorción
Coeficiente de afinidad
Coeficiente de Freundlich
Coeficiente de Freundlich
Cantidad de absorbato absorbido por tiempo
Tiempo
Constante de pseudo primer orden
Constante de pseudo segundo orden
761
CHAPTER 7.8. DECOLORACIÓN DEL COLORANTE AZUL DIRECTO 2 MEDIANTE
PEROXIDASAS
Gabriela Eleonora Moeller-Chávez (1); María Yesenia Vargas- Espinoza (2)
(1) UPEMOR, Jiutepec, Mor. México
(1) UNAM, México DF, México
RESUMEN
Las peroxidasas han surgido como una clase importante de enzimas que se pueden
utilizar para la eficiente degradación de contaminantes orgánicos. En este trabajo se
utilizaron dos métodos enzimáticos para el tratamiento del colorante azul directo 2
(AD2). Se evaluó la capacidad catalítica de dos enzimas peroxidasas la primera,
obtenida de un residuo industrial de soya y la segunda proveniente del fruto del
chayote. El objetivo de este trabajo es evaluar la capacidad de decoloración de las
peroxidasas sobre el colorante azul directo 2 (AD2) preparado en solución sintética. Se
evaluaron varios parámetros operacionales que afectan a la degradación enzimática
tales como concentración inicial del colorante, dosis de peróxido de hidrógeno, tiempo
de contacto, ph de la solución y cantidad de enzima peroxidasa. En las pruebas
realizadas la concentración del colorante (AD2) utilizada fue de 10, 20, 30, 40, 50, y 60
mg l-1. La estrategia de dosificación de la enzima fue una única adición al inicio del
tratamiento.
Palabras claves: colorantes azo, degradación, enzima peroxidasa
ABSTRACT
Peroxidases are emerging as an important kind of enzymes that can be used for the
efficient degradation of organic contaminants. In this study, two enzymatic methods for
the direct treatment of direct blue 2 (AD2) were used. The catalytic ability of two
peroxidases, obtained from an industrial waste soybean and from the fruit of the chayote
was evaluated. The objective of this work is to evaluate the ability of peroxidases to
decolorize the direct blue 2 (AD2) dye, prepared in a synthetic solution. Several
operational parameters that affect the enzymatic degradation such as initial
concentration of the dye, hydrogen peroxide dose, contact time, ph of solution and
amount of peroxidase enzyme were evaluated. During the experiments, the
concentration of the dye (AD2) used was 10, 20, 30, 40, 50, and 60 mg l-1. The strategy
for the addition of the enzyme dosage was a unique addition al the starting time (t= 0).
Key words: azo dye, degradation, enzyme peroxidase.
Introducción
762
Los colorantes y pigmentos son sustancias orgánicas que se utilizan en pequeñas
proporciones para impartir color a diferentes materiales con propósitos funcionales o
estéticos. Los colorantes son sustancias orgánicas que exhiben color debido a su
capacidad para absorber luz de la región visible del espectro desde longitudes de onda
de 4000 a 8000 Å (Garzón-Zúñiga et al., 2011).
Los colorantes azo, representan el 70% de la totalidad de colorantes utilizados en
el mundo (Tantak y Chaudhari., 2006). Los colorantes azo son compuestos químicos
xenobióticos, que se caracterizan por la presencia en su estructura molecular del grupo
cromóforo azo, constituido por dos átomos de nitrógeno unidos mediante un doble
enlace (-N=N-) (Muruganandham y Swaminathan., 2004).
En la industria textil se consumen dos tercios de la producción de colorantes azo,
la cual fue estimada en 1, 400,000 toneladas en el año 2006 (Mahmoodi y Arami, 2006),
materia prima que es utilizada para el proceso de teñido. Se estima que el 15% de la
producción mundial total de colorantes se pierde durante el proceso de teñido, el cual
se desecha en los efluentes (Maljaei, et al., 2009).
Objetivo
Evaluar la capacidad oxidante de las enzimas peroxidasas obtenidas de: a) de un
residuo industrial del procesamiento del frijol de soya b) del chayote en la decoloración
del colorante azul directo 2 (AD2).
Colorante utilizado. Se evaluó la degradación del colorante azul directo 2 (AD2). Este
colorante es clasificado como un colorante diazo debido a la presencia de sus dobles
enlaces nitrogenados (-N=N) (Figura 1). Es ampliamente utilizado en la industria textil.
Fue Suministrado por una empresa ubicada en CIVAC Morelos. Este compuesto se ha
seleccionado ya que es uno de los colorantes más usados en la industria textil
mexicana para el teñido de la mezclilla. Las propiedades químicas del colorante se
presentan en la (Tabla 1). La preparación del colorante fue como se recibió sin ningún
tratamiento adicional.
FIGURA 7. Estructura química del colorante azul directo 2. Fuente: Golka et al.,
2004; Morgan et al., 1994.
763
TABLA 1. Propiedades químicas del colorante azul directo 2 (AD2)
Propiedades
Valor
Nombre comercial
Chlorazol Black BH, Azul Directo 2
Formula molecular
C32H21N6Na3O11S3
Peso molecular
830.71 g/mol
Solubilidad en agua
45 g/L (25°C)
Colour Index
22590
No. Cas
2429-73-4
Las fuentes de peroxidasas. Las peroxidasas han surgido como una clase importante
de enzimas que se pueden utilizar para la eficiente degradación de contaminantes
orgánicos. En este trabajo se utilizaron dos métodos enzimáticos para el tratamiento
del colorante azul directo 2 (AD2). Se evaluó la capacidad catalítica de dos enzimas
peroxidasas la primera, obtenida de un residuo industrial de soya y la segunda
proveniente del fruto del chayote.
Actividad enzimática. En la determinación de la actividad enzimática se utilizó el
mismo método para las dos fuentes de peroxidasa, en donde una unidad de actividad
enzimática (AE) se define como la cantidad de enzima que produce 1 mmol en 1
minuto, referido a 1 mL de preparado enzimático o 1 mg de proteína (actividad
específica) (Marquéz, 1999). La (AE) también se define como la cantidad de enzima
que convierte 1 mol de peróxido de hidrógeno por minuto a pH 7.4 y 25 º C (Pramparo,
2008). Para la determinación de la actividad enzimática se utilizó el método de la 4aminoantipirina, la cual consiste en una valoración colorimétrica, utilizando fenol 0.2 M y
H2O2 2x10-2 M como substratos y 4-aminoantipirina 4.8 x 10-2 M como cromógeno y
una solución de buffer de fosfatos a pH 7.4. La determinación se realizó mediante un
espectrofotómetro UV-Vis a una longitud de onda de 510 nm. Para determinar la
actividad enzimática por volumen o total se utilizó la siguiente ecuación.
Actividad volumétrica: (A2 ± A1)/(t2 - t19H‫ڙ‬G9P
Donde:
A2: Absorbancia final
A1: Absorbancia inicial.
t2: Tiempo final
t1: Tiempo inicial
Ve: Volumen de ensayo
‫&ڙ‬RQVWDQWHPRODUGHGHVDSDULFLyQGHOD-aminoantipirina.
D: Distancia que atraviesa el haz de luz por la celda.
Vm: Volumen de la muestra
Ensayos de decoloración. El diseño experimental utilizado para los ensayos de
decoloración fue un diseño factorial de cribado 2^6-1 mediante el programa
Statgraphics para optimizar el proceso. Se realizaron 70 pruebas experimentales
incluyendo 6 puntos centrales, con 6 números de factores experimentales temperatura,
764
pH, tiempo de contacto, dosis de enzima, dosis de peróxido de hidrógeno, y
concentración del colorante. Las pruebas fueron totalmente aleatorias. Utilizando como
factor de respuesta el porcentaje (%) de decoloración los factores experimentales se
presentan en la (Tabla 2).
TABLA 2. Factores de operación para los ensayos de decoloración para el
colorante azul directo 2 (AD2).
Factores
Bajo
Alto
Unidades
Continuo
Temperatura
pH
Tiempo de contacto
25.0
3.5
2.0
50.0
7.5
6.0
°C
Unidades
Horas
Si
Si
Si
Dosis de enzima
1000.0
2000.0
microlitros
Si
Dosis de peróxido
1000.0
2000.0
microlitros
Si
Colorante
10.0
60.0
mg/L
Si
Cálculo para el porcentaje de decoloración (D%). La eliminación del color se controló
por análisis de absorbancia de las muestras retiradas utilizando un espectrofotómetro
de barrido (UV / visible), contra una línea de base definida por la absorbancia. La
absorbancia de cada sobrenadante clarificado se midió a la longitud de onda de
absorción máxima (Ȝmax) del colorante (576 nm). La concentración de colorante
residual en las muestras se determinó a partir de las lecturas de absorbancia utilizando
la curva de calibración a diferentes concentraciones de colorante. El % de decoloración
(D%) se calculó usando la siguiente ecuación:
ࡰΨ ൌ
۱ܑି۱‫ܜ‬
۱ܑ
X 100
Donde Ci es la concentración inicial de la muestra y Ct es la concentración de la
muestra a un determinado tiempo de reacción. Para llevar a cabo la determinación de
las concentraciones en cada una de las muestras previamente se realizó una curva de
calibración.
Resultados
Actividad enzimática. Una unidad de actividad enzimática (símbolo U) es la cantidad
de enzima que en una reacción enzimática cataliza la conversión de 1 µmol de sustrato
por minuto. Se utiliza también en combinación con otras unidades (U/mg de proteína o
U/mL) para señalar, respectivamente, la actividad enzimática específica o la
concentración de actividades, se obtuvieron actividades enzimáticas de 20 a 30 UL, la
actividad enzimática se realizó por el método de la 4 amino antipirina que es un método
colorimétrico. Se realizaron 27 cromatografías dando un volumen total de 500 mL con
actividad enzimática 20 a 30 U/L en promedio. Como se describe en la (Figura 2).
765
FIGURA 8. Actividad peroxidasa del residuo industrial de soya y del fruto del
chayote.
Diseño experimental. Después de los ensayos realizados se obtuvieron los siguientes
resultados con las posibles combinaciones de los factores evaluados mencionados en
la (Tabla 2).
Diagrama de Pareto. De acuerdo a la gráfica de Pareto estandarizado, la longitud de
cada barra es proporcional al valor del estadístico-t calculado para el efecto
correspondiente. Cualquier barra más allá de la línea vertical es estadísticamente
significativa en el nivel de significancia, establecido en 5%. De acuerdo al análisis, solo
uno fue el factor que causo efecto significativo en el tratamiento, el pH (Figura 3), para
la peroxidasa del chayote fueron dos los factores significativos el efecto del pH, y el
tiempo de contacto son los factores que afectan el porcentaje de decoloración del
colorante azul directo 2 (AD2) (Figura 4).
Diagrama de Pareto Estandarizada para % de decoloración
Diagrama de Pareto Estandarizada para Remociòn de color
B:pH
+
-
A:Temperatura
B:pH
+
-
C:Tiempo de contacto
E:Dosis de peróxido
D:Dosis de enzima
A:Temperatura
F:Colorante
D:Dosis de enzima
C:Tiempo de contacto
F:Colorante
E:Dosis de peroxido
0
2
4
Efecto estandarizado
6
8
0
2
4
Efecto estandarizado
6
8
FIGURA
9.
Diagrama
de
Pareto FIGURA
10.
Diagrama
de
Pareto
estandarizado para la decoloración (AD2) estandarizado para la decoloración (AD2)
para el residuo industrial de soya.
para el chayote.
766
Tiempo de reacción. La eficiencia de decoloración del azul directo 2 (AD2) está en
función del tiempo de contacto con la enzima, como se observa en la (Figura 5). De
acuerdo al diseño estadístico el tiempo óptimo para alcanzar la mayor eficiencia de
decoloración para la peroxidasa del residuo industrial de soya es de 2 horas alcanzando
una eficiencia de 71%. Sin embargo para la peroxidasa del chayote de acuerdo al
diseño estadístico el tiempo de contacto es de 6 horas logrando porcentajes de
decoloración hasta en un 97% (Figura 6).
80
60
40
20
0
2
% de decoloración
Remociòn de color
Superficie de Respuesta Estimada
Temperatura=37.5,pH=5.5,Dosis de peroxido=1500.0,Colorante=35.0
Remociòn de color
43.0
43.8
44.6
45.4
46.2
47.0
47.8
(X 1000.0)
48.6
2
49.4
1.8
1.6
50.2
1.4
1.2
3
51.0
1
4
5
6
Dosis de enzima
51.8
Tiempo de contacto
100
80
60
40
20
0
2
Temperatura=37.5,pH=5.5,Dosis de peróxido=1500.0,Colorante=35.0
% de decoloración
52.0
56.0
60.0
64.0
68.0
72.0
76.0
(X 1000.0)
80.0
2
84.0
1.8
1.6
88.0
1.4
1.2
3
92.0
1
4
5
6
Dosis de enzima
96.0
Tiempo de contacto
FIGURA 11. Optimización del tiempo de FIGURA 12. Optimización del tiempo de
reacción para la decoloración (AD2) reacción para la decoloración (AD2)
para el residuo industrial de soya .
para el chayote.
Concentración de H2O2. Habitualmente, las reacciones de oxidación con el peróxido de
hidrógeno, se producen rápidamente en medio básico y de forma muy lenta si el pH es
ácido. Pero se debe hacer notar que, en presencia de peroxidasas, el peróxido de
hidrógeno puede actuar como oxidante de diferentes sustratos independientemente del
pH imperante en el medio. El peróxido de hidrógeno actúa como un co-sustrato para
activar la acción enzimática de la peroxidasa. Esto contribuye en el ciclo catalítico de las
peroxidasas, donde inicialmente se oxida la enzima nativa para formar un intermediario
enzimático el cual acepta los compuestos aromáticos para llevar a cabo su oxidación
llegando a la formación de un radical libre. Por tal situación la dosis de peróxido juega
un papel muy importante en las reacciones enzimáticas ya que al tener un exceso de
este reactivo en la reacción puede llegar a causar inhibición en la actividad enzimática y
cuando está presente en cantidades pequeñas puede ser un limitante en la velocidad
de reacción muy probablemente debido a daños en la propia enzima. Esto muestra la
importancia crítica de la optimización de las concentraciones de H 2O2 en estos
enfoques de degradación de colorante a base de enzimas. En la (Figura 7) se puede
observar que 2000 µL es la dosis de H2O2 en donde se obtiene la mayor actividad
enzimática logrando decolorar un 71%. Concentraciones menores de 1000 µL
presentan una disminución en la velocidad de reacción. Por los resultados obtenidos, la
concentración de H2O2 no presenta un efecto significativo sobre la reacción catalizada
por la enzima. En la (Figura 8) se puede observar que 1000 µL es la dosis de H 2O2 en
donde se obtiene la mayor actividad enzimática logrando decolorar por arriba del 97%.
Concentraciones menores de 1000 µL presentan una disminución en la velocidad de
reacción.
767
Remociòn de color
80
60
40
20
0
1
1.2
1.4
1.6
Dosis de enzima
1.8
Remociòn de color
43.0
43.8
44.6
45.4
46.2
47.0
47.8
(X 1000.0)
48.6
2
49.4
1.8
1.6
50.2
1.4
1.2
51.0
1
Dosis de peroxido
2
51.8
(X 1000.0)
% de decoloración
Temperatura=37.5,pH=5.5,Tiempo de contacto=4.0,Colorante=35.0
100
80
60
40
20
0
1
Temperatura=37.5,pH=3.5,Tiempo de contacto=4.0,Colorante=35.0
% de decoloración
52.0
56.0
60.0
64.0
68.0
72.0
76.0
(X 1000.0)
80.0
2
84.0
1.8
1.6
88.0
1.4
1.2
1.2
92.0
1.4
1
1.6
1.8
2
96.0
Dosis de peróxido
(X 1000.0)
Dosis de enzima
FIGURA 7. Optimización dosis de FIGURA 13. Optimización dosis de
enzima y peróxido para la decoloración enzima
y
peróxido
para
la
(AD2) para el residuo industrial de soya. decoloración (AD2) para el chayote.
Temperatura y pH. La eficiencia de decoloración del azul directo 2 (AD2) para la
enzima peroxidasa del residuo industrial de soya está en función del pH y la
temperatura el diseño estadístico mostró que el proceso de decoloración es
estrictamente dependiente del pH de la solución acuosa del colorante. El pH para
obtener la mayor decoloración del (AD2) fue de 3.5 logrando un porcentaje de
decoloración del 71%. De acuerdo al diseño estadístico la temperatura óptima para
alcanzar porcentajes de decoloración del 43 al 51% fue de 25°C, como se muestra en la
(Figura 9). Para la peroxidasa del chayote de acuerdo al diseño estadístico la
temperatura de 50 ° C causa un aumento en la decoloración del 50 al 96 %, con un pH
de 3.5 como se muestra en la (Figura 10).
Remociòn de color
Tiempo de contacto=4.0,Dosis de enzima=1500.0,Dosis de peroxido=1500.0,Colorante
Remociòn de color
43.0
43.8
80
44.6
45.4
60
46.2
47.0
40
47.8
48.6
20
7.5
49.4
6.5
0
5.5
50.2
4.5
25
30
51.0
35
3.5
40
pH
45
50
51.8
Temperatura
% de decoloración
Tiempo de contacto=4.0,Dosis de enzima=1500.0,Dosis de peróxido=1500.0,Colorante
% de decoloración
52.0
56.0
100
60.0
64.0
80
68.0
60
72.0
76.0
40
80.0
20
7.5
84.0
6.5
0
5.5
88.0
4.5
25
30
92.0
35
3.5
40
pH
45
50
96.0
Temperatura
FIGURA 14. Optimización de temperatura FIGURA
15.
Optimización
de
y pH para la decoloración (AD2) para el temperatura y pH para la decoloración
residuo industrial de soya.
(AD2) para el chayote.
Concentración de colorante. La concentración del sustrato presente en la fase acuosa
es un factor clave el cual afecta la velocidad de oxidación catalizada por una enzima.
768
Los resultados se muestran en la (Figura 11) donde se puede observar que el mayor
porcentaje de decoloración se obtiene a concentraciones mayores de 40 mg/L logrando
decolorar un 51% para la peroxidasa del residuo industrial de soya. Mientras que para
la peroxidasa del chayote (Figura 12) donde se observa que a concentraciones de
colorante mayores del 50% se pueden logra eficiencia del 96% de decoloración, ya que
a concentraciones menores de 50% disminuye el porcentaje de decoloración.
Contornos de la Superficie de Respuesta Estimada
Contornos de la Superficie de Respuesta Estimada
Temperatura=37.5,pH=5.5,Tiempo de contacto=4.0,Dosis de enzima=1500.0
Colorante
50
40
30
20
10
0
1
1.2
1.4
1.6
Dosis de peroxido
1.8
Remociòn de color
43.0
43.8
44.6
45.4
46.2
47.0
47.8
48.6
49.4
50.2
51.0
51.8
2
(X 1000.0)
60
50
Colorante
60
Temperatura=37.5,pH=3.5,Tiempo de contacto=4.0,Dosis de enzima=1500.0
40
30
20
10
0
1
1.2
1.4
1.6
Dosis de peróxido
1.8
% de decoloración
52.0
56.0
60.0
64.0
68.0
72.0
76.0
80.0
84.0
88.0
92.0
96.0
2
(X 1000.0)
FIGURA 16. Optimización de dosis de FIGURA 17. Optimización de dosis de
enzima y peróxido para la decoloración enzima
y
peróxido
para
la
(AD2) para el residuo industrial de soya.
decoloración (AD2) para el chayote.
Seguimiento visual de los ensayos experimentales. En las Figuras (13 y 14) se
muestra el seguimiento visual de la decoloración del colorante azul directo 2 (AD2).
FIGURA 18. Seguimiento visual
concentración inicial de 10 mg/L.
para
la
decoloración
(AD2),
con
una
769
FIGURA 19. Seguimiento visual
concentración inicial de 60 mg/L.
para
la
decoloración
(AD2),
con
una
Conclusiones
De los resultados obtenidos se concluye que:
Con la peroxidasa de chayote se lograron altos porcentajes de remoción del colorante
azul directo 2 (AD2) hasta de un 97%. Para la enzima del residuo industrial de soya se
obtuvieron porcentajes del 71%.
Se logró optimizar el proceso de decoloración del (AD2), resultando 5 factores
influyentes en el tratamiento: temperatura, pH, tiempo de contacto, dosis de enzima y
concentración del colorante.
Por lo anterior se concluye que las peroxidasas de las dos fuentes son una alternativa
de remoción para el colorante azul directo 2 (AD2).
Referencias
GARZÓN-ZÚÑIGA, M.A., SANDOVAL-VILLASANA, A.M. AND MOELLER-CHÁVEZ, G.E. (2011).
Decolorization of the AO24 azo dye and reduction of toxicity and genotoxicity in trickling biofilters.
Water Environment Reseach 83 (2), 107-115.
GOLKA, K., KOPPS, S., MYSLAK, Z.W. (2004) Carcinogenicity of azo colorants: influence of.
MAHMOODI, N.M. AND ARAMI M. (2006) Bulk phase degradation of acid red 14 by nanophotocatalysis
using immobilized titanium (IV) oxide nanoparticles. Journal of Photochemistry and Photobiology A:
Chemistry 182, 60-66.
MALJAEI, A., ARAMI, M. AND MOHAMMAD, N. M. (2009) Decolorization and aromatic ring degradation
of colored textile wastewater using indirect electrochemical oxidation method. Desalination 249 (3),
1074 - 1078.
MARQUÉZ, M. Y M. VÁZQUEZ (1999). Modeling of enzymatic protein hydrolysis, Process Biochemistry:
35(1), 111±117.
MORGAN, D.L., DUNNICK, J.K., GOEHL, T., JOKINEN, M.P., MATTHEWS, H.B., ZEIGER, E.,
MENNEAR, J.H. (1994). Summary of the National Toxicology Program Benzidine Dye Initiative.
Environmental Health Perspectives. 102, Supplement 2, 63-78. http://www.jstor.org/stable/3431822.
MURUGANANDHAM, M. AND SWAMINATHAN, M. (2004). Decolourisation of reactive orange 4 by
fenton´s and photo-fenton oxidation technology. Dyes and Pigments 63 (3), 315-321.
PRAMPARO, L.; PRUVOST, J.; STÜBER, F.; FONT, J.; FORTUNY, A.; FABREGAT, A.;
LEGENTILHOMME, P.; LEGRAND, J.; BENGOA, (2008). C. Numerical approach for the enzymatic
elimination of phenol in a torus reactor. Poster, 18th European Symposium on Computer Aided
Process Engineering (ESCAPE-18), Lyon (France).
TANTAK, N.P. AND CHAUDHARI, S. (2006) Degradation of azo dyes by sequential fenton´s oxidation
and aerobic biological treatment. Journal of Hazardous Materials B136, 698-705.
770
CHAPTER 7.9. INCREASED BIODEGRADABILITY OF THE FUNGICIDE IMAZALIL
AFTER PHOTO-FENTON TREATMENT
Roberto Candal (1, 2); Federico Ariganello (1); Gustavo Curutchet (1, 2);
Elsa López-Loveira (1)*
(1) Instituto de Investigación e Ingeniería Ambiental, Universidad Nacional de San Martín, Campus
Miguelete 25 de Mayo y Francia, Provincia de Buenos Aires, Argentina.
(2) Consejo Nacional de Investigaciones Científicas y Técnicas, Rivadavia 1917, Buenos Aires, Argentina.
ABSTRACT
Imazalil (C14H14Cl2N2O) is a widely used fungicide for the post-harvest treatment of citric
fruits. This fungicide is usually commercialized in the form of emulsions containing 50%
of the active compound and coadjutants that help to stabilize the emulsion and disperse
the product. The presence of Imazalil and the coadjutants in waste waters, produced by
washing the fruits before packing, is a potential risk for the living organisms present in
the receptor water courses. The treatment of the waste waters containing pesticide is
mandatory to preserve environment quality. Imazalil is not easily biodegradable;
however, a resistant microbial consortium was isolated from sludge containing the
fungicide that was not able to degrade the pollutant in culture with planktonic biomass.
Photo-Fenton is an advanced oxidation process that was successfully used to degrade
pesticides in water using light, Fe(III) and H2O2, as energy source, catalyst and oxidant
respectively. Frequently, it is not possible to completely mineralize the pollutant and only
a partial reduction of total organic carbon (TOC) can be accomplished. However, the
partial oxidation of the recalcitrant compound may lead to byproducts with a higher
biodegradability than the former pollutant.
In this work, water solutions containing 500 mg/L Imazalil and 500 mg/L TOC were
partially oxidized by photo-Fenton process (FeCl3.6H2O (0,15 mM), pH = 3.0, 25 ºC and
H2O2 doses. The reaction was carried out in a batch reactor (250 mL) with top lighting
(UVA, 20 W). After 4 h illumination, Imazalil was completely degraded. There was no
change in the TOC with 9 mM of H2O2, which means that mineralization was negligible
under these conditions, but with 3 doses of 9 mM and 27 mM of H 2O2, TOC mildly
decreased and with 54 mM of H2O2 TOC decreased considerably. The four remaining
solutions were adjusted for biological treatment. The pH was raised to 7.0 and the
solution was supplemented with inorganic nutrients (SO4(NH4)2, K2HPO4, MgSO4 and
CaCl2). The solutions (100 mL) were inoculated with a consortium previously adapted to
Imazalil, and incubated at 25 ºC. After 5 days of incubation, the TOC decreased by 0%,
42%, 59%, 62% and 84% in solutions without photo-Fenton treatment and with 9, 3 x 9,
27 and 54 mM of H2O2 respectively, indicating that the oxidation treatment produces
byproducts with higher biodegradability than the former fungicide. This coupled process
shows to have great potential for treatment of Imazalil for waste water from fruit industry.
771
Keywords: biodegradation, coupled treatment, Imazalil, photo-Fenton, water treatment
Introduction
Imazalil (IMZ) is a postharvest fungicide widely for the control of fungi that use to attack
several vegetables and fruits, in particular citrus. IMZ belong to the family of the
imidazole, it is moderately soluble in water a persistent in soils. The United States
Environmental Protection Agency (US-EPA) had classified this fungicide as
carcinogenic; it is a nice example of an agrochemical that should be eliminated from
liquid effluents before its discharge in water bodies (EPA, 1999; Santa Cruz
Biotechnology, 2010).
Aqueous effluents containing pesticides can not be purified by conventional
biological treatment due to its toxicity towards the microorganism involved in the process
and/or its low biodegradability (Dunia, 2013). Specific treatment methods or a
combination of different methods should be use to eliminate these products from water
effluents.
Advanced oxidation technologies (AOTs) are considered one of the most promising
technologies for the elimination of recalcitrant pollutants present in water effluents.
Recalcitrant pollutants are characterized by high chemical stability and low
biodegradability (Malato, 2009; Soon, 2011; Ballesteros, 2009; Hincapie-Perez, 2006).
AOTs involve the generation of the powerful oxidants hydroxyl radicals (HO x),
which act as a no specific oxidant leading to mineralization, or partial oxidation, of all the
organic compounds dissolved in water. These radicals can be generated by different
processes as TiO2-photocatalysis, UVC-H2O2, Fenton process, photo-Fenton process
and other related processes.
Fenton process is based in the catalytic decomposition of H2O2 in the presence of
Fe(II). As consequence of the reaction is produced Fe(III), which can be reduced by
H2O2 to Fe(II), entering the catalytic cycle. Fenton process can be described by the
following set of reactions:
Fe(II) + H2O2
Fe(III) + H2O2
Fe(III)-OOH2+
Fe(III) + HOx + HOFe(III)-OOH2+ + H+
Fe(II) + HO2x
(1)
(2)
(3)
The rate of the process is limited by reaction (3) which lead to the regeneration of
Fe(II), necessary for the continuity of the cycle. However, Fe(II) production can be
quickly produced by the photodecomposition of Fe(HO)2+,which is produced by the
hydrolysis of Fe(III) in water:
Fe(III) + H2O
Fe(OH)2+ + hQ
Fe(OH)2+ + H+
Fe(II) + HOx
(4)
(5)
This process is called photo-Fenton and is produced under illumination with light in
the range 365-420 nm (Perez, 2002; Pignatello, 1999).
772
Although AOTs have great potential for water decontamination, if they were applied
to reduce the Chemical Oxygen Demand (COD) and/or the BOD to acceptable levels the
consumption of the needed chemicals may be too high and the treatment results
unaffordable. The combination of AOTs with biological treatment may help to reduce
costs.
AOTs can be coupled with biological processes as a pretreatment or a post
treatment. The first case is applicable when the effluent contains important amounts of
recalcitrant pollutants and a low charge of biodegradable organic matter. In this way it
may be possible to decrease the concentration of recalcitrant compounds by oxidation
with HOx radicals. The oxidized species usually display higher biodegradability and less
toxicity than the mother compounds. The treated effluent can then be directed to a
regular biological plant. The low organic charge is necessary to prevent the consumption
of HOx radicals by reaction with other biodegradable substances (as sugars, proteins,
lipids, etc). The second configuration (AOTs as a post treatment) can be used when the
organic load is high and the recalcitrant compounds are not too toxic for the
microorganism present in the biological plant. In this way the effluent firstly enters to the
biological plant, where most of the biodegradable organic load is degraded and the
Biological Oxygen Demand (BOD) is decreased to acceptable levels. Secondly, the
HIIOXHQWLV³UHILQHGÉ\DQ$27WUHDWPHQWZKHUHWKHUHPDLQLQJUHFDOFLWUDQWLVFRPSOHWHO\
eliminated. The treated effluent can be discharged to the receptor water body or can be
reuse for industrial or agricultural processes, depending on its final composition.
The key to appropriately combine chemical with bio-oxidation is to find the
adequate chemicals concentration and necessary chemical treatment time to obtain an
effluent containing substances bio-degradable enough to be mostly eliminated during
bio-treatment. The optimum time is related with the biodegradability and toxicity of the
byproducts generated during chemical oxidation. Long treatment times consume high
quantities of H2O2 producing highly oxidized byproducts, these oxidized compounds
possess low yield for the formation of biomass (biomass obtained / consumed carbon), a
fact that can lead to the need of introducing additional carbon source in the system to
support good growth. However, the lower biomass generation involves less sludge
formation in the process of removing organic carbon that favoring the cost reducing of
sludge disposal. By the other hand, short oxidation times produce byproducts with
molecules quite similar to the pesticide, which are also toxic and recalcitrant
(Ballesteros, 2009; Lapertot, 2007; Zapata, 2010; Oller, 2007)
In this work a combine photo-Fenton ± biological process for the depuration of
water containing IMZ is studied. The H2O2 dosification was optimized to obtain the
higher mineralization ratio.
Chemicals. In most of the experiments the commercial fungicide Xedrel 50® (IMZ
50g/100ml, Magan) was used as source of IMZ (Table 1 shows several characteristics of
Xedrel). For analytical purposes pure IMZ (PESTANAL® analytical grade, SigmaAldrich) was used. Hydrogen peroxide 30%, ferric chloride, acetonitrile, sodium sulfite, sodium hydroxide, sulfuric acid, were also used; all of them of analytical grade.
773
Analytical determinations. The IMZ concentration was determined by HPLC with UV
detector, using modular Shimadzu equipment coupled to a Shim-Pack VP-ODS, reverse
phase column (250 mm long, 4.6 µm particle size and 12 nm pore size). A mix
acetonitrile:water (70:30) was used as eluent (flow: 1 ml min-1). The Total Organic
Carbon (TOC) was determined using a TOC-L Shimadzu. The H2O2 concentration was
colorimetrically determined (Pupo-Nogueira, 2005). The Fe concentration was measured
by atomic absorption spectroscopy.
TABLE 1. Physico-chemical and ecological properties of the Imazalil
Pesticide
Active
Ingredient
% Xedrel 50®
Imazalil
C14H14Cl2N2O
Emulsifiable
Magan
50
concentrate
Commercial
Molecular Formula
Formulated
Solubility
TOC
in water
%
High in acid pH
57.8
Ecological Information
Very toxic to aquatic
organisms
Culture medium and condition. Appropriate culture media were used to grow the
bacteria consortium in the reactors. The liquid medium was composed of 5.0 g/L
glucose, 2.0 g/L (NH4)2SO4, 1.0 g/L K2HPO4, 0.1 g/L MgSO4, 0.01 g/L CaCl2 and
different concentrations of IMZ from Xedrel and pH 7.0. They were incubated in agitated
reactors at constant temperature (25°C).
7KH FRPSRVLWLRQ RI WKH ³EDVDO FXOWXUH PHGLXP´ ZDV VLPLODU WKDQ WKH ULFK PHGLXP EXW
without organic carbon.
Experimental set-up. IMZ-resistant microorganisms isolation. Sludge coming from a
fruit packaging industry waste water treatment plant was exposed to a selection process
by incubation in a rich culture medium with increasing concentration of Xedrel 50®. In
the first step 2.5% of sludge was incubated in rich medium with 100 ppm of IMZ (from
Xedrel) at neutral pH, 25 ºC for 3 days. In a second step, 1% of the enriched consortium
was subculture in rich media with increasing IMZ concentration, up to 500 ppm (solubility
in water). The consortium was maintained by periodical subculturing in rich medium with
500 ppm of IMZ.
Photo-Fenton treatment. Water solutions containing 500 mg/L of IMZ and 500 mg/L of
TOC from Xedrel 50® were partially oxidized by photo-Fenton process incorporating
FeCl3.6H2O (0,15 mM), and H2O2 in different doses (9 mM, 3 x 9 mM, 27 mM y 54 mM);
the temperature was maintained 25ºC constant all along the experiments, the pH was
adjusted at 3.0 using H2SO4. The reaction was carried out in batch reactors of 250 mL
with top lighting (UVA, 20 W), and magnetically stirred. The reaction kinetics was studied
by taking aliquots at several times, determining for each one the: pesticide
concentration, TOC, H2O2 and pH. Sodium sulfite or acetonitrile were used to quench
the reaction in the samples. The photo-Fenton treatment was also extended until
complete consumption of H2O2 so that it could be coupled to biotreatment. H2O2 may
damage the bacterial cells. Solutions obtained were supplemented with inorganic
nutrients (SO4(NH4)2, K2HPO4, MgSO4 and CaCl2) until reach the basal medium
concentrations and were used in biodegradation assays as described below.
774
Biodegradation assays. Biological assays with planktonic cells were carried out in
batch mode, in 100 mL shaken flasks at 25°C and initial pH 7.0 using basal culture
medium. Two concentrations of: IMZ in Xedrel (50 y 500 mg/L), with and without of a
suplementary carbon source (glucose 5,0; 2,5; 1,25 y 0,0 g/L) were studied. For the
previously photocatalyzed solutions of IMZ in Xedrel, pH was adjusted at 7.0 with NaOH
and controlled all throughout and no external carbon source was added. The following
parameters were measured at several times: pH, TOC and IMZ.
Imazalil biodegradation assays. Consortium incubated with Xedrel 50® as the only
source of carbon did not show cell growing or carbon consumption, showing that the
active principle as well as additives from Xedrel 50® cannot be biodegraded. By the
other hand in all the cultures added with glucose as carbon source, cell growing and
carbon diminution was observed. In the samples with 1.25 and 2.5 g/L glucose all the
glucose was consumed, being the obtained biomass proportional to the initial glucose
mass. However, in the samples with 5.0 g/L glucose the growing stopped before all the
glucose was consumed. This phenomenon is consequence of the diminution of pH due
to ammonium assimilation as nitrogen source. In the samples with glucose 5.0 g/L, the
final pH was as low as 3.0. Under this pH conditions, the cell growing stopped as
demonstrated by further experiments run with 5.0 g/L glucose at pH 3.0. In this case no
cell growing was observed, probing that under this pH conditions the consortium do not
grow.
In all the cases, the concentration of IMZ remains constant, meaning that no
degradation neither adsorption of IMZ took placed during cell growing, being glucose the
only carbon source used by the consortium.
Photo-Fenton process. Figure 1 shows the temporal evolution of IMZ and H2O2 during
photo-Fenton treatment of solutions containing different initial concentration of H 2O2. In
all the cases IMZ was completely degraded in approximately 4 h. As regard to H 2O2
consumption, two phases can be identified: a slow one that extends to the total
recalcitrant compound degradation followed by a fast H2O2 consumption where the
oxidant is completely degraded. TOC was measured in all the cases immediately after
H2O2 was completely consumed. In the system with 9 mM H2O2 (System B) there was
no change in the TOC, which means that mineralization was negligible under these
conditions. However in the systems with 27 (System D) and 54 (System E) mM H 2O2,
TOC decreased in 3 and 40 % respectively. When H2O2 was incorporated as three
doses of 9 mM H2O2 (each one added after total consumption of H2O2) (System C), TOC
decreased was 11%. Incorporation of H2O2 in doses resulted more efficient towards
mineralization than the incorporation of the same final quantity of H2O2 in a single dose.
In all the cases final pH was in the range 2.7-2.2; the diminution in pH may be
associated with the production of organic acids as oxidation byproducts (Hazime, 2012).
These results suggest a complex degradation mechanism, where the different
produced byproducts display different degradation degree and H2O2 consumption rates.
Complexation of Fe(III) by the oxidized byproducts cannot be disregarded. A detailed
study of the reaction mechanism is out of the scope of this work.
775
H2O2 9 mM
H2O2 27 mM
H2O2 54 mM
IMZ (9mM)
IMZ (27 mM)
IMZ (54 mM)
40
20
0
600
400
200
0
2
4
6
8
IMZ (mg/L)
H2O2 (mM)
60
0
Time (h)
FIGURE 1: Evolution of IMZ and H2O2 concentration in the solution when exposed
to the photo-Fenton treatment.
Coupled photo-Fenton ± Biological treatment. Based on the previous results, in order
to explore the effect of photo-Fenton treatment on biodegradability of this pesticide,
Xedrel 50® water solutions were photo-treated until all IMZ disappeared. The remaining
solutions were submitted to biological treatment in batch reactors containing the
previously isolates consortium. In this way both, photo-Fenton and biological treatments
were coupled. One control and four samples were exposed to biological treatment: A:
without photo-Fenton treatment, B, C, D and E: samples partially oxidized by photoFenton process in the presence of 9 mM H2O2, 3 doses of 9 mM H2O2, D: 27 mM H2O2
and E: 54 mM H2O2 respectively. The five solutions were adjusted for biological
treatment: pH was raised to 7.0 and the solutions were supplemented with inorganic
nutrients (SO4(NH4)2, K2HPO4, MgSO4 and CaCl2). After neutralization, the
concentration of soluble iron measured in the solution was less than 1.0 mg/L. The
excess of iron precipitated as a dark-orange powder. The solutions (100 mL) were
inoculated with the consortium adapted to IMZ (and to the degradation products after
photo-Fenton treatment), and incubated at 25 ºC. The adaptation of the consortium to
IMZ and it´s oxidation byproducts was done by successive incubations in basal medium
containing these compounds.
776
A
B
C
D
100
E
a
600
400
200
0
0
5
10
15
20
Degradation Ratio (%)
TOC (mgC/L)
800
b
Photo- Fenton Treatment
Biological Treatment
80
60
40
20
0
A
B
C
D
E
Time (days)
FIGURE 2a: Evolution of TOC during the biological treatment of the oxidized
samples. 2b: Relative contribution of photo-Fenton and biological treatment to
mineralization of the different systems. Systems: A: without photo-Fenton treatment,
B, C, D and E: samples partially oxidized by photo-Fenton process in the presence of 9
mM H2O2, 3 doses of 9 mM H2O2, D: 27 mM H2O2 and E: 54 mM H2O2 respectively.
Figure 2a shows the temporal evolution of TOC during biotreatment of solutions
obtained after different photo-Fenton treatment. Consortium showed biodegradation
capability in all the systems containing IMZ byproducts. At the end of biotreatment, the
TOC decreased by 42%, 48% 60%, and 51% in solutions B, C, D and E respectively,
indicating that the oxidation treatment produces byproducts with higher biodegradability
than the former fungicide. The main biodegradation was produced in the first 5 days of
experiment, after that, all the systems reached a plateau. Biodegradation rate
decreases quickly during the experiment showing the initial rate close relationship with
the degree of previous photo-Fenton treatment. This fact suggests the presence of many
byproducts, from the easily and faster biodegradable to the more recalcitrant. The pH of
the culture rises all along the assay for the four conditions tested. Figure 2b shows that
the overall mineralization percentage for the combined systems was: 0%, 42%, 59%,
62% and 84% for the systems A, B, C, D and E, respectively. The D assay probed to be
more efficient in the biological treatment; despite that there are not significant
differences between assays C and D. The E assay displayed the higher degree of
mineralization.
Table 2 shows the ratio of TOC diminution by unit of hydrogen peroxide after photoFenton treatment and after the coupled treatment.
TABLE 2. Efficiecy in the use of H2O2 in the treatment
Assay
A
H2O2
Photo Coupled
Disminution TOC Coupled
(mM)
(mgTOC/mmolH2O2)
(mgTOC/mmolH2O2)
(%)
0
x 0 1.7 0.5 2.7 x 21.6 11.3 9.2 6.8 0
B
9
C
3x9
D
27
E
54
42
59
62
84
777
The efficiency for TOC elimination by hydrogen peroxide unit consumed decreases with
increasing the dose of peroxide in the coupled process. Once IMZ was degraded into
byproducts, the biological treatment takes care for mineralization. However, even if
minimization of photo-Fenton treatment time helps to save in expensive reagents,
decreasing hydrogen peroxide dose leads to a lesser degree of refining of wastewater. It
should be assume a compromise between the efficiency of utilization of hydrogen
peroxide and the overall efficiency of process. This will depend mainly on regulatory
discharge standards
Conclusion
Here we report the isolation of a microorganism consortium, that are resistant to the IMZ
present in the commercial pesticide Xedrel 50®. The photo-Fenton process here tested
was useful for the degradation of this recalcitrant compound. We show the importance to
apply an AOT before the biological treatment, due to a significant improvement in the
biodegradability of the byproducts with respect to the original recalcitrant compound. It
should be mention that IMZ can not be adsorb or degraded by the microorganisms
growing in suspension. This coupled process shows to have great potential for treatment
of IMZ for waste water from fruit industry.
Acknowledgments
This work was supported by project FSNano-08 (ANPCyT, Argentina). GC and RC are
members of CONICET.
References
Ballesteros M., Sánchez-Pérez J. A., Casas-López J. L., Oller I, Malato-Rodríguez S. (2009). Degradation
of a four-pesticide mixture by combined photo-Fenton and biological oxidation. Water Res. 43: 653660.
Ballesteros M.M., Casas-López J.L., Oller I., Malato S., Sánchez-Pérez J.A. (2010). A comparative study
of different tests for biodegradability enhancement determination during AOP treatment of
recalcitrant toxic aqueous solutions. Ecotoxicol. Environ. Saf. 73: 1189-1195.
Dunia E. S., Doña-Rodríguez J. M., Araña J., Fernández-Rodríguez C., González-Díaz O., Pérez-Peña
J., Silva A. M. T. (2013). Optimization of the degradation of imazalil by photocatalysis: Comparison
between commercial and lab-made photocatalysts. Applied Catalysis B: Environmental. 138 - 139:
391-400.
EPA (1999). Draft Guidelines for Carcinogenic Assessment. Environmental Protection Agency, United
States.
Hazime, R., Ferronato, C., Fine, L., Salvador, A., Jaber, F., Chovelon, J. M. (2012). Photocatalytic
degradation of imazalil in an aqueous suspension of TiO 2 and influence of alcohols on the
degradation. Applied Catalysis B: Environmental 126: 90-99.
Hincapie- Perez M., Peñuela G., Maldonado M.I., Malato O., Fernandez-Ibanez P., Oller I., Gernjak W.,
Malato S. (2006). Degradation of pesticides in water using solar advanced oxidation processes.
Applied Catalysis B: Environmental 64: 272-281.
Lapertot M., Ebrahimi S., Dazio S., Rubinelli A., Pulgarin C. (2007). Photo-Fenton and biological
integrated process for degradation of a mixture of pesticides. Journal of Photochemistry and
Photobiology A: Chemistry 186: 34±40.
Malato S., Fernández-Ibáñez P., Maldonado M. I., Blanco J., Gernjak W. (2009). Decontamination and
disinfection of water by solar photocatalysis: Recent overview and trends. Catalysis Today 147: 159.
778
Oller I., Malato S., Sanchez-Perez J.A., Maldonado M.I., Gassó R. (2007). Detoxification of wastewater
containing five common pesticides by solar AOPs±biological coupled system. Catalysis Today 129:
69-78.
Pérez, M., Torrades, F., Domenech, X., Peral, J. (2002). Fenton and photo-Fenton oxidation of textile
effluents. Water Research, 36(11): 2703-2710.
Pignatello J. J., Liu D., Huston P. (1999). Evidence for an additional oxidant in the photo assisted Fenton
reaction. Environmental science and Technology 33: 1832-1839.
Pupo-Nogueira R.F.,
Oliveira M.C., Paterlini W.C. (2005). Simple and fast spectrophotometric
determination of H2O2 in photo-Fenton reactions using metavanadate. Talanta 66: 86-91.
Santa Cruz Biotechnology, Inc. (2010). MSDS Imazalil, USA.
Soon A. N., Hameed B. H. (2011). Heterogeneous catalytic treatment of synthetic dyes in aqueous media
using Fenton and photo-assisted Fenton process. Desalination 269: 1-16.
Zapata A., Malato S., Sanchez-Perez J.A., Oller I., Maldonado M.I. (2010). Scale-up strategy for a
combined solar photo-Fenton/biological system for remediation of pesticide-contaminated water.
Catalysis Today 151: 100-106.
779
CHAPTER 7.10. BIOTECHNOLOGY PROCESS FOR THE VALORIZATION OF THE
LIQUID RESIDUAL EFFLUENTS OF THE CITRIC INDUSTRY
León-Becerril, Elizabeth (1); Zárate-Cruz, César Augusto (1); Díaz-Torres, José de
Jesús (1), López-López, Alberto* (1)
(1) Unidad de Tecnología Ambiental, Centro de Investigación y Asistencia en Tecnología y Diseño del
Estado de Jalisco A.C. Av. Normalistas 800, Colinas de la Normal, 44270. Guadalajara, Jalisco, Mexico
ABSTRACT
Orange juice production generates important quantities of solid (citrus peel, pulp and
seeds) and liquid wastes (wash water, evaporator and cold press wastes). Solid wastes
have a composition high in sugars, cellulose, hemicellulose, pectin, organic acids and
essential oils. The aim of this work was to propose a biotechnology process for the
valorization of the orange wastes for a citric company in Mexico. This company has a
system for the combustion of orange peel as a source of energy, this process requires
the reduction of orange peel humidity from 82% to 35% H2O(w/w); therefore solids are
mechanically pressed generating a liquid residual effluent, referred as peel press liquor;
which represents a problem of contamination, so it is necessary to propose a process for
the valorization of this liquid effluent. Firstly, a physicochemical characterization of the
effluent was made; the effluent presents pH= 4.0, high concentrations in organic matter
(57 gBOD/l); total sugars (48.0 g/l) mainly reducing sugars; in addition of value added
byproducts such as essential oils (d-limonene, 0.86 g/l), organic acids as citric (2.1 g/l)
and succinic (3.6 g/l). Then the process consisted on the recovery of d-limonene as a
value-added product. In the next stage, an anaerobic process was employed for the
treatment of the liquid residual effluent (wastewater) in order to reduce the concentration
of organic matter and to produce in parallel methane; the absence of d-limonene avoid
the inhibition of microbial activity. The batch reactor was inoculated with biomass
stabilized from an anaerobic wastewater plant. Adaptation and growth of biomass last 20
days, removal efficiencies of organic matter reached 70 to 80%, for equivalent organic
loading rates from 20 to 30 kgCOD/m3d. The production rate of methane was
0.416 lCH4/gCODremoved at operating conditions. Then a study of anaerobic
biodegradation of organic matter with a COD concentration of 96 g/l, was conducted for
30 days, maximum COD removal efficiency is approximately 65%. It was established a
biotechnology process for recovery of d-limonene, and the subsequent wastewater
treatment by an anaerobic process, producing biogas with a high concentration in
methane, which can be used as an energy source.
Key words: anaerobic process, citric industry, D-limonene, methane, organic matter,
wastewater treatment
*Author for correspondence: [email protected]
Introduction
780
Cultivation of oranges is an important economic sector in Mexico, occupying the fifth
place in worldwide production with 3.6x106 ton in 2012, after Brazil, United States, China
and India (FAOSTAT, 2014). An important percentage of orange production (70%) is
used for manufacturing derivative products, such as juice and jams, about 50 to 60% of
the processed fruit becomes in solid waste such as citrus peel, pulp and seeds and in
liquid waste which include wash water, evaporator and cold press wastes (Wilkins et al.,
2007). These wastes should be used as resources, because of its potential to be
transformed in energy or as raw material in the manufacture of other products. Solid
wastes have a composition high in sugars, cellulose, hemicellulose, pectin, organic
acids and essential oils (Grohmann et al., 1999). In order to prevent problems related to
the disposal and environmental concerns, this waste must be properly processed,
because a large fraction is still deposited every year, causing economic and
environmental problems.
Several works has been devoted for the valorization of citrus wastes, some
applications include fertilizer; animal feed; adsorption of chemical compounds i.e.
orange peel was found to be effective (a 85%) for arsenic removal from real water
systems containing concomitant ions (Khaskheli et al., 2011); bio-oil production i.e.
bio-oil yield represents up to 60 g/100 g (by weight) of the raw material used (Rezzadori
et al., 2012); ethanol and extraction of essential oils and pectin, in this aspect
Pourbafrani et al. (2010) suggested a process obtaining 39.64 l of ethanol, almost 45 m 3
of methane, 8.9 l of D-limonene per ton of wet citrus waste.
Although, the alternatives proposed for the citrus wastes valorization, most of them
require expensive pre-treatments, depending on the process or product to be obtained
(Martin et al., 2010). For some of the valorization products, such as fuels (ethanol or
methane), it is necessary to remove D-limonene, a citrus oil present in those wastes.
D-limonene is toxic for digesting microorganisms and decreases biogas yield, or it is a
strong inhibitor for microorganisms if the liquid effluent is used for ethanol production
(Wilkins et al., 2007; Boluda-Aguilar et al., 2010; Choi et al., 2013). In contrast,
D-limonene represents a value-added product that could be recovered from this wastes;
D-limonene is employed in the manufacture of domestic household products, food and
medicines as flavoring agents, and cosmetics and bactericide products (Pourbafrani et
al., 2010). It was determined that orange peel contains 5.436 kg of oil per ton of
oranges, of which 90% of D-limonene (Braddock, 1995; Rezzadori et al., 2012).
Anaerobic digestion is one alternative that can be technical and economical viable.
This process produces methane that can be used as an energy source due to its high
heating value (35,793 kJ/m3 STP) and it is efficient because it permits the treatment of
the liquid effluent at high organic loading rates at different temperature intervals (Khanal,
2008). Anaerobic digestion has been applied with success for citrus wastes, at both
thermophilic and mesophilic conditions at laboratory scale; although there is little
information of citrus wastes digestion. Martin et al. (2010) studied the mesophilic
anaerobic digestion of orange peel waste after removing about the 70% of D-limonene;
obtaining for organic loading rates (OLR) 1.20 to 3.67 kgCOD/m 3.d, with a solid retention
time of 25 days, a methane yield of 270 to 290 NmlCH4/gCODadded, with a
biodegradability of 84-90%; acidification was observed at high OLR. Kaparaju and
Rintala (2006) worked at termophilic conditions for the anaerobic digestion of orange
pulp and peel; they obtained at batch operation, a methane production rates about
0.49 m3/kgVSadded waste. In semi-continuous digestion, loading of 2.8 kgVS/m3.d and
781
hydraulic retention time (HRT) of 26 days generated a specific methane yield of
0.60 m3/kgVS. However, this required adjusting pH (from an initial 3.2 to 8.0) by adding
CaCO3. Finally, Martin et al. (2013) proposed the co-digestion of orange peel with
glycerol derived from biodiesel manufacturing at laboratory scale; because glycerol
could reduce the inhibitory effect of some compounds and provide a correct nutrient
balance. Under mesophilic and semi-continuous conditions, a mixture of orange peel
waste-residual glycerol of 1:1 (in COD) operated at organic loading rates up to
2.10 gVS/l. The methane yield coefficient was quite constant, with a mean value of
330±51 Nml/gVSadded, while the OLR reached a mean value of 1.91±0.37 kgVS/m3.d
(17.59±2.78 kg mixture/m3.d) and the hydraulic retention time varied in a range of 8.5±
30.0 d.
A few works have been proposed for wastewater generated from the pressing of
orange peel, such as Siles et al. (2008) and Santos et al. (2010). Those authors studied
the mesophilic anaerobic digestion at laboratory scale of wastewater generated from the
pressing of orange peel, obtaining 295-297 NmlCH4/gCODremoved.
The aim of this work was to propose a biotechnology process for the valorization of
the orange wastes for a citric company in Veracruz, Mexico. This company has a system
for the combustion of orange peel as a source of energy, this process requires the
reduction of orange peel humidity from 82% to 35% H2O(w/w); therefore solids are
mechanically pressed to remove as much as liquid as possible. Also, this process
requires the addition of lime (Ca(OH)2) as binder to elevate pH from 3.5 to 9-9.5, and it
generates a liquid residual effluent, referred as peel press liquor. The solids are dried in
a rotary dryer and then burnt for generation of thermal energy, in the form of steam that
could be used for process steam requirements and/or generation of electricity. Although,
peel press liquor represents a problem of contamination, it is heavily pollutant due to its
high concentration of organic matter (100.0 gCOD /l) and total suspended solids (90.0
g/l), so it is necessary to propose a process for the valorization of this liquid effluent. A
physicochemical characterization of the liquid residual effluent was made and a proposal
of recovery of value-added products and the treatment of the effluent with production of
biogas.
Liquid residual effluent. The liquid residual effluent corresponds to the peel press
liquor, generated in the pressing of orange peel and the liquid effluent (wastewater)
generated after the extraction of D-limonene. This wastewater was treated by anaerobic
digestion. The liquid residual effluents came from a company that produces orange juice
in Veracruz, Mexico. The residual effluent was refrigerated at 4ºC until it was used for
the experiments.
Analytical methods. The following parameters were determined in the peel press liquor
and in the wastewater: pH, alkalinity, acidity, chemical oxygen demand (COD),
biochemical oxygen demand (BOD), fat, oils and grease (FOG), total solids (TS), total
suspended solids (TSS), volatile suspended solids (VSS), total dissolved solids (TDS).
All the analyses were carried out in accordance with the Standard Methods of the
American Public Health Association (APHA, 1998). Also, phenols by Folin-Ciocalteu
method (Folin and Ciocalteau, 1927) total (Dubois et al., 1956) and reducing sugars
782
(Miller, 1959) were determined. Organic acids were determined by the analytical method
5989-1265EN and volatile organic compounds were analyzed by HPGC.
Pretreatment of the wastewater. Given the high total and suspended solids content of
the wastewater, prior to anaerobic treatment, suspended solids were removed by
centrifugation for 30 minutes at 3500 rpm.
One of the main problems of treating this type of effluents is the high acidity and
null alkalinity; therefore solutions of NaOH 2M, Na2CaCO3 2M were used for pH
neutralization.
It was indispensable to provide the necessary nutrients for the appropriate
metabolism of the microorganisms involved in the process, (C:N:P = 350:7:1). Nitrogen
as NH4Cl was added to the final wastewater to be anaerobically digested (Khanal,
2008).
Anaerobic treatment. The reactor consisted of an Erlenmeyer flask glass of 2 liters with
constant mixing. Temperature was controlled at 37ºC; pH was fixed at 6.8. Biogas
(methane and carbon dioxide) generated during the anaerobic process passed through
a solution 3M NaOH for capturing the CO2 present in the biogas to Na2CO3, then
methane gas was measured by displacement of water in a column; methyl orange was
used as indicator of the saturation of NaOH solution (Figure 1).
FIGURE 1. Experimental system.
The reactor was inoculated with methanogenically activate biomass obtained from
a stabilized anaerobic wastewater treatment process. The specific methanogenic activity
was ranging from 0.15-0.89 gCOD/gVSS using acetate as substrate. The inoculation of
reactor was initially performed with 25% of the reaction volume corresponding to volatile
783
suspended solids (VSS) concentration about 10,000 mg/l.
A first stage of adaptation and growth of the biomass was performed; it consisted
in the initial feeding of the reactor feed with 9.60 gCOD/l, corresponding to the 10% of
the total COD concentration of the wastewater. COD concentration in the wastewater
feeding to reactor was gradually increased by 10% until the total COD concentration of
the wastewater about 96 gCOD/l. Each load lasted 48 hours; this time assures a COD
degradation greater than 50% and gas production in the reactor for each load. During
the batch experiments, the volume of methane was measured as a function of time;
samples of wastewater were taken for analysis of pH, alkalinity, COD before and after
feeding.
Once reached 100% of the total COD concentration, feeding of wastewater
stopped; then the stage of biologic degradation of the wastewater started. Organic
matter degradation and respectively were determined. Daily, volume samples of 10 ml
were taken from the reactor which represents less than 5% of the total volume reactor,
for analysis of pH, alkalinity and COD concentration, and methane production in function
on time were measured. At the end of this stage which lasted 30 days without feeding,
biomass growth rates measured as SSV concentration were determined.
Characterization of the liquid residual effluent. Physicochemical characteristics of
peel press liquor and wastewater (peel press liquor after the extraction of D-limonene)
are shown in Table 1.
TABLE 1. Physicochemical characteristics of the liquid residual effluent
Parameter
pH
Fat, oil and grease (g/l)
Alkalinity (g/l)
Acidity (g/l)
Total biochemical oxygen demand (g/l)
Soluble biochemical oxygen demand (g/l)
Total chemical oxygen demand (g/l)
Soluble chemical oxygen demand (g/l)
Total solids (g/l)
Total suspended solids (g/l)
Total dissolved solids (g/l)
Total volatile solids (g/l)
Total nitrogen (g/l)
Total phosphorus (g/l)
Sugars
Total sugars (g/l)
Reducing sugars (g/l)
Organic volatile compounds
D-limonene (g/l)
Terpinolene (g/l)
D-terpineol (g/l)
Peel press liquor
3.9
0.99
#0
1.30
56.50
50.50
129.30
117.15
101.58
9.85
93.28
94.68
N.I.
N.I.
Wastewater
4.0
0.48
#0
1.11
58.50
46.75
104.00
95.90
97.57
12.90
84.67
91.70
0.08
3.60
47.82
42.45
64.06
38.60
0.860
0.049
0.091
0.48
N.I.
N.I.
784
Organic acids
Tartaric acid (g/l)
Lactic acid (g/l)
Citric acid (g/l)
Succinic acid (g/l)
N.A.
N.A.
N.A.
N.A.
0.30
1.11
2.12
3.63
N.D. Not Identified
N.A. Not Analyzed.
Both liquid effluents are very pollutant; they present pH= 4.0 as a result of a high
acidity (1.30 gCaCO3/l) and null alkalinity; high concentrations in organic matter around
50.00 gBOD/l and 100.00 gCOD /l, about 90% is soluble; which can be confirmed for the
high concentration in total dissolved solids. As the same way, the same proportion is
observed between the concentrations of total dissolved solids and total solids. The
reduction of BOD and COD concentrations in the wastewater is associated with the
removal of fat, oil and greases due to the presence of D-limonene. In addition, both
effluents present a high quantity of total sugars, mainly reducing sugars, which offers the
alternative of fermentation and consequently the production of ethanol (Pourbafrani et
al., 2010).
Peel press liquor present value added byproducts such as essential oils (Dlimonene), about 1.0 g/l. In the company of study, only the 50% of D-limonene can be
recovered; this is for the low efficiency of the extraction process. Although, recovery of
this oil is economically and technically important. Other organic oils are present such as
terpinolene and D-terpineol. Organic acids as citric (2.1 g/l) and succinic (3.6 g/l) are
also present in the wastewater.
Anaerobic treatment. The anaerobic treatment of the wastewater is technically
feasible; Figure 2.a presents the efficiency of removal organic matter in the stage of
adaptation and growth of the biomass; this stage lasted 20 days.
785
100
a)
80
60
40
20
0
4
Alkalinity (g/l)
Alkalinity
7.1
pH
6.9
b)
3
6.7
6.5
2
pH
COD removal efficiency (%)
6.3
6.1
1
5.9
0
5.7
0
20
40
60
80
100
Organic matter concentration (g COD/l)
FIGURE 2. Behavior of the reactor in the stage of adaptation and growth of the
biomass. a) COD removal efficiency. b) Alkalinity and pH behavior.
The reactor is stable when the concentration is less than 70% of the total
concentration of COD, about 58 gCOD/l. However if COD concentration is greater than
this value, stability problems were observed, mainly acidification of the reactor, although
the addition of neutralization agents, such as NaOH, Na 2CaCO3 (Figure 2.b); the reactor
is not stable. About 0.08 g NaOH/gCOD is required to increase pH from 3.9 to 6.0; and
0.02 gNa2CO3/gCOD to raise pH from 6.0 to 6.8.
Removal efficiency of COD increased gradually with respect to the COD
concentration in the feeding until 40 g/l. For COD concentrations from 28.8 to 57.60 g/l
equivalents to organic loading rates from 14.4 to 28.8 kg/m 3.d, reached average
efficiencies of COD removal of 75%. This period is considered stable, in spite of the high
COD concentrations, methane was produced and COD removal efficiency was
maintained constant, and pH could be controlled with external alkalinity addition.
Then the reactor is unstable for COD concentrations up to 58.00 g/l, due to this,
removal efficiency of organic matter diminishes until 22.6%. This drop in the efficient is
associated with the accumulation of acidity, basically volatile fatty acids (VFAs) (Martin
et al., 2010; Khanal, 2008).
The methanogenic activity was determined from the production of biogas because
of organic matter degradation; the main components of biogas were basically methane
(65-70%) and carbon dioxide (30-35%) (Khanal, 2008). Biogas production rate is 0.416
lCH4/gCODremoved at operation conditions of temperature and pressure
(0.275 lCH4/gCODremoved at STP) (Figure 3). Methane production is comparable with
other results reported in the literature for orange press liquor of 0.295±
0.297 lCH4/gCODremoved at STP (Siles et al., 2008; Santos et al., 2010). Methane
production is comparable with the theorical value of 350 lCH4/gCODremoved at STP,
786
showing an acceptable methanogenic activity and indicating the stability of the reactor at
different COD concentrations. These values are accord with the specific methanogenic
activity of 0.15-0.89 gCOD/gVSS using acetate as substrate (Martin et al., 2010).
Methane gas volume (l/d)
20
16
12
8
4
0
0
10
20
30
40
50
COD removed (g/d)
FIGURE 3. Graphical determination of the methane production rate.
From the adaptation and growth of the biomass, a test for degradation of organic
matter was made taking into account the total COD concentration of the wastewater,
operating the reactor in batch mode. Figure 4 shows the degradation rate of organic
matter as COD. Maximum COD removal efficiency is approximately 65% at the end of
30 days; degradation rate in the first ten days of treatment is around 6g/d, although
degradation rate drop drastically might due to an accumulation of VFAs, therefore,
methane production rate diminishes in the same way (Khanal, 2008; Pourbafrani et al.,
2010; Pourbafrani et al., 2010; Martin et al., 2010). Biomass increased in a 16% of the
initial concentration of VSS; the maximum growth of the biomass can be seen between
days 10 and 15. At the end of the treatment, the quantity of biomass is lesser than the
initial quantity; this diminution might be due to inhibition of biomass by the high acidity of
the reactor and subsequently the poor substrate assimilation by microorganisms
provoking the biomass decay (Khanal, 2008).
COD/CODo (-)
1
40
0.8
30
0.6
20
0.4
COD
Methane
Biomass
0.2
10
0
0
0
5
10
15
20
25
Volume accumulated of methane (l)
Biomass concentration (gVSS/l)
30
Time (d)
787
FIGURE 4. Efficiency of removal organic matter, volume accumulated of methane
and biomass concentration behavior in the degradation stage.
Biotechnology process for the valorization of citric wastes. Figure 5 shows the
biotechnology process for the valorization of citric wastes, specifically the liquid residual
effluents for the citric company in Mexico. An extraction process for the recovery of
added-value products (D-limonene) is proposed and a biological treatment of the
wastewater by anaerobic digestion, resulting in the production of biogas with methane
concentrations about the 65-70%, which provides a potential source of energy (Khanal,
2008). The high concentration of total sugars, mainly reducing sugars offers the
alternative of its fermentation and consequently the production of ethanol.
D-limonene represents a value-added product that could be recovered from this
liquid effluent, D-limonene is employed in the manufacture of domestic household
products, food and medicines as flavoring agents, and cosmetics and bactericide
products. In addition, this compound is an inhibitory for microorganisms (Wilkins et al.,
2007; Boluda-Aguilar et al., 2010; Choi et al., 2013), therefore its recovery is highly
recommended.
In the citric company of study, peel press liquor is pumped to a multiple effect
evaporator, where D-limonene is recovered, almost the 50%, see Table 1.
The wastewater resulting from this stage could be treated by an anaerobic process
in order to reduce the concentration of organic matter and to produce in parallel
methane; the absence of D-limonene avoids the inhibition of microbial activity.
Juice
production
Orange wastes
Generation of
energy
Orange peel
Combustion
Solids
Pressing
Liquid residual effuent
(Press orange liquor)
Extraction
d-limonene
Liquid residual effuent
(wastewater)
Ashes
Ethanol
Destillation
Vinasses
Fermentation
Anaerobic
treatment
Methane
Wastewater
treated
FIGURE 5. Biotechnology process for the valorization of citric wastes of the citric
company in Mexico.
Conclusion
A biotechnology process was established for the valorization of citrus wastes of a citric
company in Mexico. D-limonene was recovered as a value-added product, in a 50%;
and the subsequent wastewater treatment was proved by an anaerobic process,
788
producing biogas with a high concentration in methane, which can be used as an energy
source. The removal of D-limonene was an important factor that avoided the inhibition of
the anaerobic microorganisms. Another factor that permitted the biological digestion of
organic matter was the addition of Na2CO3 to the reactor. The adaption of biomass to
substrate was possible, removal efficiencies of organic matter reached 70 to 80%, for
equivalent organic loading rates from 20 to 30 kgCOD/m 3d, and the production rate of
methane was 0.416 lCH4/gCODremoved at operating conditions. Anaerobic biodegradation
of organic matter with a COD concentration of 96 g/l reached a maximum COD removal
efficiency of approximately 65%.
Acknowledgements
Authors would thank the financial support of Project CONACYT-PEI 199475.
789
References
APHA-AWWA-WEF, 1998. Standard Methods for the examination of water and wastewater, in: Clesceri,
L.S., Greenberg, A.E., Eaton, A.D. (Eds.), American Public Health Assn-American Waste Works
Assn-Water Environment Federation, Washington DC.
Boluda-Aguilar M.; Garcia-Vidal L.; F.delP. G.-C.; Lopez-Gomez A. (2010). Mandarin peel wastes
pretreatment with steam explosion for bioethanol production. Bioresource technology. 101: 35063513.
Braddock R.J. (1995). By-products of citrus fruit. Food Technol. 49: 74-77.
Choi I.S.; Kim J.-H.; Wi S.G.; Kim K.H.; Bae H.-J. (2013). Bioethanol production from mandarin (Citrus
unshiu) peel waste using popping pretreatment. Applied Energy. 102: 204-210.
Dubois M.; Gilles K.A.; Hamilton J.K.; Rebers P.A.; Smith F. (1956). Colorimetric method for determination
of sugars and related substances. Anal. Chem. 28: 350-356.
FAOSTAT, http://faostat3.fao.org/faostat-gateway/go/to/home/E (Retrieved 01.08.14).
Folin C.; Ciocalteau V. (1927). Tyrosine and tryptophan determination in proteins. J. Biol. Chem. 73: 627650.
Grohmann K.; Manthey J.A.; Cameron R.G.; Buslig B.S. (1999). Purification of Citrus Peel Juice and
Molasses. J. Agric. Food Chem. 47: 4859-4867.
Kaparaju P.L.; Rintala J.A. (2006). Thermophilic anaerobic digestion of industrial orange waste.
Environmental technology. 27: 623-633.
Khanal S.K. (2008). Anaerobic Biotechnology for Bioenergy Production: Principles and Applications, Iowa.
Khaskheli M.I.; Memon S.Q.; Siyal A.N.; Khuhawar M.Y. (2011). Use of Orange Peel Waste for Arsenic
Remediation of Drinking Water. Waste and Biomass Valorization. 2: 423-433.
Martin M.A.; Fernandez R.; Serrano A.; Siles J.A. (2013). Semi-continuous anaerobic co-digestion of
orange peel waste and residual glycerol derived from biodiesel manufacturing. Waste management.
33: 1633-1639.
Martin M.A.; Siles J.A.; Chica A.F.; Martin A. (2010). Biomethanization of orange peel waste. Bioresource
technology. 101: 8993-8999.
Miller G.L. (1959). Use of DinitrosaIicyIic Acid Reagent for Determination of Reducing Sugar. Anal. Chem.
31: 426-428.
Pourbafrani M.; Forgacs G.; Horvath I.S.; Niklasson C.; Taherzadeh M.J. (2010). Production of biofuels,
limonene and pectin from citrus wastes. Bioresource technology. 101: 4246-4250.
Rezzadori K.; Benedetti S.; Amante E.R. (2012). Proposals for the resiudes recovery: Orange waste as
raw material for new products. Food and Bioproducts Processing. 90: 606-614.
Santos M.D.; Lopez J.A.; Perez A.F.; Martin A.M. (2010). Modelling the anaerobic digestion of wastewater
derived from the pressing of orange peel produced in orange juice manufacturing. Bioresource
technology. 101: 3909-3916.
Siles J.A.; Martín M.A.; Chica A.; Borja R. (2008). Kinetic modelling of the anaerobic digestion of
wastewater derived from the pressing of orange rind produced in orange juice manufacturing.
Chemical Engineering Journal. 140: 145-156.
Wilkins M.R.; Suryawati L.; Maness N.O.; Chrz D. (2007). Ethanol production by Saccharomyces
cerevisiae and Kluyveromyces marxianus in the presence of orange-peel oil. World Journal of
Microbiology and Biotechnology. 23: 1161-1168.
790
CHAPTER 7.11. INFLUENCE OF ORGANIC LOADING RATE ON THE HYDROGEN
PRODUCTION FROM AGRO-INDUSTRIAL RESIDUES AS SUBSTRATE USING
AFBR REACTORS
Samantha Christine Santos (1); Paula Rúbia Ferreira Rosa (2); Edson Luiz Silva*(2)
(1) Department of Hydraulics and Sanitation, São Carlos Engineering School, University of São Paulo,
São Carlos, SP, Brazil;
(2) Department of Chemical Engineering, Federal University of São Carlos, São Carlos, São Paulo, Brazil
ABSTRACT
The aim of this study was to evaluate the influence of hydraulic retention time (HRT) and
two different carbon sources (cheese whey and cassava processing wastewater) on the
stability and efficiency of anaerobic fluidized bed reactor (AFBR) in hydrogen production.
Four AFBRs were operated for 160 days with the temperature control fixed at 30 °C.
Two reactors used cheese whey as substrate (AFBR1 and AFBR2), and two other
reactors were feed with cassava processing wastewater (AFBR3 and AFBR4). The
reactors were inoculated with sludge from a UASB reactor used for swine wastewater
treatment. To inhibit methanogenesis in AFBR reactors, it was conducted heat pretreatment of the inoculum. For AFBR2, the organic load rate (OLR) ranged from 12 to 40
kg COD m-3 d-1 and the highest hydrogen yield (HY) of 3.2 mmol.g-1 COD was obtained
in the lowest OLR. When cassava processing wastewater was used as substrate for
AFBR3 the OLR ranged from 4 to 12 kg COD m -3 d-1 with a fixed HRT of 12 h, and
showed a maximum HY of 2.0 mmol g-1 COD (OLR 10 kg m-3 d-1). The main soluble
metabolites observed at all four bioreactors were ethanol, methanol, acetic, butyric,
propionic and lactic acids. The ability to produce H2 continuously, the organic acids
distribution and the bioreactor performance were influenced by factors such as
modifying the substrate organic source and also by decreasing the hydraulic retention
time and consequent increased organic loading rates applied. Cheese processed
wastewater fermentation showed a higher potential to produce hydrogen under the
studied conditions.
Keywords: Cheese whey, cassava wastewater, continuous anaerobic fluidized bed
reactors, hydraulic retention time.
Introduction
The increasing demand for renewable energy sources has become a global concern.
Currently, only 12% of the world's power is produced from renewable sources
(International Energy Agency, 2013), which demonstrates the strong and significant
dependence on fossil fuel use. The burning of fossil fuels has been generating a global
increase of greenhouse gas emissions and therefore the reduction of these features is
needed to minimize environmental problems (Show et al. 2012).
*Author for correspondence: [email protected]; [email protected]
791
Hydrogen gas is seen as a solution for sustainable energy source because it has
high conversion efficiency, is recyclable and non-polluting. Moreover, considering that
nowadays one of the biggest issues is replace the fossil fuels implications, H2 is an
inexhaustible and attractive idea (Mohan et al. 2008). Therefore, the recovery of
industrial waste through anaerobic processes represents a significant opportunity to
combine waste treatment and produce energy using renewable organic sources (Kargi
et al. 2012).
Cheese whey is the main waste product of the dairy industry, featuring a pressure
environmental impact by physic-chemical characteristics, including minerals (0.46 to
10%), suspended solids (0.1 to 22 kg m-3), phosphorus (0.006 to 0.5 kg m-3) and organic
matter (0.6 to 102 kg m-3). This wastewater is 99% biodegradable and at worldwide level
this residue had an output of 40x106 tons of cheese whey, representing a promising
activity to hydrogen biological production from this wastewater (Prazeres et al. 2012).
Cassava is a tropical root crop originally from Brazilian Amazonia that provides the
staple food of an estimated 800 million people worldwide, and the wastewater generated
by its processing has also the potential of being used as substrate for fermentative
hydrogen production. After 1 kg of fresh cassava root being peeled, washed, grated and
mixed with water during the starch production process, about 0.2 kg of starch, 0.4±
0.9 kg of residue, and about 5±7 liters of carbohydrate-rich wastewater are generated
(Cappelletti et al. 2011).
In this sense, hydrogen production from real wastewaters via dark fermentation has
DWWUDFWHG D VLJQL¿FDQW DPRXQW RI UHVHDUFK LQWHUHVW EHFDXVH RI KLJK +2 production rates
(HPR) and the ability to utilize pretreatment liquors derived from underutilized agriculture
residue (Sathyanarayanan et al. 2014). Some aspects that influenced hydrogen
production in anaerobic fluidized bed reactors have been studied, such as: adaptation of
the inoculum source of carbon in the reactor (Koskinen et al. 2007), optimization of
biomass adaptation to different hydraulic retention times (Barros e Silva, 2012; Shida et al.
2012; Zhang et al., 2007), substrate concentration (Lin et al. 2006; Zhang et al. 2007),
metabolites distribution (Lin et al. 2006; Shida et al. 2012; Zhang et al. 2007), support
material influence (Barros and Silva, 2012), applied temperature (Munoz -Paez et al.
2012) and upflow velocity (Reis and Silva, 2011). These studies were performed using
simple substrates, such as glucose and only a few were executed in AFBRs using
wastewater organic sources (Santos et al. 2014; Ferreira Rosa et al. 2014). In this
scenario, the present work aimed to evaluate the OLR effect and elucidate the
consequences of using different real organic sources (cheese whey and cassava
processing wastewater) under hydrogen production using four anaerobic fluidized bed
reactors.
Organic agro-industrial substrates. The cheese whey used as substrate in this study
were supplied by Cargill Agricultural S/A. This substrate was diluted to different
concentrations (3000 to 10.000 mg COD L-1). Already, the feeding substrate of cassava
processing wastewater was subjected to acid hydrolysis with sulfuric acid and heated at
120°C for 30 min, before being used in the feed of reactors, at different inlet
concentrations (2000 a 15.000 mg COD L-1). To all substrate feeding inlet it was added
792
micronutrients, according to Barros and Silva (2012) and the pH was controlled in the
range of 4.0 to 4.5 (NaHCO3 added to cheese whey and NaOH to cassava wastewater).
Inoculum source. All four reactors were inoculated with sludge from a UASB reactor
used in the treatment of swine wastewater. This inoculum was subjected to a heat
treatment according to Kim et al. (2006) to obtain an improvement in hydrogen
production by eliminating the methanogenic activity of metanogenics Archaea.
Bioreactors design and set-ups. This study was carried out by four reactors (total
each volume of 770 cm3, reactor height of 80 cm and internal diameter of 3.5 cm).
These anaerobic fluidized bed reactors (AFBR) for biohydrogen production was made by
acryl and jacketed with a water bath to maintain the temperature at 30 °C. Figure 1
shows a schematic illustration of the reactor. Polystyrene was used as support material
for biomass adhesion.
FIGURE 1. Installation scheme of the anaerobic fluidized bed reactor.
The anaerobic conditions were conducted from the use of nitrogen. The feeding of
the reactors were made with cheese whey powder (AFBR1 and AFBR2) and cassava
wastewater (AFBR3 and AFBR4). Pre-treated inoculum (heat pre-treated) was used
(10% v/v) and added to the feeding solutions. In this preliminary step, the system was
maintained in batch mode for 72 hours to enable the production of hydrogen by the
microorganisms, and then in continuous mode. When the steady state condition was
reached (variation of less than 10% in hydrogen production), changes in hydraulic
retention time and variations in the influent concentrations were made, as shown in
Table 1. Before introducing cassava wastewater to the reactors, the inoculum was
activated using a mixture of glucose and cassava in the influent as substrate.
793
The reactors were operated for 160 days, divided into four and five operational phases
for cheese whey and cassava wastewater as substrate, respectively. The composition of
biogas and soluble metabolite was monitored in function of time.
TABLE 1. Operating conditions of the reactors.
Substrate
Reactor
AFBR1
Cheese
whey
AFBR2
AFBR3
Cassava
wastewater
AFBR4
Phase
HRT
(h)
1
2
3
4
1
2
3
4
14
12
10
8
6
6
6
6
Substrate
concentration
(mg L-1)
5000
5000
5000
5000
3000
5000
7000
10.000
1
12
3000
2
3
4
5
12
12
12
10
2000
3000
5000
5000
1
12
15.000
2
12
7000
Carbon sources ratios
(%)
OLR
(*)
-
8.5
10
12
15
12
20
28
40
50% glucose and 50%
cassava
100% cassava
100% cassava
100% cassava
100% cassava
50% glucose and 50%
cassava
50% glucose and 50%
cassava
3
12
7000
100% cassava
4
12
9000
100% cassava
5
10
9000
100% cassava
HRT: hydraulic retention time; (*): Organic loading rate: kg COD m-3 d-1
6
4
6
10
12
30
14
14
18
22
Analytical methods. Periodic physical-chemical analyzes were performed on samples
collected from the influent and effluent, as well as biogas formed, to all four anaerobic
fluidized bed reactors. For monitoring purposes, constantly measures of liquid flow,
influent and effluent pH and bed height were performed. Determinations on influent and
effluent analyzes comprised total soluble carbohydrates (Dubois et al. 1956), organic
acids and alcohols by high-performance liquid chromatography (HPLC, Shimadzu ®) and
chemical oxygen demand (APHA: Standard Methods for the Examination of Water and
Wastewater, 2012). In biogas formed at the bioreactors, parameters, ie H2 yield,
volumetric production (gas meter type TG-1, Ritter®, Germany) and the composition
biogas (gas chromatograph, Shimadzu GC -2010®) were evaluated.
794
Biohydrogen production. Figure 2 shows the hydrogen production rate (HPR),
methane production rate (MPR), yield hydrogen (HY) and biogás content achieved at the
reactors AFBR1 and AFBR2, with cheese whey as substrate. The reactors performance
at each operational phase is shown in Table 2.
a)
b)
HPR/ M PR (L.d-1 . L -1 )
1
0,5
0
12
c)
H Y A FB R1
1
0
10
H2 C ontent
15
20
25
1,5
20
1
15
10
0,5
5
0
0
10
12
O LR (kg.m-3 .d-1 )
14
25
30
35
O LR (kg.m-3 .d-1 )
40
d)
C H4 C ontent
HY
30
8
2
(kg.m-3 .d-1 )
2
6
3
H Y (molH 2 .mol-1 lactose)
O LR
M PR A F BR2
4
14
B iogas Conte nt (%)
HPR/M PR (L.d-1 . L -1 )
1,5
10
HPR A F B R2
5
2
8
H Y (molH 2 .mol-1 lactose)
M PR A F BR1
H2 C ontent
C H4 C ontent
3
20
2,5
15
2
1,5
10
1
5
0,5
0
B iogas Conte nt (%)
HPR A F B R1
2,5
0
10
15
20
25
30
35
O LR (kg.m-3 .d-1 )
40
FIGURE 2. Performance of the reactors with cheese whey as substrate during OLR
increased (a) hydrogen and methane production at AFBR1 (b) Hydrogen and methane
production at AFBR2 (c) H2 yield and biogas content at AFBR1 (d) H2 yield and biogas
content at AFBR2.
For AFBR1, the volumetric hydrogen production increased from 1.92 to 2.16 L d -1 L1, conformed OLR increased from 8.5 to 10 kg COD m-3 d-1 (Figure 2a). Methane was
detected at the biogás formed, with productions of 0.40 L d -1 L-1 and 1.29 L d-1 L-1 (OLRs
of 12 and 15 kg COD m-3 d-1, respectively), resulting in decreased hydrogen production
rate, from 1.76 L d-1 L-1 (12 kg COD m-3 d-1) to 1.25 L d-1 L-1 (15 kg COD m-3 d-1). For
hydrogen content in the biogás it was observed higher porcentages (19 to 25%)
considering OLR variation (8.5 to 12 kg COD m -3 d-1 (Figure 2c). However, at the
organic loading rate of 15 kg COD m-3 d-1 it was verified lower values for H2, 9%, at the
hydraulic retention time of 8 h. During the operational phases with 12 and 15 kg COD
m-3 d-1 applied, methane was observed with 4 and 15%, respectively. The verification os
this trend (methane production) at low pH is already common in the literature, from the
use of cheese whey as substrate, impairing the performance of biological reactors to
produce H2 (Azbar et al. 2009; Carrillo-Reyes et al. 2012; Castello et al. 2009. The
795
maximum HY value for AFBR1 was 1.9 mol H2 mol-1 lactose (3.2 mmol.g-1 COD) at the
lowest OLR applied. From the increase in OLR from 8.5 to 15 kg COD m -3 d-1, there was
a decrease at HY from 1.9 to 0.7 mol H2 mol-1 lactose.
TABLE 2. Performance of the reactors AFBR1 and AFBR2 using cheese whey as
substrate.
Reactor HRT
(h)
OLR
HPR
MPR
HY
HY
(Kg COD.m-3.d-1)
(L.d-1.L-1)
(L.d-1.L-1)
(mol.mol-1
lactose)
(mmol.g-1
COD)
2.6
14
8.5
1.92
0
1.9
AFBR1
12
10
2.16
0
1.2
2.4
10
12
1.73
0.4
1.1
1.7
8
15
1.25
1.3
0.7
1.0
6
12
3.1
0
2.3
3.2
6
20
4.7
0
1.2
2.7
AFBR2
6
28
2.7
0.7
0.7
1.2
6
40
0.7
1
0.2
0.2
HRT: hydraulic retention time; organic loading rate; HPR: hydrogen production
rate; MPR: methane production rate; HY: hydrogen yield
For the reactor AFBR2, the increase OLR from 12 to 20 kg COD m -3 d-1 resulted in
higher hydrogen production rate, from 3.1 to 4.7 L d -1 L-1 (Figure 2b), however showed a
decrease to 0.7 L d-1 L-1 at the applied OLR of 40 kg COD m-3 d-1 because the presence
of methane (1 L CH4 d-1 L-1). The hydrogen content in the biogas was stable (15-17%) in
accordance with the OLR increase from 12 to 28 kg COD m-3 d-1. For the OLRs of 28
and 40 kg COD m-3 d-1 it was verified the presence of methane with 4 and 8% (Figure
2d), respectively. Regarding hydrogen yield, with increasing OLR from 12 to 40 kg COD
m-3 d-1, there was a decrease of HY from 2,3 mol H2 mol-1 lactose to 0,2 mol H2 mol-1
lactose, showing that in this case, the increase in OLR inhibited the production of
hydrogen.
Therefore, the maximum HY obtained from cheese whey as organic substrate was
2.3 mol mol-1 lactose (AFBR2) and 1,9 mol mol-1 lactose (AFBR1). These results are
superior to those obtained in a previous study by Rosa et al. (2014), that evaluated
cheese whey fermentation and obtained HY of 1.1 mol H2 mol-1 lactose, applying a HRT
of 6 hours, with an initial substrate concentration of 5000 mg COD L-1. Davila Vasquez
et al. (2009) achieved similar HY of those obtained in this study, using a CSTR at
differents HRT (10, 6 and 4 hours). The authors obtained the highest yield of 2,4 mol
mol-1 lactose at the highest HRT applied (10 hours), ie similar to the behavior obtained in
this study.
Continuous hydrogen production from cassava wastewater was also carried for
hydrogen production. Therefore, as well for cheese whey, hydrogen yields (HY), biogas
content, hydrogen production rate (HPR) and methane production rate (MPR) are shown
in Figure 3, considering this organic substrate.
In AFBR3, the H2 content of the biogas ranged from 4-16% and maximum value
was obtained in phase 4 (OLR of 10 kg COD m-3 d-1). In the phases 4 and 5 were found
796
methane in the biogas content with 2 and 8%, respectively. Regarding HPR, the
increase of OLR from 4 to 10 kg COD m-3 d-1 resulted in an increase from 0.3 to 1.66 L
H2 d-1 L-1, decreasing to 0.7 L H2 d-1 L-1 at OLR of 12 kg COD m-3 d-1due to the
appearance of methane (0.4 L CH4 d-1 L-1). Total replacement of glucose by cassava
wastewater had a positive effect on HY, increasing from 0.5 to 1.3 mmol g -1 COD in
phase 1 to 2. From the exclusive use of cassava as a carbon source, maximum HY of
2.0 mmol g-1 COD was obtained in phase 4 (10 kg COD m-3 d-1). Thus, the OLR of 10 kg
COD m-3 d-1 was considered optimal OLR, making it possible to obtain highest values of
HY and HPR from the use of cassava wastewater in this type of bioreactor.
a)
HY
% H2
b)
% C H4
HY
3
% H2
1
5
0
0
1
2
3
4
5
30
25
2
20
15
1
10
5
0
0
0
1
2
3
Phases
c)
H PR
Biogas Content (%)
10
H Y (mmol H 2.g-1 C O D)
15
2
0
4
5
Phases
d)
M PR
2
H PR
M PR
2,5
H PR/ M PR (L .d -1. L -1)
H PR/ M PR (L .d -1. L -1)
% C H4
3
Biogas Content (%)
H Y (mmol H 2.g-1 C O D)
1,5
1
0,5
0
0
1
2
3
Phases
4
5
2
1,5
1
0,5
0
0
1
2
3
4
5
Phases
FIGURE 3. Performance, in each operational phase, of the reactors feed with
cassava wastewater: (a) HY and biogas content in AFBR3 (b) HY and biogas content in
AFBR4 (c) HPR and MPR in AFBR3 (d) HPR and MPR in AFBR4.
HY: hydrogen yield; HPR: hydrogen production rate; MPR: methane production rate.
In AFBR4 the content of hydrogen in the biogas ranged from 1 to 25%, with
maximum in phase 2 (25%) and 3 (14%). Similarly in AFBR3, for phases 4 and 5 were
found methane content in biogas with 14 and 27%, respectively. The strategy of
reducing the OLR from 30 to 14 kg COD m-3 d-1 (phases 1 and 2, with 50% glucose and
50% cassava) resulted in an HPR increase from 0.12 to 2.1 L H2 d-1 L-1. From the
exclusive use of cassava as substrate (phase 3), there was observed that the OLR
increased resulted in a decrease of HPR (1.1 to 0.03 L H2 d-1 L-1), showing that different
operational strategies of the reactors resulted in different behaviors and trends. The
values of MPR were 0.9 L CH4 d-1 L-1 (phase 4) and 1.5 L CH4 d-1L-1 (phase 5). The
maximum HY of 1.53 mmol g-1 COD was obtained in phase 2 which used a mixture of
50% glucose and 50% cassava wastewater. From the exclusive use of cassava as
carbon source (phases 3, 4 and 5), the increasing of OLR (14 to 22 kg COD m -3 d-1)
resulted in a decrease of HY from 0.8 to 0.01 mmol g-1 COD, showing that the highest
OLRs also inhibited the production of hydrogen in this reactor. Aktusu et al. (2009)
evaluated the effect of substrate concentration on the production of H 2 from starch as
substrate. The increase of OLR from 15 to 60 kg.m -3 d-1 resulted in a drop in HY of 60%,
same behavior observed in the reactor AFBR4.
797
However, for AFBR3, the exclusive use of cassava wastewater (phase 2 and 3)
cause no modifications at HY. Phase 4 (OLR of 10 kg COD m -3 d-1) resulted in maximum
HY of 2 mmol g -1COD. From this OLR there was a decrease of this parameter, to 0.68
mmol g -1COD. This same behavior was observed by Thanwised et al. (2012). The
authors evaluated the effect of OLR (16-130 kg COD m-3 d-1) on the hydrogen
production from tapioca wastewater. In a first phase, the authors applied an OLR of 16
kg CODm-3 d-1 which resulted in 0.4 mmol g-1 COD. With the increase in OLR to 22 kg
COD m-3 d-1, the HY remained stable at 0.5 mmol g-1 COD. The optimal OLR was 31 kg
COD m-3 d-1 with maximum values of HY (0.745 mmol g-1 COD). From this OLR, the
authors reported lower HY values (0.22 mmol g -1 COD). These authors associate this
drop due to modified channels in which fermentation was used to produce other
products, such as ethanol, acetic and butyric acids competition instead of hydrogen.
In comparison among the four reactors, the highest HY of 3.2 mmol.g-1 COD were
obtained by AFBR2 from the use of cheese whey as substrate. This fact can be
explained by the complexity of cassava wastewater used as a carbon source when
compared to cheese whey.
Soluble metabolites distribution. In Figure 4, the main metabolites composition
associated with reduced HRT and increased OLR is displayed. For AFBR 1 (cheese
whey as substrate), the OLR increased (8.5 to 12 kg COD m -3 d-1) was accompained by
etanol variation from 20 to 29%. However, the OLR applied of 15 COD m -3 d-1 showed
decreases in both the percentage of ethanol (15%) and the parameter HY. Methanol
and butiric acid remained at stable porcentages,from 17 to 21% and 20 to 29%,
respectively. Applying the OLR of 8.5 and 10 kg COD m -3 d-1 resulted in acetic acid
decreases from 31 to 17%. In other organic loading rates this metabolite was observed
with average values of 27%. For propionic acid, the OLR increased of 8.5 to 15 kg COD
m-3 d-1, resulted in a HPr accumulation from 1% to 16%.
The OLR variation from 12 to 40 kg COD m-3 d-1 at AFBR2 showed that there was a
drop of ethanol from 38 to 12%, which was accompanied by a HY decrease. Methanol
and butyric acid ranged from 19 to 31% and from 14 to 19%, respectively. For propionic
acid, only at the OLR of 12 kg COD m-3 d-1 the values of this metabolite were below 1%,
whereas for other OLRs, the percentages were between 10 and 14%. Acetic acid had an
accumulation of 15 to 38% through OLR increase from 12 to 40 kg COD m -3 d-1,
impairing the reactor performance.
It is observed that the metabolic pathway for the hydrogen and ethanol production
was strongly influenced by HRT reduction and/or OLR increased. For both reactors
using cheese whey as substrate, higher OLRs applied, resulted in an inhibition of the
production of hydrogen and ethanol. Moreover, to the reactor AFBR2, with the highest
OLRs used, there was an accumulation of HAc. This accumulation of acids affected the
hydrogen yield, as reported by other authors (Searmsirimongkol et al. 2011; Singh et al.
2013).
798
Ethanol
M olar pe rce ntage of soluble
me tabolites (%)
a)
M olar pe rce ntage of soluble
me tabolites (%)
b)
Methanol
Acetic acid
Propionic acid
Butyric acid
100
80
60
40
20
0
8,5
Ethanol
Methanol
10
12
-3
O LR (kg.m .d-1 )
Acetic acid
Propionic acid
15
Butyric acid
100
80
60
40
20
0
12
20
28
O LR (kg.m-3 .d-1 )
40
FIGURE 4. Composition of the main metabolites using cheese whey as substrate
(in molar percentages) (a) AFBR1 (b) AFBR2.
Van Ginkel and Logan (2005) reported the inhibitory effect of acetic and butyric
acid, adding 25 mM of these acids in the feed of the continuous reactor, with glucose as
a substrate, and reported a reduction in hydrogen yield attributing this decrease to the
presence of 13 and 22% of HAc and HBu.
Considering the reactors AFBR3 and AFBR4, feed with cassava wastewater,
ethanol, lactic acid, methanol, butyric acid, propionic acid and acetic acid , were found to
be major soluble metabolites during the experiments (Figure 5). In AFBR 3, for phase 1 to
4, ethanol and methanol were stable with values between 35-40% and 14-17%,
respectively, but decreased to 15% and 8% in phase 5, also reflecting the drop of HY
(Figure 5 c), as well as reactors with cheese whey. Moreover, to phase 5 (highest OLR
applied of 12 kg COD m-3 d-1) it was observed the presence of valeric acid (18%), which
also affect performance of the reactors. Lactic acid was found only in phase 1, with 8%
and no longer detected for the other phases. The percentages of acetic acid (12-30%)
were similarly the values of butyric acid (12-24%). Propionic acid was detected in
amounts less than 14%. For phases from 1 to 4, the metabolic pathway of ethanolacetate was dominant with a sum of ethanol and acetate than 65% of the sum of soluble
metabolites (SMP).
799
Regarding in AFBR4, with the reduction of OLR (phase 1 to phase 2), there was an
increase of 19 to 54% and 0.6 to 11% of ethanol and butyric acid, respectively. There
was also a decrease from 46 to 23% and from 19 to 2% lactic acid and propionic, acid
respectively. From the exclusive use of cassava wastewater (phase 3), there was
observed an ethanol drop from 54 to 34%, and an increase of 9 to 17% of acetic acid;
10 to 20% of butyric acid. In phases 4 and 5, the concentrations of lactic acid and
ethanol reached zero, and there was the appearance of valeric acid (14-19%).
FIGURE 5. Soluble metabolites as relative percentages using cassava (a) in
AFBR3 (b) in AFBR4 (c) H2 yield and main soluble metabolites (molar %) of lactic acid
and ethanol in AFBR3 (d) H2 yield and main soluble metabolites (molar %) of lactic acid
and ethanol in AFBR4.
For both reactors, even with different operating strategies, both had the presence
of valeric acid. According Saady et al. (2013), valeric acid may decrease the yield of
hydrogen during fermentation in the dark, using mixed culture. According to the authors,
the Equation 1 and 2 show the formation of valeric acid from the hydrogen consumption,
thus justifying the lowest values of HY in phase 5 for AFBR3 (0.68 mmol g-1 COD) and
the phases 4 and 5 for AFBR4 (0.2 and 0.01 mmol g-1 COD).
CH3CH2COO-+2CO2+6H2 ĺ&+3(CH2)3COO-+4H2O
3CH3COO- +3H2+ 2H+ ĺ&+3(CH2)4COO-+4H2O
(1)
(2)
Furthermore, the reduction of alcohols in phase 4 and 5 was due to the appearance of
methane, where ethanol and methanol are converted to methane, as shown in the
Equations 3 and 4 (Saady et al. 2013).
800
4CH32+ĺ&+4+CO2+2H2O
2CH3CH2OH +CO2 ĺ&+4+ 2CHCOOH
(3)
(4)
Conclusion
Hydrogen production using cheese whey and cassava processing wastewater as
organic substrate in AFBRs was investigated in this study. From the cheese whey as
substrate, the máximum H2 yields of 2.6 mmol.g-1 COD (AFBR1) and 3.2 mmol.g-1 COD
(AFBR2) were verified at the applied OLR of 8.5 and 12 kg COD m -3 d-1, respectively.
Already, from cassava wastewater, HY of 2 mmol.g-1 COD (AFBR3) and 1.5 mmol.g-1
COD ((AFBR4), values lower than those fed reactors with cheese whey. Metabolic
pathway of ethanol was strongly influenced by the hydraulic retentio time reduction
and/or increasing the OLR for all the reactors. Besides this, cheese whey wastewater
fermentation showed a higher potential to produce hydrogen under the studied
conditions.
Acknowledgments
The authors thank FAPESP for financial support (Thematic Project, process
2009/15984-0) and CNPq research fellowship (Process146730/2010-0).
References
Akutsu, Y., Lee, D.-Y., Chi, Y.-Z., Li, Y.-Y., Harada, H., Yu, H.-Q. 2009. Thermophilic fermentative
hydrogen production from starch-wastewater with bio-granules. International Journal of Hydrogen
Energy, 34(12), 5061-5071.
APHA, AWWA, WEF., 2012. Standard methods for the examination of water and wastewater. 22th edition,
American Public Health Association, Washington, D.C.
Azbar, N., Dokgoz, F.T.C., Keskin, T., Korkmaz, K.S., Syed, H.M. 2009. Continuous fermentative
hydrogen production from cheese whey wastewater under thermophilic anaerobic conditions.
International Journal of Hydrogen Energy, 34(17), 7441-7447.
Barros, A.R., Silva, E.L. 2012. Hydrogen and ethanol production in anaerobic fluidized bed reactors:
Performance evaluation for three support materials under different operating conditions.
Biochemical Engineering Journal, 61, 59-65.
Cappelletti, B.M., Reginatto, V., Amante, E.R., Antonio, R.V. 2011. Fermentative production of hydrogen
from cassava processing wastewater by Clostridium acetobutylicum. Renewable Energy, 36(12),
3367-3372.
Carrillo-Reyes, J., Celis, L.B., Alatriste-Mondragon, F., Razo-Flores, E. 2012. Different start-up strategies
to enhance biohydrogen production from cheese whey in UASB reactors. International Journal of
Hydrogen Energy, 37(7), 5591-5601.
Castello, E., Garcia y Santos, C., Iglesias, T., Paolino, G., Wenzel, J., Borzacconi, L., Etchebehere, C.
2009. Feasibility of biohydrogen production from cheese whey using a UASB reactor: Links
between microbial community and reactor performance. International Journal of Hydrogen Energy,
34(14), 5674-5682.
Davila-Vazquez, G., Berenice Cota-Navarro, C., Manuel Rosales-Colunga, L., de Leon-Rodriguez, A.,
Razo-Flores, E. 2009. Continuous biohydrogen production using cheese whey: Improving the
hydrogen production rate. International Journal of Hydrogen Energy, 34(10), 4296-4304.
Dubois, M., Gilles, K.A., Hamilton, J.K., Rebers, P.A., Smith, F. 1956. Colorimetric method for
determination of sugars and related substances. Analytical Chemistry, 28(3), 350-356.
801
Ferreira Rosa, P.R.; Santos, S.C.; Silva, E.L., 2014. Different ratios of carbon sources in the fermentation
of cheese whey and glucose as substrates for hydrogen production and ethanol production in
continuous reactors. International Journal of Hydrogen Energy, v. 39, p. 1288-1296.
IEA.
International
Energy
Agency.
Key
World
Energy
Statistics
(2013).
Available:
http://www.iea.org/publications/freepublications/publication/kwes.pdf.
Kargi, F.; Eren, N.S.; Ozmihci, S. 2012. Hydrogen gas production from cheese whey powder (CWP)
solution by thermophilic dark fermentation. International Journal of Hydrogen Energy, 37, 22602266.
Kim, D.-H., Han, S.-K., Kim, S.-H., Shin, H.-S. 2006. Effect of gas sparging on continuous fermentative
hydrogen production. International Journal of Hydrogen Energy, 31(15), 2158-2169.
Koskinen, P.E.P.; Kaksonen, A.H.; Puhakka, 2007. L.A. The relationship between instability of H2
production and compositions of bacterial communities within a dark fermentation fluidized-bed
bioreactor. Biotechnology Bioengineering, v. 97, p. 742-758.
Lin, C. N.; Wu, S. Y.; Chang, J. S. 2006. Fermentative hydrogen production with a draft tube fluidized bed
reactor containing silicon-gel-immobilized anaerobic sludge. International Journal of Hydrogen
Energy, v. 31, p. 2200-2210.
Mohan, S.V., Mohanakrishna, G., Sarma, P.N. 2008. Integration of acidogenic and methanogenic
processes for simultaneous production of biohydrogen and methane from wastewater treatment.
International Journal of Hydrogen Energy, 33(9), 2156-2166.
Muños, K.P, Steinmetz, H. 2012.Evaluation of pre-treatment on the first stage of an anaerobic digester for
enhancing bio-hydrogen production and its associated energy balance. Energy Procedia, 29, 469
± 479.
Prazeres, A.R., Carvalho, F., Rivas, J. 2012. Cheese whey management: A review. Journal of
Environmental Management, 110, 48-68.
Rosa, P.R.F.; Santos, S.C.; Sakamoto, I.K.; Varesche, M.B.A; Silva, E.L., 2014. Hydrogen Production
from cheese whey with ethanol-type fermentation: Effect of hydraulic retention time on the microbial
community composition. In: Bioresource Technology, 161, 10-19.
Reis, C.M. dos.; Silva, E.L. 2011. Effect of upflow velocity and hydraulic retention time in anaerobic
fluidized-bed reactors used for hydrogen production Chemical Engineering Journal, v. 172, p. 2836.
Santos, S.C.; Rosa, P.R.F.; Sakamoto, I.K.; Varesche, M.B.A; Silva, E.L., 2014. Continuous thermophilic
hydrogen production and microbial community analysis from anaerobic digestion of diluted sugar
cane stillage. In: International Journal of Hydrogen Energy, 30, 9000-9011.
Shida, G.M., Sader, L.T., Cavalcante de Amorim, E.L., Sakamoto, I.K., Maintinguer, S.I., Saavedra, N.K.,
Amancio Varesche, M.B., Silva, E.L. 2012. Performance and composition of bacterial communities
in anaerobic fluidized bed reactors for hydrogen production: Effects of organic loading rate and
alkalinity. International Journal of Hydrogen Energy, 37(22), 16925-16934.
Searmsirimongkol, P., Rangsunvigit, P., Leethochawalit, M., Chavadej, S. 2011. Hydrogen production
from alcohol distillery wastewater containing high potassium and sulfate using an anaerobic
sequencing batch reactor. International Journal of Hydrogen Energy, 36(20), 12810-12821.
Saady, N.M. 2013. Homoacetogenesis during hydrogen production by mixed cultures dark fermentation:
Unresolved challenge. International journal of hydrogen energy, 38, 13172-13191.
Singh, L., Wahid, Z.A., Siddiqui, M.F., Ahmad, A., Ab Rahim, M.H., Sakinah, M. 2013. Application of
immobilized upflow anaerobic sludge blanket reactor using Clostridium LS2 for enhanced
biohydrogen production and treatment efficiency of palm oil mill effluent. International Journal of
Show, K.Y.; Lee, D.J.; Tay, J.H.; Lin, C.Y. 2012. Biohydrogen production: Current perspectives and the
way forward. International Journal of Hydrogen Energy, v. 37, p. 15616-15631.
Thanwised, P., Wirojanagud, W., Reungsang, A. 2012. Effect of hydraulic retention time on hydrogen
production and chemical oxygen demand removal from tapioca wastewater using anaerobic mixed
cultures in anaerobic baffled reactor (ABR). International Journal of Hydrogen Energy, 37(20),
15503-15510.
Van Ginkel, S., Logan, B.E. 2005. Inhibition of biohydrogen production by undissociated acetic and butyric
acids. Environmental Science & Technology, 39(23), 9351-9356
802
Zhang, Z.-P., TAY, J.-H., Show, K.-Y., Yan, R., Liang, D.T., Lee, D.-J., JIang, W.-J. 2007. Biohydrogen
production in a granular activated carbon anaerobic fluidized bed reactor. International Journal of
803
CHAPTER 7.12. STUDY OF THE EFFECT OF STERILIZATION AND INOCULATION
OF LIGNOCELLULOLYTIC FUNGI FOR TREATMENT OF THE AZO DYE BLACK
TERASYL SR IN PEAT PACKED BIOFILTERS FOR ORGANIC MATTER AND
COLOR REMOVAL IN THE TEXTILE INDUSTRY
M. P. Mejía -López (1); G. E. Moeller ±Chávez* (2)
(1) F.I. -UNAM, Cuernavaca Mor., México
(2) UPEMOR, Environmental Technology Department, Cuernavaca Mor., México
ABSTRACT
Azo dyes are widely used in the textile industry, they are carcinogenic and recalcitrant
molecules. This research is focused on the development of a procedure at bench scale
to treat an azo dye (Black terasyl SR) widely used in the mexican textile industry (thread
production and dyeing), by a biofiltration process using peat as the packing material.
The main objective was to study the effect of inoculation with Pleurotus ostreatus a
lignocellulolytic fungi and the effect of sterilization in the performance of the biofilter to
remove organic matter, color and toxicity. The research consisted in constructing,
packing, operating and evaluating four reactors. One reactor was packed just with peat,
the second one used sterilized peat inoculated with P. ostreatus, the third reactor was
packed with sterilized peat and the last one was packed and inoculated with P. ostreatus
without sterilization. The reactors were operated for five months. Water treated in this
work originates from an effluent having a high concentration of COD (8000 mg/L) and a
concentration of color 566 mg/L. At very high concentrations worked with dilutions. The
highest volumetric organic work load was 624.7 mg/L d this load was used for a short
time due to the rapid decline of the inoculum strain. Thus the load was reduced to 161.7
mg/L d. Results indicated that the biofilter packed with peat with no sterilization had the
best performance: 50% color removal and 56 % organic matter removal with the
complete elimination of toxicity.
Key words: azo dye, biofiltration, P. ostreatus.
Introduction
Many pollutants are generated in the textile industry: phenols, sulfides, chromium and
dyes, among others. A widely used dye type is the azo group. They are used in the
textile, paper, food, cosmetic and pharmaceutical industries. (Bishop & Jiang, 1994).
Azo dyes cause environmental problems mainly to the following reasons:
Their organic structures pass through wastewater treatment systems without being
eliminated or removed and produce highly coloured discharges.
804
They cause inhibition of the biological processes used in the WWT facilities. Dyes
could cause an acute or chronic toxic effect in living organisms exposed to them.
The environmental problematic of highly coloured effluents is due to their
persistence. Sometimes the dye concentration is too low as 1 ppm or less and highly
notorious and toxic.
Biological treatment, both aerobic and anaerobic or combined are considered to be
effective for the removal of toxic compounds and organic matter. Different types of
microorganisms are used for this purpose: fungi and bacteria (Haug et al., 1991; Guo et
al., 2008).
This study is focused in comparing the performance capacity of biofilters packed
with peat (the mixed bacterial population developed in the packing material) compared
to the lignocellulolytic enzymes produced by P. ostreatus inoculated in the packing
material.
Textile industry effluents containing Black Terasyl SR are treated and color
removal and organic matter capacity of biofilters packed and inoculated with Pleurotus
ostreatus is evaluated. P. ostreatus is responsible of the carbon recycling from lignine. It
has non specific enzymes with the potential use to transform pollutants and xenobiotic
compounds. In recent years it has been reported that organic matter and color removal
is effective using inoculated peat packed reactors using a Basidiomycetal fungal type
(Dávila et al., 2003, Moeller & Garzón, 2003).
Some reasons of why this is a feasible technology for organic matter and color
removal are:
Conventional WWTP are not effective in removing this type of pollutants and due to
these, highly colored effluents are discharged in water bodies. As a consequence new
restrictions for discharge have been established in many countries. (Vandevivere et al.,
1998)
Because of the simplicty, efficiency and economy of this technology compared to
conventional systems, it is preferred in many countries (low investment costs, ease of
operation and low maintenance activities (Buelna, 1994).
The experimental system was constituted by four reactors (peat packed biofilters); two of
them were packed with peat and inoculated with P. ostreatus (Table 1). One of them
was previously sterilized and the other was not. Both biofilters were compared in its
performance and operated using two organic loads (625 mg/L d and 162mg/L). Two
additional biofilters were packed just with peat, one was sterilized and the other was not.
The four reactors were fed from the upper part and the effluent was collected in the
lower part of the biofilter. Figure 1 shows a photograph of the biofilter used. Reactor
details as reported by Dávila (2005). The experiment was designed in a way that it was
possible to evaluate the effect of the inoculated peat with P. ostreatus with and without
sterilization. (R2 and R4) and to compare their performance with two other biofilters
without inocula and with and without sterilization (R3 and R1).
TABLE 1. Reactors and packing material characteristics
805
Reactor
1
2
3
4
Identification
R1
R2
R3
R4
Packing material characteristics
Peat (not sterilized)
Sterilized peat inoculated with P. ostreatus
Peat (sterilized)
Non sterilized peat inoculated with P. ostreatus
FIGURE 1. Biofilter
As for the results of COD (Figure 2) for the first organic load (624.7 mg/L d), a decrease
was observed in the COD output of all biofilters. From day 26 to 40 (early days with
second load) increased COD for R2 and R4. After day 40 decrease in effluent COD for
the four biofilters were observed.
In the case of color concentration (Figure 3), where data of the four biofilters are
presented compared to the influent. It was noted that R2, R3 and R4 have similar
behavior as the first load. However R1, is noted, showing low concentrations of dye in
the effluent. For the second load (161 mg/L d), R1, R2 and R4 showed the same
behavior, except R3, that exhibits the higher concentrations of dye in the effluent.
In Figure 2 COD results are shown, where it can be observed that the removal efficiency
for R1 was the best achieved (53.4 % and 56.7% for the two organic loads).
Also, in figure 3 it is shown that R1 achieved the best results for the removal of dye
concentration (24.2 and 50.1% for the two organic loads).
For these results it was found that the best way to remove color is only with unsterilized
peat. R1 mostly removes color and also saves time and material, because it is not
necessary to be inoculated with P. ostreatus. The color removal was carried out by
physical processes in the peat and to the microorganisms attached to it.
806
FIGURE 2. Chemical Oxygen Demand, COD behaviour (Influent, R1, R2, R3 and
R4)
FIGURE 3. Color concentration behaviour (Influent, R1, R2, R3 and R4).
Conclusion
Results are concentrated and shown in Table 2. From these results it can be concluded
that:
The best performance for color, organic matter, NTK, N-NH3, P and toxicity removal was
R1 (biofilter packed with non sterilized peat).
The mixed population developed in R1 achieved the best performance for the biofilters
compared with those that were sterilized and inoculated with P. ostreatus.
R2 and R4 were negatively affected by the presence of enzymes secreted by the fungi
that lead to low color and organic matter removal. Enzymes secreted by P. ostreatus
were not enough to achieve high color removal.
807
It was observed that with the lower organic load a better organic matter and color
removal was achieved.
The sterilization process affected negatively the performance of the reactors. TABLE 2. Reactor performance for color and COD
Apparent color
R1
R2
R3
R4
%
Organic
load
625mg/L
d
31.3
13.2
19.5
19.7
%
Organic
load
162mg/L
d
38.9
30.2
30.1
27.4
Color
COD
concentration
%
%
%
%
%Organic
%Organic
Organic
Organic Organic
Organic
load
load
load
load
load
load
162mg/L
625mg/L
625mg/L
625mg/L 162mg/L
162mg/L
d
d
d
d
d
d
24.7
31.8
24.2
50.1
53.4
56.7
13.4
28.2
12.4
45.4
29.9
41.6
15.4
24.2
15.2
35
31.6
56.8
16.3
26.2
16
46
42.1
47.8
Real Color
References
Bishop, P. and Jiang, H. (1994), Aerobic biodegradation of azo dyes in biofilms. Wat. Sci. Tech. 29, 525230.
%XHOQD*³/DELRILOWUDFLyQHQOHFKRRUJiQLFRXQDQRYHGRVDDOWHUQDWLYDSDUDHOWUDWDPLHQWRGHHIOXHQWHV
industriales \ GH SHTXHxRV PXQLFLSLRV´ LQ VSDQLVK ,, &RQJUHVR \ H[SRVLFLyQ LQWHUQDFLRQDO GH OD
industria del medio ambiente. México±Canadá±Estados Unidos, México D.F., (1994)
Dávila, S. V. (2005). Degradación Fúngica del Colorante Naranja ácido 24 in spanish. Tesis de Maestría
en Ingeniería, Facultad de Ingeniería, UNAM, México, DF, México, 114pp.
Dávila, S. V., Moeller, Ch. G., Garzón, Z. M. A. y Farfan, G. Fungal colour reduction of azo dyes in
wastewater. Proceedings of the IWA Conference on Enviromental Biotechnology. Kuala Lumpur,
Malasia, (2003).
Guo, M., Lu F., Liu, M., Li T., Pu, J., Wan N., Liang P. Zhang C. (2008). Purification of recombinant
laccase from Trametes versicolor in Pichia mehanolica and its use for the decolorization of
anthraquinone dye. Biotechnol Lett. DOI 10. 1007/s 10529-008-9817-z.
Haug W. Schmidt, A., Nortermann, B., Hempel, D., Stolz, A. y Knackmuss H. (1991). Mineralization of
sulfonated azo dye Mordant Yellow 3 by a 6 aminonaphthalene-2-sulfonate-degrading bacterial
consortium Appl. Environ. Microbiol. 57, 3144-3149.
Moeller, Ch. G. & Garzón, Z. M. A. (2003). Desarrollo de tecnologías no convencionales para el
tratamiento de efluentes de la fabricación de colorantes tipo azo.in spanish. Proyecto Interno
Instituto Mexicano de Tecnología del Agua, IMTA. Mexico.
Vandevivere P, Verstraete W. (1998). Treatment and reuse of wastewater from the textile industry: review
of emerging technologies. Journal of chemical technology and biotechnology. 72, 289-302
Notation
COD
P. ostreatus
Chemical Oxigen Demand
Pleurotus ostreatus
808
WWT
NTK
N-NH3
P
WWTP
Wastewater treatment
Kjeldahl nitrogen
Ammonia nitrogen
Phosphorus
Wastewater treatment process
809
CHAPTER 7.13. EFFECT OF THE COSOLVENT IN THE DEGRADATION OF
NAPHTHALENE BY CATALYTIC OZONATION
Claudia M. Aguilar-Melo* (1); Jorge I. Chairez-Oria (1); Julia L. Rodríguez-Santillán (2)
Tatiana Poznyak (2)
(1) Unidad Profesional Interdisciplinaria de Biotecnología ± IPN. México D.F., México
(2) Escuela Superior de Ingeniería Química e Industrias Extractivas - IPN, México DF, México
ABSTRACT
Naphthalene (NA) is a polycyclic aromatic hydrocarbon that has been listed as a priority
pollutant by Environmental Protection Agency of United States (US EPA). NA toxic
properties in aquatic systems increase the removal difficulty by conventional methods
(for example, the microbiological effect). Ozonation and catalytic ozonation (O 3-cat)
processes are sound alternatives of degradation for this compound. However, the
existence of certain compounds in water, such as: phosphates, carbonates, sulphates,
surfactants and co-solvents (in real systems, the chemicals increase the pollutants
concentration in water bodies) can be detrimental to the efficiency of O 3 and O3-cat
systems. In this work, the NA (20 mg L-1) degradation by conventional and catalytic
ozonation in the presence of a co-solvent (ethanol) was studied. This solution was
proposed to simulate some aspects of real wastewaters where not only water acts as
solvent. Two proportion of the mixture water:ethanol were selected 70:30 and 50:50 with
the purpose of studying the co-solvent effect on NA degradation system by ozonation.
O3-cat process was used with nickel oxide as catalyst (100 mg L -1). The degradation
analysis of NA by O3-cat in two different proportions of co-solvent showed that in the
case of 70:30 (water:ethanol) a 95% of NA degradation in 60 min was obtained, while in
the case 50:50 (water:ethanol), only 55% of NA degradation was achieved. The O 3
process showed similar results of degradation to the initial compound in comparison with
catalytic system. According to these degradation results, there is an inhibition effect in
pollutant removal by ethanol. The higher ethanol concentration, the lower elimination
rate of NA was obtained (by 35% during the 60 min). UV-Vis analysis was used to
evaluate the effect of co-solvent and catalyst presence over the byproducts dynamics.
Even when the NA was almost not affected by these factors, byproducts were
significatively modified by the presence of co-solvent and catalyst.
Key words: catalytic ozonation (O3-cat), co-solvent (ethanol), naphthalene (NA), nickel
oxide, ozonation.
810
Introduction
Naphthalene (NA) is the smallest aromatic polycyclic hydrocarbon with only two fused
benzene rings. This hydrocarbon has industrial and commercial applications as active
ingredient in mothballs, fabrication of carbaryl pesticides, naphthylamines, anthranilic,
phthalic acids, synthetic resins, and so forth.Besides, this organic is the major aromatic
component in crude oil, petrol and creosotes (McConkey, et al., 2002). As a result of NA
industrial applications, it can be found in air, water and soil. The toxicity of NA compound
and its by-products has been proved by several studies. In humans, acute exposition to
NA causes liver and kidney damage, cataracts formation, changes in haemoglobin
oxidation, neurological and DNA damage, among other affections (Bagchi et al., 1998;
Stohs et al., 2002; Bagchi et al., 1998). The US EPA has categorized the NA as a
priority pollutant to be eliminated from the environment due to its carcinogenic and
teratogenic activity (US EPA, 2003).
Research in degradation of NA has been carried out under several systems
including photolysis (McConkey et al., 2002; Bertilsson et al., 2002), biological systems
(Annweiler et al., 2002), photocatalysis (Lair et al., 2008; Woo et al., 2009) and
ozonation (Zhu et al., 1999). Ozonation is one of the techniques with the shortest
periods of degradation (hours) and high degradation capacity. Therefore it has received
a lot of attention to decompose complex and toxic contaminants (Beltrán et al., 2004).
Ozone possess a high oxidation potential (2.07 eV) and it can act by either, a direct
reaction with contaminant molecules or an indirect reaction via generation of hydroxyl
radicals (OHÚ) derived from ozone decomposition with superior oxidation potentials
(2.80 eV). Both mechanisms occur simultaneously while reacting with pollutants,
nevertheless either via can prevail manipulating the system conditions. For instance, any
increment in pH rises ozone decomposition increasing OHÚ presence. An important
limitation of ozonation process is the transformation of organic matter to aldehydes and
organic acids which remain unchanged during the treatment preventing complete
mineralization (Poznyak et al., 2007). Additionally, the reaction constants of ozone are
not always high (Beltrán et al., 2004; Nawrocki et al., 2010).
In order to overcome the aforementioned ozonation disadvantages, the use of
catalyst has been widespread for increasing the reaction rate and the mineralization
degree in a process denominated catalytic ozonation (O3-cat) (Nawrocki et al., 2010). In
the O3-cat systems, the increment of contaminants mineralization is predominantly
product of OH· reactions resulted from ozone decomposition (Rodríguez et al., 2012).
Such systems are divided in homogenous and heterogeneous according to the
characteristics of the catalyst in the solution. In heterogeneous catalysis, a solid catalyst
like a metallic oxide, immobilized metal, metal doped, zeolite, or activated carbon is an
active agent in the reaction system (Nawrocki et al., 2010). Among the metallic oxides
used as catalyst, TiO2, MnO2 and Al2O3 are the most common compounds. However, in
studies made by Rodríguez et al., (2013) over degradation of 2-4 Diclorophenoxyacetic
acid (compound highly resistant to degradation) with different metallic oxides (TiO 2,
SiO2, Al2O3 and NiO), it was demonstrated that the higher mineralization was obtained
by NiO.
On the other hand, the presence of certain compounds in water, such as:
phosphates, carbonates, sulphates, surfactants and co-solvents (in real systems, such
811
chemicals increase the pollutants concentration in water bodies) can be detrimental to
the efficiency of O3 and O3-cat systems.
In this work, the degradation by conventional and catalytic ozonation of the NA (20
mg L-1) in the presence of a co-solvent (ethanol) was studied. This systems was
proposed in order to simulate some aspects of real wastewaters where not only water
acts as solvent. Two proportions of the mixture water:ethanol were selected (70:30 and
50:50) with the purpose of studying the effect of co-solvent in the degradation system of
NA by O3 process. NiO was selected as catalyst by its effective degradation of complex
organics showed in previous studies. This study analysed the NA decomposition as well
as a preliminary examination on the byproducts evolution.
All the reagents used in all the ozonation experiments were analytic degree (J.T baker).
Absolute ethylic alcohol was used as co-solvent (Chemical Meyer, 99.5%). Nickel (II)
oxide catalyst (sigma-aldrich) was used in the nanoparticles form with diameters smaller
than 50 nm (99.8%). The reagents used in the analysis of NA degradation were HPLC
grade (Acetonitrile by Mallinckrodt).
A solution of NA (20mg L-1) was prepared using as solvent two mixtures of water:
ethanol (70:30 and 50:50). The diluents preparation was made with distillate water. NA
crystals were mixed in the eluent preparation through agitation during 24 hours at
ambient temperature and protected from light. The final pH of the sample was 5.6.
Ozonation. Ozone was generated from dry oxygen (type corona of discharge) in an
equipment HTU500G (AZCO Industries Limited ± Canada) with a maximum flux O2 - O3
of 500 mL min-1. The ozonation process was carried in batch at ambient temperature in
a 500 mL glass reactor.
The reactor has an inlet of air in the bottom were the mixture O 2 ± O3 is distributed
through a diffusor in a ceramic plaque. At the top of the reactor, there is an outlet for the
gas mixture which is monitored by a detector of ozone in the gas phase (Ozone analyzer
BMT 963 with UV detector) and sent to a computer. The flux diagram shows the
ozonation system (Figure 1).
Ozonation was carried out in 400 mL of the NA model solution (20 mg L -1) in the
mixtures of water:ethanol 70:30 and 50:50. The reaction was followed at ambient
temperature, without pH control, under a maximum ozone flux of 5.5 ± 0.5 mg L-1 during
60 min. The catalytic ozonation systems were made under the conditions described
above. The concentration of Nickel oxide was 100 mg L -1. During the treatment, aliquots
of 3.0 mL were taken at different time intervals for analytical studies. All the experiments
were made by triplicate.
Analytical methods.The NA degradation was monitored by UV-Vis spectroscopy in a
Perkin Elmer Lambda 25 equipment. The analysis was made in the range of 200 to 400
nm. The samples belonged to the systems O3-cat were filtered before entering the
equipment with membranes of 0.2 µm.
The degradation of NA was followed by a High-Performance Liquid Cromatograph
(HPLC) PerkinElmer Flexar with a Diode Array Detector (DAD). The column was a
Platinum C18 Altech 250 mm x 4.6 mm, mobile phase water:acetonitrile 30:70, flux 0.3
812
mL min-1 and wavelength of de 210 nm. The samples were filtered in membranes of 0.2
µm.
V-‐3
I-‐4
V-‐1
P2
V-‐2
V-‐4
Ozone Generator
Bot
0.5
ón
Ozone Analyzer
Computer
Reactor
Oxygen
FIGURE 1. Flux diagram of the ozonation process.
Results and discussionN
The HPLC analysis of the ozonated (conventional and catalytic) samples showed an
inhibitory effect in NA degradation when the ethanol concentration was increased. The
effect can be observed in the HPLC normalized area graphics of NA degradation (Figure
2). The treatments with 30% of co-solvent reached a 95% degradation of NA while the
ones with 50% of ethanol had a reduction of 55%.
2a
2b
FIGURE 2. Results of NA degradation follow by HPLC (A/A0). a) O3 treatment. b) O3
-cat.
70:30 y
50:50 water:ethanol.
The rate of degradation was also affected by the presence of co-solvent. In
treatments with a minor concentration of co-solvent, the relation A/A0 had a fast rate of
reduction between the 5 and 30 min after which the curve slope diminished. Treatments
with 50% of co-solvent demonstrated a degradation rate which was slowdown and the
813
curve slope was unaffected.
A similar change in degradation rates has been reported in ozone. However,
remarkable differences have being notorious in the degradation of ozonation byproducts
which are less reactive to ozone with degradation rates smaller than the initial
compounds. The UV-Vis analysis (Figure 4 and 5) shown the disappearance of some
characteristic bands after 30 min.
A more relevant change in degradation rates was a phenomenon only observed
when the smaller co-solvent concentration was used. There is scarce information related
to the presence of co-solvents in the ozonation process and O3-cat. The ozone
decomposition in aqueous media in the presence of organic acids was reported by
Staehelln (1985). That study describes the effect of formic acid, methanol, t-butanol,
bicarbonate as well as ions carbonate and phosphate. The primary alcohol methanol
was described as a promoter in the conversion of OHÚ to the superoxide anion (.O2-),
which is a weaker oxidant than the precursor (Staehelln et al., 1985). The conversion
could be one of the causes for the degradation rate reduction in treatments with higher
concentrations of ethanol. The ozonograms obtained from the treatments with O 3 and
O3-cat (Figure 3) shown a difference in the ozone concentration at the outlet of the
reactor. The greater the the solvent concentration, the bigger degradation of the
molecule of ozone. The outlet ozone concentration remains constant after 60 min of the
treatment.
FIGURE 3. 7KUHDWPHQWV¶R]RQRJUDPVLQHOXHQWVwater:ethanol 70:30 and 50:50.
In addition, in the catalysed treatments, the degradation rates of NA remain
unchanged for both co-solvents concentrations (Figure 2a and 2b). The inefficiency
effect of O3-cat has been reported before by Al-Hayek et al., (1989) in a system of O3,
O3+Al2O3, and O3+Fe/Al2O3 for phenol degradation. The results shown that the addition
of Al2O3 and Fe/Al2O3 did not improve phenol degradation, whilst the total organic
814
carbon (TOC) shown an important reduction as a result of the catalytic action (O 3: 40%,
O3+Al2O3: 70%, and O3+Fe/Al2O3: 90%). Therefore, the effect of catalyst in not always
observed in the initial compound but in the byproducts and mineralization system.
The UV-Vis analysis of the samples allowed observing a general effect of cosolvent and catalyst during treatments. These analyses were useful to get an indirect
evaluation of the possible presence of byproducts (as a result of absorbance analysis
under certain wavelength intervals). Figure 4 shows the spectrum of initial solutions of
NA with 30 and 50% ethanol/water proportion. A signal observed in the range of 250 to
290 nm is a distinctive spectrum of aromatic groups. The increment of absorbance in
samples with higher concentration of co-solvent may be a consequence of UV radiation
absorption by ethanol.
FIGURE 4. UV-Vis spectra of NA in the diluent 70:30 (line in black) and 50:50 (line
in blue) water: ethanol.
Effect of reaction time over the byproducts decomposition. The samples analyzed
by UV-Vis in treatments with 30% of ethanol treated by either conventional processes as
well as O3-cat were evaluated at two different reaction times. The increment of reaction
time (Figure 5a and 5b) decreased the UV/Vis signal measured in the range of 250 to
290 nm (range in which aromatic compounds are located). When the reaction time was
1800 seconds (Figure 5b), the absorbance in this range is near to 1 and it is not
detected characteristics bands of NA. Therefore, one can expect the byproducts
accumulation
with
aromatic
structures
as
Cinnamaldehyde,
1,2Benzenedicarboxaldehyde, 2-Formylcinnamaldehyde. These compounds have been
reported as regular byproducts from the breakdown of the NA molecule, which still
retains an aromatic ring. This characteristic can increase the absorbance signal intensity
at these wavelengths.
The spectra obtained after 3600 seconds of treatment in the presence of 30%
ethanol (Figure 5b) clearly showed an absorbance decrease at the characteristic
wavelengths of aromatic compounds. Furthermore, it is expected that byproducts
815
obtained from the NA degradation retains aromatic rings form the initial compound. By
comparing the absorbance in the range of 210 to 230 nm for both treatment times
(Figure 5a and 5b) clear differences in the absorption spectrum can be detected.
Absorbance signal intensity is less when the reaction time is 3600 seconds. This result
may indicate a short chain byproducts decomposition such as organic acids.
When the treatment was evaluated using the 50:50 water:ethanol solvent, the
absorption spectra (Figure 5c and 5d), showed relevant differences in comparison with
treatments executed when the solvent was 70:30 water ethanol. In samples
characterized after 1800 seconds of reaction, it is still visible the UV-Vis absorbance
signal which is characteristic of NA. Moreover, this band remains detectable even after
3600 seconds of reaction. This fact is consistent with the dynamics of degradation
obtained by HPLC, where NA degradation is not completed at the end of treatment.
However, absorbance intensity at these wavelengths can be a consequence of some
byproducts such as 1-naphthol, 2-naphthol 1-4 naphthoquinone, 5-hydroxy-1,4naphthoquinone. These compounds have been reported as intermediates in the NA
degradation under different treatment methods. All these compounds have two rings:
one aromatic and the other one affected by the presence of substituents (Lair et al,
2008).
The spectra for both treatment times (1800 and 3600) at wavelengths between 210
to 230 nm, showed no difference. Therefore, the increment in the process time for
systems with higher co-solvent concentration, did not yield to the generation or
degradation of byproducts such as organic acids which are recalcitrant structures in
ozonation processes.
Effect of catalyst presence over the byproducts decomposition. The addition of NiO
in the degradation processes of NA under the presence of co-solvent, did not show an
increment in the rate of degradation of the initial compound (Figure 2). However, the UVVis analysis of samples obtained in system with 30% ethanol demonstrated significant
differences between the conventional and catalytic systems (Figure 5a and 5b).
Treatments with the catalyst presented minor UV-Vis signal intensity compared with the
conventional system in wavelengths near to 240 nm. These signals are resulting from
the byproducts generation.
Similarly, the absorbance measure in the range of 200 and 220 nm is lower when
the catalyst is used. This condition seems to indicate a minor presence of organic acids.
Therefore, the addition of catalyst in treatments with 30% ethanol, has relevant effect on
byproduct formation.
The use of the catalyst in treatments with 50% co-solvent (Figure 5c and 5b)
expose differences in the spectrum obtained from the conventional system. In this
system, a relevant difference is observed between the spectrum signals measure at
wavelengths below 250 nm. So, it is likely that also ethanol interfered in the byproduct
degradation process under the presence of the catalyst. This analysis also showed that
catalyst addition did not increment the degradation rate of NA but it has an effect on the
byproducts degradation.
Effect of co-solvent concentration over byproducts decomposition. The increment
of the ethanol concentration from 30 to 50% (Figure 5b and 5d) evidenced an inhibition
effect over ozonation process. Treatment with 30% ethanol after 3600 seconds (Figure
816
5b) showed that UV-Vis absorbance in the range of 250 and 290 nm is less than 0.5. On
the contrary, the systems with a higher concentration of co-solvent (Figure 5d), showed
an absorbance near to 1.0. Therefore, one can expect the presence of compounds with
aromatic nature.
Treatments with higher ethanol exhibited intense signals in the region from 200 to
230 nm. This fact demonstrated, that ozonation process was less efficient when the
ethanol concentration is increased. This result is consistent with the results obtained in
the NA degradation obtained by HPLC.
5a
5b
5c
5d
FIGURE 5. Analysis UV-Vis of the NA degradation treatments. a) 70:30
water:ethanol, 1800 seconds of ozonation. b) 70:30 water:ethanol, 3600 seconds of
ozonation. c) 50:50 water:ethanol, 1800 seconds of ozonation. c) 50:50 water:ethanol,
3600 seconds of ozonation. O3 treatment (line in black). O3 ±cat (line in red).
Conclusion
The treatments of O3 and O3-cat allowed the degradation of NA in the presence of a cosolvent (ethanol). The concentration increament of this co-solvent inhibited the NA
degradation (treatments with 30% of ethanol had a 95% of NA degradation while the
treatment with 50% of ethanol had only a 55% of NA degradation). The addition of a
catalyst made no difference in the degradation rates of NA, but the UV-Vis spectrum
showed important differences after 1800 and 3600 second of reaction. Thetrefore, the
817
presence of catalyst showed differences in the byproducts formation if the co-solvent
was 30% but these differences were almost imperceptibles when the co-solvent was
increment to 50%. Therefore, the presence of co-solvent in higher concentrations (50%
versus 30%) has a negative effect on the NA ozonation byproducts decomposition.
However, this effect was not observed when the NA decomposition was evaluated.
Acknowledgements
The authors acknowledge Consejo Nacional de Ciencia y Tecnología and Instituto
Politecnico Nacional support number 20141483 provided for the development of this
investigation.
References
AlǦHayek, N.; Legube, B.; Doré, M. (1989). Ozonation catalytique (Fe III/A12O3) Du phénol et de ses
produits d'ozonatio. Environmental Technology Letters. 10(4): 415-426.
Annweiler E., Richnow H.H., Antranikian G., Hebenbrock S., Garms C., Franke S., Francke W., Michaelis
W. (2000). Naphthalene Degradation and Incorporation of Naphthalene- Derived Carbon into
Biomass by the Thermophile Bacillus thermoleovorans. Applied and environmental microbiology.
66(2): 518-523.
Bagchi M., Bagchi D., Balmoori J., Ye X., Stohs S.J. (1998). Naphthalene-induced oxidative stress and
DNA damage incultured macrophage J774A.1 cells. Free Radical Biology & Medicine. 25 (2): 137143.
Bagchi, D.; Bagchi, M.; Balmoori, J.; Vuchetich, P.J.; Stohs, S.J. (1998). Induction the oxidative stress and
DNA damage by chronic administration of naphthalene in rats. Res Commun Mol Pathol
Pharmacol. 101(3):249-257.
Beltrán F. (2004). Ozone reaction kinetics for water and wastewater systems. Washington D.C.: Lewis
publishers.
Bertilsson S., Widenfalk A. (2002). Photochemical degradation of PAHs in freshwaters and their impact on
bacterial growth ± LQÀXHQFHRIZDWHUFKHPLVWU\+\GURELRORJLD±32.
Lair A., Ferronato C., Chovelon J-M., Herrmann J-M. (2008). Napthlene degradation in water by
heterogeneous photocatalysis: An investigation of the influence of inorganic anions. Journal of
photochemistry and photobiology A: Chemistry. 193: 193-203
McConkey, B.J.; Hewitt, L.M.; Dixon, D.G. y Greenberg, B.M. (2002). Natural sunlight induced
photooxidation of naphthalene in aqueous solution. Water, air, and Soil Pollution. 136: 347- 359.
Nawrocki, J.; Kasprzyk-Hordern, B.; (2010). The effiency and mechanims of catalytic ozonation. Applied
catalysis B: Environmental. 99: 27- 42.
Poznyak, T.; Colindres, P.; Chairez, I. (2007) Treatment of Textile Industrial Dyes by Simple Ozonation
with Water Recirculation. Journal of the Mexican Chemical Society. 51(2): 81-86.
Rodríguez, J.L.; Valenzuela, M.A.; Pola, F.; Tiznado, H.; Poznyak, T. (2012). Photodeposition of Ni
nanoparticles on TiO2 and their application in the catalytic ozonation of 2,4-dichlorophenoxyacetic
acid. Journal of Molecular Catalysis A: Chemical 353± 354: 29± 36.
Rodríguez, J.L.; Valenzuela, M.A.; Poznyak, T.; Lartundo, L.; Chairez, I. (2013). Reactivity of NiO FOR 24-D degradation with ozone: XPS studies. Journal of hazardous materials. 262:472-481.
Staehelln, J. y Hoigne, J. (1985). Decomposition of Ozone in Water in the Presence of Organic Solutes
Acting as Promoters and Inhibitors of Radical Chain Reactions. Environ. Sci. Technol. 19(12): 12071213.
Stohs S.J., Ohia S., Bagchi D. (2002). Naphthalene toxicity and antioxidant nutients. Toxiclogy 180 (1):
97 ± 105.
U.S. Environmental Protection Agency. (2003). Contaminant Candidate List Regulatory Determination
Support Document for Naphthalene.
U.S. Environmental Protection Agency. (2003). Health Effects Support Document for Naphthalene.
818
Woo O. T., Chung W. K., Wong K.H., Chow Alex T., Wong P.K. (2009). Photocatalytic oxidation of
polycyclic aromatic hydrocarbons: Intermediates identification and toxicity testing. Journal of
Hazardous Materials. 168 (2±3): 1192±1199
Zhu, S; Zhang, Q.; Wang, L.; Chen, J.; Lian, H. (1999). Effect of ozonation of naphthalene derivatives on
their elimination, TOC, and biodegradability. Environmental contamination and toxiclogy. 63: 101108
NOTATION
HPLC
NA
O3-cat
OHÚ
.O 2
PHA
US EPA
High-Performance Liquid Cromatograph
Naphthalene
Catalytic ozonation
Hydroxyl radicals
Superoxide anion
Polycyclic aromatic hydrocarbon
Environmental Protection Agency of United States
819
CHAPTER 7.14. APLICATION OF AN ELECTROOXIDATION PROCESS USING PbO2
ANODES FOR THE ATRAZINE REMOVAL
Juan Napoles-Armenta (1); Celestino García-Gómez (1); Pablo GortáresMoroyoqui*(1)
(1) ITSON, Biotecnología y Ciencias Alimentarias, Ciudad Obregón, México.
ABSTRACT
In recent years, concern has arisen from the contamination of water bodies by some
contaminants found in low concentrations (ng/L or µg/L). Today it is usual to speak
about emerging contaminants such as pharmaceuticals, personal care products,
surfactants, industrial additives, plasticizers, pesticides and others chemical compounds.
Atrazine is an herbicide of interest because of in aquatic species can cause feminization
and even toxicity. The objective of this work was to assess the effect of current intensity
(I), electrolysis time (t), and hydraulic recirculation time (HRT) on the removal of atrazine
from water. The initial concentration in each treatment was 5 mg/L. A cylindrical reactor
with concentric electrodes was used. The electrooxidation processes was investigated
using Ti/PbO2 (anode) and Ti (cathode) circular electrodes in the presence of NaSO 4 as
electrolyte. Electrodes were distributed, from center to periphery, (cathode-anodecathode). A factorial 23 experimental design with a composite central experimental
design were carried out to evaluate the effect of I, t, and HRT on the atrazine removal.
Five levels were studied for each variable (I, t, and HRT) with a total of 20 different
experimental conditions or treatments. Samples were analyzed using UV
spectrophotometry at 223nm. Data analysis was performed using the Design Expert
program 9.Micropollutant removal was affected by for all variables studied, showing
significant difference each one of them. The longer time the greater the atrazine
removed (F0.0001). Also it is directly affected by current intensity (F0.0001), which
implied a positive effect. In contrast, HRT showed a negative effect on the atrazine
removal (F0.0075). The best removal was 68.04%, which was achieved with 2A of I, 90
minutes of t, and 3.87minutes of HRT. This process seems to be a promising technology
that can be used as tertiary treatment to remove contaminants emerging the water.
Key words: atrazine, electrooxidation, micropollutant, treatment, water
-----------*Author for all correspondence
820
Introduction
With the rapid increase in population as well as industrialization, the demand and over
exploitation of natural resources have increased and thus the generation of urban,
agricultural, industrial solid waste, among others, causing various disturbances in the
environment.
Water is a resource vital for the development of the daily activities of mankind,
then, its pollution has become of great interest around the world. Lakes, rivers, canals
and other water bodies are contaminated by anthropogenic activity or natural processes.
Behbahani et al. (2010) reported that the discharge of industrial waste water, such as
semiconductor, aluminum, and glass manufacturing industries contribute to water
pollution, especially in groundwater.
The development of the industry have led to the production of numerous synthetic
chemicals, this has led to an increase in the number of contaminants to be considered
potentially threatening to the environment. Today, it is common to speak about emerging
contaminants such as pharmaceuticals, personal care products, surfactants, industrial
additives, plasticizers, pesticides and a variety of chemical compounds. These
compounds are usually found in small amounts (mg/L or ng/L), which are generally
unregulated or regulated by most countries. They are also called micropollutants due to
concentrations found in water bodies. These turn out to be of great scientific interest due
to their emissions in the environment can increase the presence of resistant bacteria,
also due to their physicochemical properties (high water solubility and low
biodegradability) are able to reach any natural environment and pose a serious risk to
the drinking water. Micropollutants cause significant effects such as disrupting the
endocrine system, blocking hormonal functions, and affecting the health of humans and
animal species even when they are present in low concentrations (Garcia et al., 2011).
Maternal exposure to atrazine in drinking water has been associated with low birth
weight and fetal heart defects. In addition atrazine is an herbicide of interest because of
in aquatic species can cause feminization and even toxicity.
The treatment of wastewater generated by the industry is still a major problem of
environmental pollution due to its enormous amount and diversity. Some wastewater
may have extremely high chemical oxygen demand (COD), on average about
2000mg/L, being strong odor and dark color. Therefore, proper treatment of these
effluents is essential before being discharged into water bodies (Zodi et al., 2010). The
main processes for removing contaminants from wastewater are physical, chemical and
biological processes that occur in conventional processing methods, where pollutants
are removed. However, they are not efficient to remove micropollutants. Many
researchers have sought methods suitable for wastewater treatment. Electrochemical
technologies have reached a promising stage of development and now also can be used
effectively for the removal of these compounds (Farhadi et al., 2012.). Furthermore,
Zaroual et al. (2008) reported that in recent years, the electrocoagulation has been
successfully used to treat a variety of industrial wastewater. Aleboyeh et al. (2007)
mentioned that electro coagulation process has been applied to treat a variety of dye
effluents. IWA Publishing (2010) describes processes for removing trace contaminants
from water and waste water, among which is the microfiltration, adsorption, coagulation,
membrane and biological processes. Compared with traditional flocculation and
821
coagulation, electrocoagulation, in theory, has advantages in the removal of small
colloidal particles; they have a higher probability of being coagulated by the electric field
than set them in motion. Electrooxidation is a process that degrades pollutants without
forming other waste contaminants such as blood clots or flocs. Electrooxidation requires
a supporting electrolyte to enhance electrical conductivity. Although, sodium chloride is
the most commonly used for electrolytic processes, its use remains controversial due to
the possible formation of organic chlorine by-products. Recent work has shown that
sodium sulfate is the best supporting electrolyte for the electrochemical process
considering economic, efficiency, and environmental aspects (Garcia et al., 2011).
When an investigation is conducted is important to give certainty to the results, for
which there are statistical packages helping with this. Garcia et al. (2011) used a central
composite design for the analysis of results, which consists of 2 k factorial matrix with
additional experiments. This is used to determine the optimum operating conditions. This
design includes uniformly distributed points in the space of encoded variables. One
advantage is the ability to explore the entire experimental region and the usefulness of
response interpolation. The matrix allows the description of a region around an optimal
response. Because of the importance of emerging contaminants and the need to
remove them in aquatic systems, the target for this research was to evaluate the effect
of current, time, and time recirculating hydraulic on the removal of atrazine from
wastewater.
Shyntetic wastewater solution. Atrazine analytical quality was used, to prepare a
standard calibration curve. An atrazine solution stock (20 mg/L) was prepared using
distilled water. Atrazine was solubilized by stirring for 24 hours at room temperature.
This solution was kept in refrigeration at 4°C. Subsequently, a volume of 250 ml of
atrazine stock solution was gauged with one litter of distilled water. Then it was kept
under stirring for 10 minutes. The resulting mixture with 5mg/l was use to carried out all
the experiments. Sodium sulphate was used as electrolyte support (400mg/L) to
increase electrical conductivity.
Experimental unit. A cylindrical electrooxidation reactor was manufactured using acrylic
material with 4 cm of radius and 29 cm of height, with one litter of total volume. The
trading volume was 550ml. Two circular titanium mesh electrodes were used as cathode
and a circular titanium mesh electrode lead dioxide coated was used as the anode. All
electrodes were concentrically arrangement with interposed the center to the periphery
(cathode-anode-cathode) to maximize conductivity efficiency. The spacing between
electrodes was 1 cm. HRT was maintained constant with a peristaltic pump at various
flow rates. The current intensity was adjusted using a power supply with a maximum
output of three amperes.
Studied variables. It was sampled at begin and the end of each treatment. Atrazine
concentration in the solution was determined by UV spectrophotometry at 223nm.
Atrazine maximum absorption was obtained by scanning with the spectrophotometer,
the resulting wavelength of 223nm and it was used to evaluate the residual atrazine
concentrations. A calibration curve of known concentration (0.5 to 5mg/L) versus relative
822
absorbance was used to calculate the concentration of the atrazine and to estimate
removal efficiency. A preliminary experiment using 23 factorial design was performed
considering time (t) current intensity (I) and hydraulic recirculation time (HRT) to
evaluate their effect on atrazine removal. In the case of the current intensity the low
value was 1 and the up value was 2 Amperes; in the time the values were 150 minutes
and 300 minutes, and the HRT was set at 1.83 minutes and 2.37 minutes for low and up
values, respectively. A total of 8 different experiments resulted using Design Expert
program 9. In addition a central composite design was considered resulting six central
values and six axial values, with a total of 20 experiments. In total 5 levels were
evaluated for each variable (table 1).
TABLE 1. Levels of the variables evaluated.
Table 2 shows the variable combination for each of the 20 experiments. Initial and final
concentrations were obtained for each experiment and removal efficiency was
calculated according to the following equation:
%R= C0 ± Cf X 100
C0
Equation (1)
% R: Percentage of atrazine removed.
C0 : Initial concentration of atrazine.
Cf : Final concentration of atrazine
TABLE 2. Matrix treatments for removal of atrazine.
823
Results obtained in this investigation can be seen in Table 3. This shows the effect of
current, time and HRT on the removal of atrazine. The last column shows the power
consumption of each experiment.
TABLE 3. Removal of atrazine and energy consumption for treatment.
824
Results were analyzed with the program Design Expert 9. Optimal value for the
studied variables maximizing the percentage removal of atrazine and minimizing energy
consumption was obtained. The optimum value obtained is presented in Table 4.
TABLE 4. Optimum value.
According to Pareto analysis (Figure 1) the three studied variables have a
significant effect on the atrazine removal. Time and current intensity have a direct effect
on the removal, in other words the higher the time and current intensity the higher the
removal. Time present higher effect than current intensity; whereas HRT has an indirect
effect, this is that when HRT increases the atrazine removal decreases. The Pareto
chart showing the analysis of the effect of variables on the removal of atrazine was
observed.
825
FIGURE 1. Importance of variables for the experimental design.
Also the effect of time and current intensity on the atrazine removals can be
observed using the response surface plot (Figure 2). In addition significant effect of t and
I are demonstrated for ANOVA analysis (F = 0001). In the case of HRT the indirect
effect on atrazine removal is significant (F= 0.0075).
FIGURE 2. Effect of the current intensity and time on the removal of atrazine.
Conclusion
Atrazine removal is affected by the current, time and HRT showing significant difference.
The best removal was 54.50%, which was achieved with 2A of current intensity (I),
826
54.04minutes of time (t), and 2.54 minutes of HRT.The results obtained showed that the
electrolysis time and current intensity have direct influence on the atrazine removal.
While the effect of the hydraulic recirculation time, is inversely proportional to the
removal. All variables showed a statistically significant effect. This process seems to be
a promising technology that could be used as tertiary treatment to remove emerging
contaminants from the water.
References
Baghchehsaraee, B.; Nakhla, G.; Karamanev, D.; Margaritisk, A. (2009). Effect of extrinsic lactic acid on
fermentative hydrogen production. International Journal of Hydrogen Energy. 34: 2473-2579.
Bocchini, D.A.; Alves-Prado, H.F.; Baida, L.C.; Roberto, I.C.; Gomes, E.; Silva, R.D. (2002). Optimization
of xylanase production by Bacillus circulans D1 in submerged fermentation using response surface
methodology. Process Biochemistry. 38: 727-731.
Aleboyeh, N. Daneshvar, M.B. Kasiri. 2007. Optimization of C.I. Acid Red 14 azo dye removal by
electrocoagulation batch process with response surface methodology. Chemical Engineering and
Processing, 47; 827±832.
S. Bonora, A. Pisi, S. Ottani, D. Cesini, A. Maris, M. Di Foggia, Raman and SERS study on ibuprofen
metal
complexes
with
biomedical
interest,
Vibrational
Spectroscopy
(2014),
http://dx.doi.org/10.1016/j.vibspec.2014.04.009
García-Gómez C., Gortáres-Moroyoqui P. y Drogui P. 2011. Contaminantes emergentes: efectos y
tratamientos de remoción. Química Viva, pp. 96±105.
IWA Publishing. 2010. Treatment of Micropollutants in Water and Wastewater. Editado por Jurate
Virkutyte, Rajender S. Varma and Veeriah Jegatheesan. Londres.
M. Behbahani, M.R. Alavi Moghaddam, M. Arami. 2010. Techno-economical evaluation of fluoride
removal by electrocoagulation process: Optimization through response surface methodology.
Desalination, 271; 209±218.
S. Farhadi, B. Aminzadeh, A. Torabian, V. Khatibikamal, M. Alizadehfard, Comparison of COD removal
from pharmaceutical wastewater by electrocoagulation, photoelectrocoagulation, peroxielectrocoagulation and peroxi- photoelectrocoagulation processes, Journal of Hazardous Materials
(2010), doi:10.1016/j.jhazmat.2012.03.013
Salim Zodi, Olivier Potier, François Lapicque, Jean-Pierre Leclerc. 2010. Treatment of the industrial
wastewaters by electrocoagulation: Optimization of coupled electrochemical and sedimentation
processes. Desalination, 261; 186±190.
Z. Zaroual, H. Chaair, A.H. Essadki, K. El Ass, M. Azzi. 2008. Optimizing the removal of trivalent
chromium by electrocoagulation using experimental design. Chemical Engineering Journal 148
(2009) 488±495.
Notation
I
t
TRH
COD
Current intensity
Time
Hydraulic recirculation time
Chemical oxygen demand
827
CHAPTER 7.15. CONGO RED REMOVAL USING A COMBINED SYSTEM:
ANAEROBIC TREATMENT COUPLED TO ELECTROCHEMICAL SYSTEMS
Itzel Romero-Soto (1); Celestino García-Gómez (1); Luis Álvarez-Valencia (1,2); Edna
Meza-Escalante (3); Pablo Gortáres-Moroyoqui*(1).
(1) ITSON, Departamento de Biotecnología y Ciencias Alimentarias, Ciudad Obregón, México.
(2) UANL, Facultad de Ciencias Químicas, San Nicolás de los Garza, México.
(3) ITSON, Departamento de Ciencias del Agua y Medio Ambiente, Ciudad Obregón, México.
ABSTRACT
In this work, the removal efficiency of Congo Red (CR) was improved by coupling an
UASB reactor (up-flow anaerobic reactor) to electrooxidation (EO) or electrocoagulation
(EC) cells as post-treatment. Removal efficiency of color and chemical oxygen demand
(COD) was monitored. The UASB reactor was inoculated with anaerobic granular
sludge, showing removal efficiencies of 96% and 84% for color and COD, respectively,
during 60 days of operation. The effluent from UASB reactor, with a concentration of 3.4
mg of CR/L and 168 mg of COD/L, was collected and treated independently in the cells
of EC and EO to compare these post-treatments. The system of EC was operated using
two pairs of electrodes, anodes of Fe and cathodes of stainless steel, with 110 cm 2 of
surface area. In EO were also used two pairs of electrodes, titanium (Ti) as cathodes
and Ti coated with PbO2 as anodes. In both systems, the effect of current intensity,
recycle time, and operation time were evaluated in order to compare their effect on the
removal efficiency of the color and COD remaining from UASB reactor. The highest
removal efficiencies with EC were 54.32% to 53.61% and for EC 91.25% and 99% for
color and COD, respectively. These results indicate that the EO is a suitable system as
post-treatment to ensure complete removal of color and by-products (e.g. aromatic
amines) derived from the anaerobic biotransformation of azo dyes in industrial
wastewater, as compared to EC system.
Key words: anaerobic treatment, congo red, electrocoagulation, electrooxidation.
Introduction
The textile industry is one of the major sources of pollution (Hendrickx and Boardman,
2007). Wastewater discharged form this sector contains dyes and other chemicals used
in different stages of dyeing process, producing effluents with high loads of organic
matter and chemical with structures poorly biodegradable, which alters the ecological
balance (Gutiérrez, 2004). Azo dyes are most commonly class of dyes used in the textile
industry, as its link makes better absorption in the fibers, however, this distinctive
property alters the ecological balance human health after they are introduced into the
environment. Azo dyes are considered carcinogenic and highly toxic (Hendrickx and
Boardman, 2007).
-----------*Author for all correspondence: [email protected]
828
Azo dyes are electron-accepting contaminants due to the presence of electrophilic
functional groups in their structures, making it difficult to treat them by convectional
aerobic processes. On the other hand, these pollutants can undergo anaerobic reductive
biotransformation, producing compounds susceptible to aerobic biodegradation.
However, anaerobic reduction of recalcitrant pollutants occurs slowly as a result of
toxicity effects on anaerobic consortia or due to electron transfer limitations;
consequently, anaerobic bioreactors could have deficient performance and can even
collapse. Moreover, the biotransformation products (e.g. aromatic amines) could also
adversely affect anaerobic sludge (Alvarez, 2012). To effectively remove azo dyes and
aromatic amines produced by anaerobic biotransformation process, it is necessary to
implement an additional treatment after anaerobic system. Electrochemical methods are
a new trend of advanced systems and seem to be suitable to be used as post-treatment
for the complete removal of azo dyes. Electrochemical methods consist in
electrocoagulation (EC), electrooxidation (EO), electrofloculation, and electroflotation.
EC systems apply a current to dissolve Fe (or steel) or Al anodes immersed in the
polluted water to release the corresponding metal ions yielding different Fe(II) (and/or
Fe(III)) or Al(III) hydroxide species depending on the pH of themedium. These species
act as coagulants or destabilizing agents that neutralize charges and separate colloids
and ionic products from the wastewater by sedimentation, producing some sludge. The
most popular electrochemical technique for wastewater remediation is the
electrochemical oxidation, frequently called anodic oxidation when non-chloride
solutions are treated. This procedure involves the oxidation of pollutants in an
electrolytic cell by: (i) direct electron transfer to the anode and (ii) indirect or mediated
oxidation with heterogeneous ROS formed from water discharge at the anode, such as
SK\VLVRUEHG 2+ RU FKHPLVRUEHG ³DFWLYH R[\JHQ´ 6LUpV DQG %ULOODV 7KH
objective of this study was to evaluate the removal efficiency of Congo Red (CR) using a
combined process consisted in an anaerobic treatment (UASB reactor) coupled to
electrochemical systems (EC and EO) as post-treatment.
Chemicals and Electrodes. CR was used as azo model compound (Sigma Aldrich,
purity >65%). Acetate was used as carbon and energy source (Sigma Aldrich, purity
>97). All electrodes (Fe, Stainless tell, Ti, and Ti-PbO2) were supplied by Baoji Qixin
Titanium Co., Ltd. (Hong kong, China).
Anaerobic Treatment. An UASB reactor with capacity of 900 mL was used as biological
treatment (Figure 1). The reactor was inoculated with anaerobic granular sludge (23.31
g VSS/L) collected from a full scale UASB reactor treating effluents from beverage
industries. The reactor was feed with 100 mg of CR/L and 1000 mg of acetate/L as
carbon source; and with the following mineral medium: NH4Cl (0.28 g/L), K2HPO4 (0.25
g/L), , MgSO4.7H2O (0.1 g/L), CaCl2.2H2O (0.01 g/L), and NaHCO3 (3 g/L) and was
operated at different hydraulic residence time (HRT) for 12 weeks. During the first month
the reactor was operated with an HRT of 12 h (Organic Loading Rate (OLR) of 2 g COD
/ L-d); then, at 10 h (OLR of 2.4 g COD / L-d); and finally at 8 h (OLR of 3 g COD/ L-d).
Daily samples were taken of influent and effluent for analysis of color and COD
829
concentration. The effluent was collected and stored at 4 °C during the three months of
operation. Once reached the required volume for post-treatment was thawed and
homogenized to be used in electrochemical systems.
FIGURE 1. Photography of the UASB reactor.
Electrochemical Treatments. For electrochemical systems were used acrylic reactors
with capable 0.9 L and 1.12 L for EC and EO, respectively, with dynamic flow interaction
as shown in Figure 2. Two pairs of electrodes were used for the EC, Fe as anodes and
cathodes of stainless steel, with an active surface area of 110 cm 2. In EO were used Ti
anodes coated with PbO2 and Ti cathodes. Also a source of brand power Single phase
BK Precision®. Statistical Design Expert 7.1 program was used, using the central
composite experimental design, with six focal points, evaluating a total of 20 tests, with
reference to the experimental variables shown in Table 1.
FIGURE 2. Dynamic design of electrochemical reactors.
TABLE 1. Experimental Operations variables.
830
Variable
-1.41
-1
0
+1
+1.41
2
2.3408
Electrooxidation
1
1.5
OT (Min) 32.7282 60
100
140 167.2717
20
30
CI (A)
RT(Min)
0.6591
3.1820
10
36.8179
Electrocoagulation
CI (A)
0.6591
1
1.5
2
2.3408
OT (Min)
4.8865
10 17.5
25
30.1134
RT (Min)
3.1820
10
30
36.8179
20
CI: current intensity; OT: operation time; RT: recirculation time
Analysis. Samples were collected at two points in the UASB reactor, influent and
effluent. For each sample, the dye concentration was measured at the maximum
wavelength (495) of CR using an UV spectrometer (Thermo Scientific®). The pH was
determined with a Hach potentiometer and COD concentration by the Hach method 600
nm.
The results indicate that the color and COD removal efficiencies obtained were 98.62%
and 83.94%, respectively; which are very similar to those reported by Yetilmezsoy and
Solis ((2009), where similar conditions were handled. Figure 3 shows COD and color
removal efficiencies at different HRT. Despite the variations in HRT, similar removal
efficiencies were obtained, this means that despite of the gradual increment of organic
load, the reactor remained the same removal efficiencies in terms of color and COD.
831
FIGURE 3. COD and color removal efficiencies achieved by using the UASB
reactor.
The results of EC system for COD and color removal are shown in Table 2. The
best results were found at the higher Cl tHVWHG$+RZHYHUWKHUHPRYDOGRHVQ¶W
exceed 60% and this is because during experimentation is reached the saturation point
(Lai & Lin, 2014).
TABLE 2. COD and color removal using electrocoagulation.
EXPERIMENT CI (A) OT (MIN) RT (MIN) R. COD (%) R. COLOR (%)
1
2
25
30
52.67
51.21
2
2
25
10
52.3
52.9
3
2
10
30
34.91
49.91
4
2
10
10
33.97
50.9
5
1
25
30
44.7
38.12
6
1
10
30
30.01
35.12
7
1
25
10
43.43
37.12
8
1
10
10
28.95
33.33
9
1.5
20
20
47.34
38.42
10
1.5
17.5
20
48.72
37.87
11
1.5
17.5
20
48.99
39.12
12
1.5
17.5
20
47.34
38.23
832
13
1.5
17.5
20
46.57
36.65
14
1.5
17.5
20
47.57
38.32
15
0.65
17.5
20
13.21
10.72
16
2.34
17.5
20
54.32
53.61
17
1.5
4.88
20
16.42
12.12
18
1.5
30.11
20
53.49
52.12
19
1.5
17.5
3.182
46.53
36.45
20
1.5
17.5
36.81
49.45
41.93
COLOR
53.61
Design-Expert® Software
R DQO
54.32
10.72
ctual Factor
: TRH = 20.00
54
Removal of color (%)
1 = B: TIEMPO
2 = A: INTENSIDAD
13.21
X1 = A: INTENSIDAD
X2 = B: TIEMPO
43
Actual Factor
C: TRH = 20.00
32
21
10
2
25.00
1.75
21.25
1.5
C: Current (A )
17.50
1.25
13.75
1
10.00
A : Time (min)
57
Removal of COD (%)
esign-Expert® Software
Figure 4 and 5 are the surface response plots indicating the effect of current
intensity and operation time on the removal of color and COD, respectively. The highest
removal efficiencies were achieved at 25 min of operation and 2.0 A of current intensity,
with values of ~53% for COD and color.
46
35
24
13
25.00
2.00
21.25
1.75
17.50
B: Time (min)
1.50
13.75
1.25
10.00
A : Current (A )
1.00
FIGURE 4. Effect of current intensity FIGURE 5. Effect of current intensity
and operation time on the color and operation time on the COD
removal.
removal.
The most important variable in the EC process was the current intensity, in Figure 6 the
graph with the percentage shares of each of the evaluated variables is displayed.
833
FIGURE 6. Contribution of variables on color removal.
For the system of EO, the removal efficiencies are given in Table 3. Unlike the EC
system, the most important variable in the process of EO was the operation time (Figure
9). During this process were achieved removal efficiencies above 90%. EO process had
better results than EC; nevertheless, high operating times were required. However, EO
is considered a better process because of it tends to remove contaminants completely
by oxidizing them until CO2 and H2O. Conversely, EC reaches a saturation point limiting
the removal of pollutants, although longer times of operation are applied; besides,
sludge production requires further treatment. (Rodriguez-Fernandez, 2010; Kobia et al,
2004; Suhas & Gupta, 2009).
TABLE 3. COD and color removal using electrooxidation
EXPERIMENT OT (min) RT (min) CI (A) R. COLOR (%) R. COD (%)
1
140
10
1
63.11
60.65
2
140
30
2
87.42
90.16
3
140
10
2
84
83.77
4
140
30
1
67.14
63.93
5
60
30
2
55.42
54.09
6
60
10
1
10.87
42.62
7
60
10
2
48.28
47.54
8
60
30
1
11.28
45.45
9
100
20
1.5
47.28
49.9
10
100
20
1.5
52.85
52.45
11
100
20
1.5
53.63
51.23
12
100
20
1.5
53.26
51.5
834
13
100
20
1.5
47.17
14
100
20
1.5
49.51
15
167.27
20
1.5
85.85
16
32.72
20
1.5
15.23
17
100
3.18
1.5
45.1
18
100
36.81
1.5
48.12
19
100
20
0.65
17.83
20
100
20
2.34
91.25
49.44
48.97
99
29.98
46.84
47.11
26.47
99
Figure 7 is the surface response plot of the effects of current intensity and
operation time on the color removal, which shows that the higher ranks in the
experimental variables the greater the color removal, the same was obtained in the
removal of COD and is shown in Figure 8.
Actual Factor
B: TRH = 20.00
92
Removal of color (%)
X1 = A: TIEMPO
X2 = C: INTENSIDAD
R DQO
99
10.87
26.47
X1 = A: TIEMPO
X2 = C: INTENSIDAD
70.75
Actual Factor
B: TRH = 20.00
49.5
28.25
7
2
140.00
1.75
120.00
1.5
C: Current (A)
100.00
1.25
80.00
1
60.00
A: Time (min)
99
Removal of COD (%)
R COLOR
91.25
80.75
62.5
44.25
26
2
140
1.75
120
1.5
C: Current (A )
100
1.25
80
1
A : Time (min)
60
FIGURE 7. Effect of current intensity and FIGURE 8. Effect of current intensity
operation time on the color removal.
and operation time on the COD
removal.
The most important variable in the process was the operation time. Figure 9 shows the
contribution (or the effect) as percentage of each variable studied.
835
FIGURE 9. Effect of the variables on the removal of color.
Conclusion
The most important variables for both electrochemical processes studied were operation
time and current intensity. The cumulative removal efficiencies using the UASB reactor
coupled to EO process were 98.5% and 99.8% for color and COD, respectively; while
92.6% for color and 99% removal for COD when EC was used. In both systems was
obtained water quality reuse according to the rule (which rule), but in EC system
requires a treatment of subsequent sedimentation to reduce clots present.
References
A la Torre, F., (2007). Design of reactors packed for color removal in industrial wastewater. Master Thesis
in Engineering, Facultad de Ingeniera, UNAM, Mexico. DF, Mexico, 137pp.
Alvarez, L., (2010). Immobilization of humic substances on metal-oxides (nano) particles and their impact
in redox processes. Doctoral Thesis. Instituto Potosino de Investigación, Científica y Tecnológica,
A.C. San Luis Potosí, S.L.P. México.
Beltran H.; Gonzales, M., (1991). Treatment of wastewater in the textile industry. In: Chemical
Engineering. Vol. 23, No.273 p.131-134.
Chen, G., (2004). Electrochemical technologies in wastewater treatment. Sep. Purif. Tech .: 38, 11-41.
Domenech, X.; Litter M. I.; Jardim, W., (2001). Advanced oxidation processes for the removal of
pollutants. Red Cyted VIII-G, Comisión Nacional de Energía Atómica, Buenos Aires, 1, 3-26.
Fai, G. and Geyes, J., (1993). Water purification and treatment and removal wastewater, Editorial Limusa,
first edition.
Forgacs, E., Cserhati, T., Oros, G., (2004). Removal of synthetic dyes from wastewaters: a review.
Environmental. Pp. 953-971.
García, M., (2009). Treatment of industrial wastewater. Block III. Study of real water in the textile industry.
Gogate P.; Pandit A., (2004a). A review of imperative technologies for wastewater treatment I: oxidations
technologies at ambient conditions, Advances Environmental Research, 8, pp. 501-51.
Gogate P.; Pandit A., (2004b). A review of imperative technologies for wastewater treatment II: hybrid
methods, Advances Environmental Reserch, 8, pp. 553-97.
836
Gupta, V. and Suhas, (2009). Application of low-cost adsorbents for dye removal. A review. Journal of
environmental Management. 90: 2313-2342.
Gutierrez, A., (2004). The clothing industry in Mexico Economic Analysis. Universidad Metropolitana
(UAM). Mexico.
Hendrickx and Boardman, (2007). Pollution Prevention Studies in the Textile Wet Processing Industry.
Kobya, M.; Orhan T.; Bayramoglu, M., (2004). Treatment of textile wastewaters by electrocoagulation
using iron and aluminum electrodes. In: Journal of Hazardous Materials. Vol. 100, No. 1-3, p.163178.
Lai, L. and Lin, S., (2014) Treatment of chemical mechanical polishing wastewater by electrocoagulation:
system performances and sludge settling characteristics. In: Chemosphere Vol 54, No. 3, pp. 235242.
Linares-Hernández; Martínez-Miranda; Barrera-Díaz; Pavón-Romero; Bernal-Martínez; Lugo-Lugo,
(2011). Oxidation of persistent organic matter in industrial wastewater by electrochemical treatment.
ACI: VOL. 2 (1), pp. 21-36.
Mollah, M .; Morkovsky, P .; Gomes, J .; Kesmez, M .; Parga, J .; Cocke, D., (2004); Fundamentals,
present and future perspectives of electrocoagulation, Journal of Hazardous Materials 114, 199210.
Pound, J.; Borchert, M.; Vigelanhn, L.; Storm, T., (2004). Two stage biological treatment of a diazo
reactive textile dye and the fate of the dye metabolites. Chemosphere, 56; 167-180.
Rios, J., (2007). Study the feasibility of a system of electrocoagulation water treatment from industrial
laundries reuse purposes. Bolivia. Chapter I, the water. Page. 22-24.
Rodie B. and Hardenberg., (1987). Wastewater treatment. Chapter IV. Sanitary engineering. Mexico DF.
Rodriguez-Fernandez , (2010). Advanced treatment of industrial wastewater. Report Tegnological
Surveillance. Madrid.
Sires, I.; Brillas E., (2012).Remediation of water pollution caused by pharmaceutical residues based on
electrochemical separation and degradation technologies: A review. Journal Environment
International 40. Pp. 212±229.
Solis R., (2009). Coagulation surface water using cassava starch (Manihot esculenta Crantz), mixed with
conventional coagulants. Thesis. Universidad Autonoma de Juárez, Tabasco, Villahermosa,
México.
Yetilmezsoy, K.; Ilhan, F.; Sapci-Zengin, Z.; Gonullu, M., (2009). Decolorization and COD reduction of
UASB pretreated poultry manure wastewater by electrocoagulation process.: A post-treatment
study. Journal of Hazardous Materials pp.120-132.
Zaroual, Z.; Azzi, M.; Saib, N.; Chainet, E., (2006). Contribution to the study of electrocoagulation
mechanism in basic textile effluent. In: Journal of Hazardous Materials. Vol 131, No.1-3.; p.73-78.
837
CHAPTER 7.16. OPTIMIZATION OF PHENOL REMOVAL USING TI/PBO2 ANODE
WITH RESPONSE SURFACE METHODOLOGY
García-Gómez, Celestino (1); Nápoles-Armenta, Juan (1) y Gortáres-Moroyoqui, Pablo
(1)*
(1) Departamento de Biotecnología y Ciencias Alimentarias. Instituto Tecnológico de Sonora. 5 de
Febrero 818 sur. C. P. 85000 Obregón, Sonora, México.
ABSTRACT
In this study Response Surface Methodology (RSM) was employed to investigate the
effects of different operating conditions on the removal of phenol by the electroxidation
with Ti/PbO2 anode. Central Composite Design (CCD) was used for the optimization of
the electroxidation process and to evaluate the effects and interactions of process
variables: current intensity, electrolysis time and recirculation flow rate on the removal of
phenol. A sample of phenol concentration of 10 mg L-1was used in the experimental
study. The optimum conditions for phenol removal were established as 1.12 A current
intensity, 40 min electrolysis time and 188 ml min-1 recirculation flow rate and was able
to remove 78.97±1.72%. The decay kinetics was fitted a first-order reaction.
Key words: electrooxidation, removal, phenol, PbO2, Response Surface Methodology.
Introduction
Phenolic compounds are commonly present in wastewater produced by several
industries, such as petroleum refineries, formaldehyde resins, pharmaceuticals, textile,
dyes, coke, plastics, coal conversion and chemical plants (Buscaa et al., 2008). They
are among 129 pollutants listed by the Environmental Protection Agency (EPA) due to
their toxicity and difficulty to be degraded (Keith et al., 1979). Thus, various methods for
the treatment of phenolic wastewater have been used, such as wet air oxidation (Lefevre
et al., 2011), adsorption (Su et al., 2011), chemical oxidation (Manojlovic et al., 2007),
photocatalysis (Xu et al., 2011), biological treatment (Ahmad et al., 2012), and ozone
oxidation (Joshi et al., 1982). But there are few sufficiently efficient processes for the
removal of these toxic organic pollutants. The electrochemical oxidation technology has
become a promising method for the toxic, bio-refractory, and highly concentrated
organic wastewater treatment because of its simplicity, easy control, strong oxidation
performance, and environmental compatibility (Zhu et al., 2008; Zhuo et al., 2011; Wei et
al., 2011; Zhu et al., 2013). This technique is based on hydroxyl radical (OH) production,
which is a powerful oxidant able to react efficiently and non-selectively with several
types of organic pollutants such as phenol (Alves et al., 2013; Fenga et al., 2013; Luoa
et al., 2014; Urtiaga et al., 2014). The electrochemical oxidation of aqueous solutions,
which contain organic pollutants, by the use of Ti/catalyzer anode, proceeds in two steps
(Zhu et al., 2013). The first step is the anodic discharge of the water, forming hydroxyl
radicals which are absorbed on the active sites of the electrode (M)
838
ଶ ൅ ՜ ሺιሻ ൅ ା ൅ Ǧ (1)
After this the absorbed hydroxyl radical oxidizes the pollutant (R):
ሺιሻ ൅ ՜ ൅ ൅ ା ൅ Ǧ
(2)
where RO represents the oxidized pollutant which can be produced continuously
by the hydroxyl radicals which are also constantly formed, since the anodic discharge of
the water goes on.
In the electrochemical oxidation (EO) process, the removal of organic pollutant
mainly depends on the properties of the anodes (Feng et al., 2003; Li et al., 2005;
Szpyrkowicz et al., 0DUWÕQH]et al., 2008). Consequently, a wide variety of anodes
with a thin layer of metal oxides such IrO2 (Chatzisymeon et al., 2010), RuO2 (Tran et
al., 2009), SnO2 (Hu et al., 2010), and PbO2 (Samet et al., 2010) on a base metal,
usually titanium, have effectively been used on organic pollutant wastewater treatment.
PbO2 electrode has become popular due to its low electrical resistivity, favorable
overpotential for oxygen evolution, good chemical stability, a strong ability to produce
hydroxyl radicals and low cost compared to noble metals (Suryanarayanan et al., 2006;
Panizza et al., 2003). Hence, there is a great interest in the application of PbO2 anodes
for the oxidation of organics. The main goal of this study was to assess the removal of
phenol in aqueous medium (initial concentration of 10 mg L í1) by electrooxidation
process using Ti/PbO2 concentric anode electrode using a response surface
methodology (RSM).
Preparation of the synthetic solution. Phenol analytical grade reagent (99%) was
obtained from Fermont (México). Synthetic solution of phenol was prepared in glass
beakers containing 10 mg of phenol in 1000 ml of distilled water, which was solubilized
using magnetic stirrer (500 rpm) at environmental temperature (25 oC) during 24 h. The
resulting mixture constituted the synthetic phenol with final concentration of 10 mg L -1.
Na2SO4 was used as supporting electrolyte, which was added at a concentration of 400
mg L-1 to increase the conductivity. Na2SO4 was an analytical grade reagent supplied by
Fermont (México). All phenol removal experiments were performed with a volume of 650
ml.
Electrochemical cell. The reactor unit used had 650 ml of capacity and was made
of acrylic material with a work dimension of 15 cm (height) × 12 cm (diameter). The
electrolytic cell was comprised of one anode and two cathode electrodes in the form of
concentric tubes. The anode electrode (10 cm diameter × 0.2 cm thick) were titanium
coated with lead oxide (Ti/PbO2) and titanium (Ti) was used as cathodes (11 cm
diameter and another one 9 cm diameter, both with 0.2 cm thick). The inter-electrode
gap was 10 mm in the electrolytic cell. The electrodes were vertically installed inside the
electrolytic cell, the anode was located between the two cathodes and were supplied
by Baoji Qixin Titanium Co., Ltd. (Hong kong, China). The system worked with a
recirculation upflow. 839
Experimental design for phenol EO. Experimental design of the EO process for
phenol removal was carried out by using the RSM. RSM is a collection of mathematical
and statistical techniques that are beneficial for the modeling and analysis of problems in
which a response of interest is influenced by several variables and the objective is to
search the optimum conditions of variables to predict targeted responses (Myers, et al.,
2002). RSM consists of an empirical modeling technique devoted to the evaluation of
relations existing between a group of controlled experimental factors and the observed
results. RSM is an important branch of experimental design and a critical methodology in
developing new processes, optimizing their performance and improving design and
formulation of new products. The most extensive applications of RSM are in industrial
research, particularly in situations where several input variables influence the process
performance measure. This process performance measure is called the response and
the input variables are called independent variables (Myers et al., 2002).
Response surface procedures are not primarily used for the purpose of
understanding the mechanism of the system or process; rather their purpose is to
determine the optimum operating conditions or to determine a region for the factors in
which certain operating specifications are met (Myers et al., 2002). In this study, the
Central Composite Design (CCD), which is a widely used form of RSM, was selected for
the optimization of EO process used for the removal of phenol. The CCD is an ideal
design tool for sequential experimentation and allows testing the lack of fit when an
adequate number of experimental values are available. In the present study a threefactorial and a five-level central composite experimental design, with six replicas at the
center point leading to a total number of twenty experiments was employed for response
surface modeling. The variables (independent factors) used in this study were: current
intensity (X1), electrolysis time (X2) and recirculation flow rate (X3). Phenol removal
efficiency (Y1) and energy consumption (Y2) were considered as the dependent factor
(response). The actual values of process variables and their variation limits were
selected based on the values obtained in preliminary experiments and coded as shown
in Table 1.
TABLE 1. Experimental range and levels of independent process variables.
Independent
variables
Current
intensity
(A)
Electrolysis
time
(min)
Recirculation flow
rate (ml min-1)
Factor
X1
Experimental field
-Į-1.68)
Low (-1)
0.66
1
Middle (0)
1.5
High (1)
2
Į
2.34
X2
13.18
20
30
40
46.82
X3
155.27
188
236
284
316.73
For the evaluation of experimental data, the response variable was fitted by a secondorder model in the form of quadratic polynomial equation given below;
௞
௞
௞
(3)
ଶ
ൌ ܾ଴ ൅ ෍ ܾ௜ ܺ௝ ൅ ෍ ܾ௜௜ ܺ ൅ ෍ ෍ ܾ௜௝ ܺ௜ ܺ௝ ൅ ݁௜ ௜ୀଵ
௜ୀଵ
௝
௜ୀଵ
840
where Y is the predicted response (phenol removal efficiency, %) used as the
dependent variable; Xi (i=1, 2 and 3) are the variables or independent factors and b 0, bi
(i=1, 2 and 3), bii and bij (i=1, 2 and 3; j=1, 2 and 3) are the model coefficients,
respectively. Experimental data were analyzed using Design-Expert 7 program including
ANOVA to obtain the interaction between the process variables and the responses.
Two-dimensional contour plots and three-dimensional curves of the response surfaces
were developed using the same program.
Optimization procedure. Optimization of the EO for phenol removal was carried out by
using the Design-Expert program 7 version. The optimization module in Design-Expert
searches for a combination of factor levels that simultaneously satisfy the requirements
placed on each of the responses and factors. According to the optimization step of the
program, the desired goal for each variable (current intensity, electrolysis time and
recirculation flow rate) and responses (phenol removal efficiency and energy
consumption) should be chosen. The possible goals are: maximize, minimize, target,
within range, none (for responses only) and set to an exact value (factors only). The
goals are then combined into an overall desirability function. The program seeks to
maximize this function. The goal seeking begins at a random starting point and proceeds
up the steepest slope to a maximum. There may be two or more maximums because of
curvature in the response surfaces and their combination into the desirability function.
By starting from several points in the design space chances improve for finding the best
local maximum.
Experimental results. In optimizing a response surface, an adequate fit of the model
should be obtained to avoid poor or ambiguous results (Myers, et al., 2002). This is
important to ensure the adequacy of the employed model. Table 2 shows the analysis of
variance (ANOVA) of regression parameters of the predicted response surface quadratic
model for phenol removal by EO process using the results of all experiments performed.
As it can be seen from the table, the model F-value of 13.92 and 52.23 with a low
probability value of Pr>F=0.0002 and <0.0001 indicate that the model is significant for
SKHQROUHPRYDODQGHQHUJ\FRQVXPSWLRQUHVSHFWLYHO\7KH³$GHTXDWH3UHFLVLRQÚDWLRRI
the model is 12.095 and 22.792 for phenol removal and energy consumption,
respectively. The value of correlation coefficient (R2 of 0.9261 and 0.9075 for phenol
removal and energy consumption, respectively) indicates that only 7.39% of the total
variation for phenol removal could not be explained by the empirical model and
expresses good enough quadratic fits to navigate the design space and 9.25% for the
case of energy consumption. Joglekar and May suggested that R2 should be at least
0.80 for a good fit of a model. The R2 value (0.875) obtained in the present study for
these response variables was higher than 0.80, indicating that the regression models
explained the reaction well. Hence, the response surface model developed in this study
for predicting phenol removal efficiency and energy consumption was considered to be
satisfactory.
The final regression model in terms of coded factors for phenol removal has been
expressed by the following second-order polynomial equation:
841
(4)
ଵ ൌ ͸ͻǤʹͶ ൅ ͷǤ͹Ͷܺଵ ൅ ͳͶǤͺ͹ܺଶ െ ͶǤ͸͹ܺଷ
െ ͲǤʹ͸ܺଵ ܺଶ െ ͲǤ͵ͳܺଵ ܺଷ
െ ͲǤ͹ʹܺଶ ܺଷ
െ͸Ǥͳͳܺଵଶ െ ʹǤ͵͵ͳܺଶଶ ൅ ͳǤ͹͸ܺଷଶ
VXEMHFWHGWRíXi
In terms of actual factors, an empirical relationship between phenol removal efficiency
and the variables has been expressed by the following second-order polynomial
equation:
(5)
ଵ ൌ ǦʹͲǤ͵ͺ ൅ ͺͻǤͶͶଵ ൅ ͵Ǥ͵Ͳଶ ǦͲǤ͵ͻ ଷ ǦͲǤͲͷଵ ଶ ǦͲǤͲͳ͵ଵ ଷ ǦͳǤͶͻଶ ଷ ǦʹͶǤͶͶଵଶ ǦͲǤͲʹଶଶ ൅ ͹Ǥ͸ͷଷଶ
VXEMHFWHG WR $ FXUUHQW LQWHQVLW\ HOHFWURO\VLV WLPH DQG
155.27 ml min-1 Recirculation flow rate ml min-1.
TABLE 2. ANOVA results for the response surface quadratic model for CBZ
removal and energy consumption.
Source
CBZ removal
Model
Residual
Lack of fit
Pure error
Energy
consumption
Model
Residual
Lack of fit
Pure error
Analysis of variance
d.f. a
Sum of square
Mean of square
F-Value
Pr>F
9
10
5
5
4448.64
355.10
345.83
9.27
494.29
35.51
69.17
1.85
13.92
0.0002
37.31
0.0006
3
16
11
5
484.20
49.35
48.20
1.14
161.40
3.08
4.38
0.23
52.33
<0.0001
19.16
0.0022
a Degree
of freedom
= 0.9261 for CBZ removal.
R2 = 0.9075 for Energy consumption.
R2
The surface response and contour plots of the quadratic model with one variable
kept at central level and the other two varying within the experimental ranges are shown
in Figs. 1±3. In Figure. 1, the response surface and contour plot were developed as a
function of electrolysis time and current intensity while the recirculation flow rate was
kept constant at 236 ml min-1 being the central level. 80% phenol removal efficiency was
realized at 236 ml min-1 recirculation flow rate while current intensity was above 1.4 A
and electrolysis time were higher than 37 min.
842
FIGURE 1. The effect of the electrolysis time and current intensity on phenol
removal (recirculation flow rate: 236 ml min-1).
The effect of the variables, i.e., electrolysis time and recirculation flow rate on
phenol removal efficiency is illustrated in Figure 2. When current intensity was kept
constant at 1.5 A, phenol removal efficiency increased with increasing electrolysis time
and the recirculation flow rate. As seen from the contour plot, important phenol removal
could be achieved at 1.5 A for electrolysis time higher than 35 min and the lowest
recirculation flow rate. In order to achieve high level removal efficiencies (>80%),
electrolysis time values become relatively important for recirculation flow rate lower than
200 ml min-1. The interrelation between the variables, therefore, is important in terms of
optimization, however, the electrolysis time has a more significant impact on phenol
removal efficiency.
FIGURE 2. The effect of the electrolysis time and recirculation flow rate on phenol
removal (current intensity: 1.5 A).
To study the effect of current intensity and recirculation flow rate on phenol
removal the experiments were carried out with current intensity varying from to 0.66 to
2.34 A and under different recirculation flow rates at electrolysis time of 30 min. The
results were displayed in Fig. 3. This figure clearly shows that phenol removal efficiency
843
increased with increasing current intensity at lower recirculation flow rate. 80% phenol
removal efficiency was achieved in less than 200 ml min-1with the current intensity
higher than 1.5.
FIGURE 3. The effect of the current intensity and recirculation flow rate on phenol
removal (electrolysis time: 30 min).
Similar to interrelation between the current intensity and electrolysis time, it can be
concluded from the correlation between the electrolysis time and the recirculation flow
rate that the electrolysis time has a more significant impact on phenol removal efficiency.
The actual (measured) values selected as combinations to reflect a range of
current intensity, electrolysis time and recirculation flow rate were used to assess the
model responses in terms of specific numerical values in addition to the statistical
evaluation given above. The comparison of actual and predicted values of phenol
removal are listed in Table 3. The agreement between the actual and predicted values of
phenol removal is satisfactory and in accordance with the statistical significance of the
quadratic model presented in Table 2.
TABLE 3. Comparison of actual and predicted values of phenol removal.
Experiment plan X1 (A) X2 (min) 1 20 1 40 1 20 1 40 2 20 2 40 2 20 2 40 0.66 30 2.34 30 1.5 13.18 X3 (mL min-‐1) 188 188 284 284 188 188 284 284 236 236 236 ܻଵ௔ ǣ
Actual Removal efficiency (%) 52.28 80.37 42.27 64.34 61.58 85.49 47.17 71.35 35.73 66.71 30.79 ܻଵ௕ ǣ Predicted Removal efficiency (%) 54.14 79.48 41.55 66.88 64.22 83.99 47.91 70.56 37.54 68.88 32.45 Relative deviation ൫ܻଵ௔ െ ܻଵ௕ ൯ -‐1.86 0.89 -‐2.28 -‐2.54 -‐2.64 1.50 -‐0.74 0.79 -‐1.81 -‐2.17 -‐1.66 ܻଶ ǣEnergy Consumption (kWh m-‐3) 5.63 11.27 3.33 6.66 15.87 21.09 10.90 21.09 3.11 18.30 4.55 844
1.5 1.5 1.5 1.5 1.5 1.5 1.5 1.5 1.5 46.82 30 30 30 30 30 30 30 30 236 155.27 316.73 236 236 236 236 236 236 93.15 76.22 70.75 69.62 68.99 71.71 69.09 68.67 67.62 92.85 77.67 71.58 71.24 70.42 70.19 70.42 66.83 65.81 0.3 -‐1.45 -‐0.83 3.07 -‐1.43 3.52 0.44 -‐1.84 2.81 16.17 8.31 9.00 9.40 10.22 9.81 10.22 10.63 10.63 The main objective of the optimization is to determine the optimum values of
variables for phenol removal with EO from the model obtained using experimental data.
In optimization, the desired goal for the response phenol removal efficiency was chosen
to a maximum, the energy consumption like minimum and the variables of current
intensity, electrolysis time and recirculation flow rate were selected to be within range.
The optimization results of the process variables for phenol removal are shown in Table
4. The desirability function value was found as 0.702 for these optimum conditions. As
seen from the table, EO is applicable for complete phenol removal under reasonable
operating conditions. As seen from the table, EO is applicable for phenol removal under
reasonable operating conditions.
TABLE 4. Optimum operating conditions of the process variables for phenol
removal efficiency.
Current
intensity (A)
1.12
Electrolysis
time (min)
40
Recirculation flow Removal
rate (ml min-1)
efficiency (%)
188
81.17
Energy consumption Desirability
(kWh/m3)
11.10
0.702
To confirm the model adequacy and the validity of the optimization procedure,
additional experiments were performed under the projected optimal conditions based on
the results from the model that are given in Table 4. During the series of experiments,
the maximal value of the CBZ removal obtained under optimal process conditions has
been confirmed experimentally, the residual concentration of CBZ was analyzed, the
experimental response was 78.97±1.72% and recorded at the end of these experiments,
a value that was very near to that proposed by the model (81.17% of CBZ degradation).
These results testify that the RSM is a powerful tool for optimizing the operational
conditions of EO for phenol removal.
The EO using optimal conditions was applied to the phenol removal, the results of
RNO degradation are shown in Fig. 4. The experiments were carry out by triplicate to
verify the effectiveness and reproducibility of the EO performance of phenol removal. As
it is observed from this results, the removal rate of phenol increased with treatment time.
Apparent rate constant for oxidative removal of phenol, obtained from the slop of straight
lines of the inset in Fig. 4 was 0.0211 min-1, the exponential decrease of phenol
concentration indicates a first order reaction NLQHWLFZLWKDFRUUHODWLRQFRHI¿FLHQW52) of
0.965.
845
FIGURE 4. Removal of CBZ using optimal conditions process. The insert represent
the kinetic analysis following first order reaction.
Conclusion
In the present study, the performance of electrochemical treatment of phenol was
studied focusing on the influence of operating parameters such as current intensity,
electrolysis time and recirculation flow rate by using RSM with CCD. The results
obtained from the present study revealed that RSM was a suitable method to optimize
the operating conditions of EO for phenol removal. The response surface models
developed in this study for predicting phenol removal efficiency were considered to be
adequately applicable. Analysis of variance showed a high coefficient of determination
value (R2 =0.9261), ensuring a satisfactory adjustment of the second-order regression
model with the experimental data. The optimum conditions of phenol removal were 1.12
A current intensity, 40 min electrolysis time and 188 recirculation flow and whit this
conditions the system is able to remove more of 80%. This process appears to be a
promising technology that can be used as tertiary treatment to remove emerging
pollutants from wastewaters.
References
846
Ahmad, S.A.; Shamaan, N.A.; Arif, N.M.; Koon, G.B.; Shukor, M.Y.A.; Syed, M.A. (2012). Enhanced
phenol degradation by immobilized Acinetobacter sp. strain AQ5NOL 1. World J Microbiol
Biotechnol. 28: 347±52.
Alves, S. A.; Ferreira, T. C.R.; Migliorini, F. L.; Baldan, M. R.; Ferreira, N. G.; Lanzaa, M. R.V. (2013).
Electrochemical degradation of the insecticide methyl parathion using a boron-GRSHGGLDPRQG¿OP
anode. J. Electroanal. Chem. 702; 1±7.
Buscaa, G.; Berardinelli, S.; Resini, C.; Arrighi, L. (2008). Technologies for the removal of phenol from
fluid streams: A short review of recent developments. Journal of Hazardous Materials. 160: 265±
288.
Chatzisymeon, E.; Fierro, S.; Karafyllis, I.; Mantzavinos, D.; Kalogerakis, N.; Katsaounis, A. (2010).
Anodic oxidation of phenol on Ti/IrO2 electrode: experimental studies. Catal Today. 151; 185±199.
Eckenfelder, W.W. (1989). Industrial Water Pollution Control, McGraw Hill Book Company, NewYork.
Fenga, L.; van Hullebuscha, E. D.; Rodrigo, M. A.; Esposito, G.; Oturan, M. A. (2013). Removal of residual
anti-LQÀDPPDWRU\ DQG DQDOJHVLF SKDUPDFHXWLFDOV IURP DTXHRXV V\VWHPV E\ HOHFWURFKHPLFDO
advanced oxidation processes. A review, Chem. Eng. J. 228; 944±964Joshi, M.J.; Shambaugh,
R.L. (1982). The kinetics of ozone±phenol reaction in aqueous solution. Water Res.16: 933±8.
Feng, Y.J.; Li, X.Y. (2003). Electro-catalytic oxidation of phenol on several metal-oxide electrodes in
aqueous solution. Water Research. 37; 2399±407.
Hu, F.; Cui, X.; Chen, W. (2010). Pulse electro-codeposition of Ti/SnO2±Sb2O4±CNT electrode for phenol
oxidation. Electrochem Solid State Lett. 13; F20±3.
Keith, L.H.; Telliand, W.A. (1979). ES&T special report: priority pollutants. I. A perspective view.
Environmental Science & Technology. 13: 416±23.
Manojlovic, D.; Ostojic, D.; Obradovic, B.; Kuraica, M., Krsmanovic, V.D.; Puric, J. (2007). Removal of
phenol and chlorophenols from water by new ozone generator. Desalination. 213: 116±22.
0DUWÕQH]-Huitle, C.A.; Battisti, A.D.; Ferro, S.; Reyna, S.; Cerro-López, M.; Quiro, M.A. (2008). Removal of
the pesticide methamidophos from aqueous solutions by electrooxidation using Pb / PbO2 , Ti /
SnO2 , and Si / BDD electrodes. Environmental Science & Technology. 42; 6929±35
Myers, R.H.; Montgomery, D.C. (2002). Response Surface Methodology: Process and Product
Optimization using Designed Experiments, second ed., John Wiley & Sons, USA.
Lefevre, S.; Boutin, O.; Ferrasse, J. H.; Malleret, L.; Faucherand, R.; Viand, A. (2011). Thermo dynamic
and kinetic study of phenol degradation by a non-catalytic wet air oxidation process. Chemosphere.
84: 1208±15.
Li, X.Y.; Cui, Y.H.; Feng, Y.J.; Xie, Z.M.; Gu, J.D. (2005). Reaction pathways and mechanisms of the
electrochemical degradation of phenol on different electrodes, Water Research. 39; 1972±81.
Luoa, Y., Guoa, W.; Ngoa, H. H.; Nghiemb, L. D.; Hai, F. I.; Zhang, J. S.; Liang, X. C. (2014). A review on
the occurrence of micropollutants in the aquatic environment and their fate and removal during
wastewater treatment. Sci. Total Environ. 619±641.
Panizza, M.; Cerisola, G. (2003). Influence of anode material on the electrochemical oxidation of 2naphthol: Part 1. Cyclic voltammetry and potential step wxperiments. Electrochim Acta. 48; 3491±
3497.
Samet, Y.; Agengui, L.; Abdelhe´di. R. (2010). Anodic oxidation of chlorpyrifos in aqueous solution at lead
dioxide electrodes. J Electroanal Chem. 650; 152±8.
Suryanarayanan, V.; Nakazawa, I.; Yoshihara, S.; Shirakashi, T. (2006). The influence of electrolyte
media on the deposition/dissolution of lead dioxide on boron- doped diamond electrode²a surface
morphologic study. J Electroanal Chem. 592; 175±82.
Su, J.; Lin, H.; Wang, Q.; Xie, Z.; Chen, Z. (2011). Adsorption of phenol from aqueous solutions by
organomontmorillonite. Desalination. 269: 163±9.
Szpyrkowicz, L.; Kaul, S.N.; Neti, R.N.; Satyanaray, S. (2005). Influence of anode material on
electrochemical oxidation for the treatment of tannery wastewater. Water Research. 39; 1601±13.
Xu, X.; Yi, Z.; Chen, D.; Duan, X.; Zhou, Z.; Fan, X.; Jiang, M. (2011). Evaluation of photocatalytic
production of active oxygen and decomposition of phenol in ZnO suspensions. Rare metals. 30:
188±91.
Tran, L. H.; Drogui, P.; Mercier, G., Blais, J. F. (2009). Electrochemical degradation of polycyclic aromatic
hydrocarbons in creosote solution using ruthenium oxide on titanium expanded mesh anode. J
Hazard Mater. 164; 1118±29.
847
Urtiaga A., Fernandez-Castro P., Gómez P., Ortiz I. (2014). Remediation of wastewaters containing
tetrahydrofuran. Study of the electrochemical mineralization on BDD electrodes. Chem. Eng. J. 239;
341±350.
Zhu, X.; Tong, M.; Shi, S.; Zhao, H.; Ni, J. (2008). Essential explanation of the strong mineralization
performance of boron-doped diamond electrodes. Environ Sci Technol. 42: 4914±20.
Zhuo, Q.; Deng, S.; Yang, B.; Huang, J.; Yu, G. (2011). Efficient electrochemical oxidation of
perfluorooctanoate using a Ti/SnO2±Sb±Bi anode. Environ Sci Technol. 45: 973±9.
Zhu, X.; Ni, J.; Lai, P. Advanced treatment of biologically pretreated coking wastewater by electrochemical
oxidation using boron-doped diamond electrodes.
Wei, J.; Feng. Y.; Sun, X.; Liu, J.; Zhu, L. (2011). Effectiveness and pathways of electrochemical
degradation of pretilachlor herbicides. J Hazard Mater. 189: 84±91.
848
CHAPTER 7.17. REMOTION OF COLOR AND COD BY ELECTRO-OXIDATION FOR
TEXTILE EFFLUENT
Vianey Ariadna Burboa-Charis (1); Petia Mijaylova-Nacheva (2);
José de Jesús Treviño-Resendez (2); Celestino García-Gómez (1);
Itzel Celeste Romero-Soto (1); Pablo Gortáres-Moroyoqui*(1)
(1) ITSON, Cd. Obregón, Sonora, México;
(2) IMTA, Juitepec, Morelos, México;
ABSTRACT
Wastewater from the textile industry is characterized by a high concentration of chemical
oxygen demand (COD) and color concentrations. The presence of dyes in water can
inhibit photosynthesis, when they are accumulated at the environment. In addition, they
represent a serious public health problem because of many of the compounds used in
the textile industry, or their degradation products may be carcinogenic, mutagenic or
may cause allergic reactions. The purpose of this study was to evaluate the effect of
current intensity and operation time on color and COD removal by electro-oxidation. The
effluent comes from a textile industry in Jiutepec, Morelos, which was previously passed
on a membrane biological reactor. Color and COD concentration before to be inside to
the EO system were 2493.33 Co-Pt and 2605.45 mg L-1, respectively. A cylindrical
acrylic reactor height of 19.5 cm was used, with an operating capacity of 1.25 L, working
with three electrodes; two cathodes (Ti) and one anode (TI/PbO2) with cathode-anodecathode arrangement. An experimental matrix was performed in the statistical package
Statgraphics Centurion with a 2k (22) factorial design and central composite design
method (three central points) and four extreme levels (low and high) Variables studied
were current intensity and operation time. The removal efficiency for the electrochemical
system was 81.36 for COD and 93.12% for color.
Keywords: color, COD, cost, electro-oxidation, response surface method, textiles
wastewaters.
Introduction
Synthetic dyes are used extensively by several industries including textile dyeing (60%),
paper (10%) and plastic matter (10%). The textile industry uses about 10,000 different
dyes. The worldwide annual production of dyestuffs is over 7105 tons (Poon et al. 1999).
Textile manufacturing is one of the largest industrial producers of wastewater;
approximately 125±150 L of water are used for 1 kg of textile product (U.S. EPA,
1995).The textile industry wastewater represent one of the most difficult to treat
effluents. Because they contain high concentrations of dyes, refractory organic
contaminants, toxic compounds, inhibitor components, tensoactive subtances and
chlorinated components, among others. Regulatory requirements and the need to save
and / or reuse water in industry, makes it necessary to investigate on processes to
improve the removal of poorly biodegradable components, colloidal particles, viruses,
849
bacteria and the possibility to eliminate color from effluent in the production process
(Kural et al, 2001).
The application of electrochemical treatment in controlling pollution from industrial
effluents has advantages over conventional treatments (chemical and biological), such
as lower installation area, less sludge generated, versatility to handle a wide range of
concentrations, contaminant selectivity, low costs for chemical reagents, lower
equipment requirements, ease of operation and low maintenance (Mendoza et al, 2000).
Electrochemical oxidation or electro-oxidation (EO) is the most popular electrochemical
procedure for removing organic pollutants from wastewaters. This technique has been
recently used for decolorizing and degrading dyes from aqueous solutions. It consists in
the oxidation of pollutants in an electrolytic cell by direct anodic oxidation (or direct
electron transfer to the anode), which yields very poor decontamination, or by chemical
reaction with electrogenerated species from water discharge at the anode such as
physically adsorbed (Martínez-Hutle and Brillas, 2009). The main objective of this study
was to evaluate color and COD removal by electro-oxidation system of an effluent from
the textile industry. As specific objectives were to determine the effects of time and
intensity of color and COD removal by electro-oxidation, treatment costs, and ultimately
get an optimal experiment.
Materials y Methods
Sampling and characterization of wastewater. The effluent comes from a textile
industry in Jiutepec, Morelos, which was previously passed on a membrane biological
reactor (MBR). The MBR effluent was accumulated to be treated in the electro-oxidation
system. Sampling was according to the procedure established in regulations
(SEMARNAT, 2003a). The physicochemical characterization of wastewater was for
COD (Standard Methods: 5220D), color (Manual HACH, 2010) and field parameters (pH
and Temperture).
Reactor Design. The electrochemical cylindrical reactor was constructed, using acrylic
material with an operating capacity of 1.25 L, a height of 19.5 cm and 6 cm tapered
portion; it worked with three electrodes (two cathodes of Ti and one anode of Ti / PbO 2)
with an arrangement of cathode, anode and cathode measures 3.3, 4.7 and 6.4 cm in
diameter respectively and 15 cm long for all.
Preliminary tests and variables studied. Preliminary tests were conducted using
different current intensities (3, 6 and 9 A) and two recirculation flows (120 and 480
ml/min), with a fixed operation time of 120 minutes. Samples of 10 mL were taken, every
30 minutes from the reactor and analyzed to determine COD, color, temperature and pH
during electrochemical treatment. Anode and cathodes were washed and dried, after
each experiment to avoid passivation problems. The response variables were the
percentage of color removal and COD removal.
Color and COD removal by: experimental design and data analysis. According to
the results obtained in preliminary testing an experimental matrix was performed in the
850
statistical package Statgraphics Centurion with a 2k (22) factorial (where k is the number
of factors) and central composite design method (three central points and four extreme
levels) to evaluate the effect of current intensity and operation time on the COD and
color removal. A total of five levels were tested for each variable (low extreme, low,
central, high and high extreme) with eleven experimental conditions or different
treatments were carried out. Recirculation flow rate was constant (480 ml /min). The
data analysis was done by the response surface method, Pareto chart and main effect
graphics. The optimal experiment was generated in the Design Expert program 9,
maximizing color removal percent and minimizing operation time.
Economic aspect. The economic study only considered energy consumption for
electrochemical treatment in dollars per kWh. Energy consumption per volume of treated
effluent in Kwhm-3, is calculate by equation 1, where I is current intensity (A), V is cell
voltage (V), t is operation time (h) and Vs is operation volume (m3) (Martínez-Huitle and
Brillas, 2009).
Energy consumption= IVt/Vs
(1)
The cost of energy was calculated by equation 2, where EC is energy consumption
(Kwhm-3), P is the energy cost (dollar per Kw-h), and R is the rate Mexican pesos (MP)
to US dollars. The energy price for agricultural irrigation (9-CU) with a value of 0.52
Mexican pesos per kilowatt-hour (Commisión Federal de Electrician, 2013-2014), and a
R values of 13.1055 MP per dollar (Secretaría de Economía, 2014) were considered at
the present study.
Cost of treatment= ECRP
(2)
Results and Discussion
Characterization of wastewater. Characteristics of textile industry wastewater used at
the present study were: COD between 3,500 and 9, 000 mg/L, Color 4,000 to 10,000
units Pt-Co, pH range 4 - 9 (7.2 in average). In 2005, Badani et al. recorded COD values
in the range of 1280-5600 mg/L in wastewater from textile industry,
Remotion of color and Cod by Electrooxidation: experimental design and data
analysis. Results of color and COD removal are shown in Table 1. Also, it shows that
T3 and T7 are the experiments where the higher COD and color removal is obtained.
TABLA 1. Treatments performed in electro-oxidation
851
Treatment
T1
T2
T3
T4
T5
T6
T7
T8
T9
T10
T11
Operation
Time
(min)
204.85
120
120
180
35.1472
180
120
120
60
120
60
Current
Intensity
(A)
7.5
3.96
11.03
10
7.5
5
7.5
7.5
10
7.5
5
%R
Color
%R
COD
84.12
72.43
94.52
89.10
55.93
89.14
96.92
90.48
78.12
92.48
72.43
60.99
36.62
77.25
73.20
34.01
56.70
59.22
58.42
55.01
53.94
44.77
Figure 1 and 2 show color and COD removal under different operation time and current
intensity. The maximum and minimum values for color removal were 96.92 and a 55.93
respectively with R2 = 0.8812; the maximum and minimum values for COD removal were
77.25 and 34.01 respectively with R2 = 0.8812.
FIGURE 1. Effect of current intensity and operation time on color removal
In 2007, Korbath and Tanyolac studied an electrochemical oxidation of simulated
textile wastewater with iron electrodes. They reported COD, color and turbidity removals
852
around 93.9%, 99.5%, and 82.9%, respectively. The electrochemical treatment of
industrial textile wastewater was optimized using response surface methodology (RSM),
where applied potential and electrolyte concentration were minimized while COD, color
and turbidity removal percent were maximized at 100% pollution load. The optimal COD,
color and turbidity removals, reported by them, was 61.6%, 99.6% and 66.4%,
respectively, whereas COD removal obtained in the present study was 20% higher than
found it by Korbath and Tanyola. However, color removal is similar still above 90% in
both studies.
FIGURE 2. Effect of intensity and time on COD removal
Palma et al. (2013) studied the electrochemical degradation of indigo carmine on
anodes Ti / IrO2-SnO2-Sb2O5, in a cylindrical reactor of 150 ml. The color was rapidly
removed entirely and the COD dropped by 75% from its initial value (336 mg / L O2) in 4
hours. COD (90%) removal achieved in the present study were higher than obtained
(75%) by Palma et al (2013). However, both studies prove the feasibility and efficiency
of electrochemical oxidation as attractive for the treatment of water from the textile
technology.
Current intensity and operation time affect directly (positive effect) color removal.
However, after optimum point, the effect of both variables in color removal is negative. In
case of COD removal, the effect of both variables is positive, in other words the higher
the current intensity and higher the operation time the higher the COD removal (Figure 3
and 4).
853
FIGURE 3. Pareto Chart for significant variables for color removal
FIGURE 4. Pareto Chart for significant variables for COD removal
The optimal conditions obtained were: operation time 151.657 minutes and current
intensity 10 A. Under this conditions the optimal COD and color removal were estimated:
95.2% for color and 73.237% for COD (Figure 5 and 6).
854
FIGURE 5. Effect of operation time and current intensity on color removal
FIGURE 6. Effect of operation time and current intensity on COD removal
Economic aspect. The cost for energy consumption is shows on Table 2. Treatment T3
and T7 are the highest color and COD removal, but they are not represent the lowest
cost (12.74 and 5.99 dollars per m3, respectively). However, treatment T7 could be
chosen as the best of eleven treatments.
TABLE 2. Cost of treatment in different conditions of energy consume
855
Treatment Operation Current
time
intensity
(min)
(A)
Ti
204.85
7.5
T2
120
3.96
T3
120
11.03
T4
180
10
T5
35.1472
7.5
T6
180
5
T7
120
7.5
T8
120
7.5
T9
60
10
T10
120
7.5
T11
60
5
Voltage
(V)
Vs
(m3)
12.7
8.95
18.2
15.8
13.6
9.85
12.6
9.65
12.3
13.7
9
0.00125
0.00125
0.00125
0.00125
0.00125
0.00125
0.00125
0.00125
0.00125
0.00125
0.00125
Energy
Cost
Consumption (Dollar m-3)
(kWhm-3)
260.16
10.32
56.77
2.25
321.35
12.74
379.2
15.04
47.8
1.89
118.2
4.68
151.2
5.99
115.8
4.59
98.4
3.9
164.4
6.52
36
1.42
The optimal experiment or treatment cost was also calculated, which would have a
energy consumption of 284.10 kWh m-3 and a final cost of 11.27 dollars per m 3 of
treated effluent.
Conclusion
Using electro-oxidation system 81.36% and 93.12% for COD and color removal were
obtained. Under optimal conditions 95.2 % and 73.237% COD and color removal would
be achieved, with a cost of 11.27 dollars per m 3 treated. However, treatment T7
demonstrates greater economic viability with a cost of 5.99 dollars per m 3 treated
(96.92% and 59.22 of color and COD removal, respectively). The system does not meet
the values that establish the Mexican Official Standards (SEMARNAT, 2003a) for
discharging wastewaters into national waters body and water resources (SEMARNAT,
2003b).
856
References
Badani, H. Ait-Amar, A. Si-Salah, M. Brik, W. Fuchs (2005). Treatment of textile waste water by
membrane Bioreactor and reuse. Desalination 185 (2005) 411±41.
%DKDGÕU . .RUEDKWL $EGXUUDKPDQ 7DQ\RODF Electrochemical treatment of simulated textile
wastewater with industrial components and Levafix Blue CA reactive dye: Optimization through
response surface methodology. Journal of Hazardous Materials 151 (2008) 422±431
Comisión
Federal
de
Electricidad
(CFE)
(2013-2014).
http://app.cfe.gob.mx/Aplicaciones/CCFE/Tarifas/Tarifas/Tarifas_industria.asp?Tarifa=CMAA&Anio
=2014
Kural E., Koyuncu & Topack D. (2001). Pilot-scale, nanofiltration membrane separation for wastewater
management in the textile industry. Water Sci. Technol., 43 (10) 233±240.
Martínez-Huitle, Carlos A., Brillas, Enric (2009). Decontamination of wastewaters containing synthetic
organic dyes by electrochemical methods: A general review. Applied Catalysis B: Environmental 87
(2009) 105±145
Palma Goyes, R., Macías Uribe, J., González, I., Torres- Palma, R. (2013). Tratamiento de aguas
residuales provenientes de la industria textil mediante oxidación electroquímica. Revista
Colombiana de Materiales Vol.4, pag. 93 ± 108.
Poon CS, Huang QP, Fung C (1999) Degradation of cupropheny Yellow RL by UV/H2O2/Ultrasonication
(US) process in aqueous solution. Chemosphere 38:1005±1014.
Secretaría de Economía (SE) (2014). http://portalweb.sgm.gob.mx/economia/es/tipos-de-cambio/449tablas-peso-mexicano-us-dolar.html
Secretaría de Medio Ambiente y Recursos Naturales (SEMARNAT) (2003a). Reforma de la Norma Oficial
Mexicana 001 (NOM-001-SEMARNAT-1996). Diario Oficial de la Federación 23 de abril de 2003,
México, D. F.
Secretaría de Medio Ambiente y Recursos Naturales (SEMARNAT) (2003b). Reforma de la Norma Oficial
Mexicana 003 (NOM-003-SEMARNAT-1997). Diario Oficial de la Federación 23 de abril de 2003, ,
México, D. F.
U.S. EPA, Pollution prevention studies in the textile wet processing industry, Virginia, USA, 1995.
857
CHAPTER 7.18. REMOCIÓN DE METALES PESADOS EN SOLUCIÓN ACUOSA
EMPLEANDO LA BIOMASA DE LA PERSEA AMERICANA
Díaz-Muñoz Lizbeth Liliana (1); Bonilla-Petriciolet Adrián* (1)
(1) instituto Tecnológico de Aguascalientes, Av. Adolfo López Mateos #1801 Ote. Fracc. Bona Gens C.P.
20256, Aguascalientes, México.
RESUMEN
En el presente estudio, la biomasa de la Persea Americana (BPA) se ha empleado, en
su forma natural y modificada con ácido tartárico (AT), como sorbente para la remoción
de Cd2+ y Zn2+ en solución acuosa. Para la modificación química de esta biomasa se
realizó un diseño de experimentos basado en un arreglo ortogonal L 16 donde se estudió
el efecto de las variables de impregnación de la BPA sobre la capacidad de remoción
de metales pesados en el agua. Estas variables fueron: relación masa ± volumen de la
BPA y la solución de AT, concentración del ácido tartárico, temperatura de la
modificación química y tiempo de contacto. Los experimentos de remoción se realizaron
con soluciones mono-componente de estos metales a 30 °C, 150 rpm y 24 h. Los
resultados indican que la BPA en su forma natural presenta una capacidad máxima de
remoción de 5.60 mg/g para el Cd2+, mientras que para el Zn2+, la capacidad de
remoción máxima es 2.59 mg/g. Por otra parte, la BPA modificada con AT mostró una
capacidad de remoción de 10.93 y 7.41 mg/g para el Cd 2+ y Zn2+, respectivamente. En
general, la capacidad de remoción de metales pesados utilizando la BPA modificada
con AT puede aumentar hasta un 195 y un 286% para el Cd2+ y el Zn2+,
respectivamente, con respecto al material en su forma natural. Estos resultados indican
que la BPA empleada en este trabajo es un sorbente alternativo, de bajo costo y
prometedor para remover iones de Cd2+ y Zn2+ en soluciones acuosas.
Palabras clave: contaminación del agua, metales pesados, persea americana, sorción.
ABSTRACT
In this study, the biomass of Persea Americana (BPA) has been used, in its natural form
and modified with tartaric acid (AT), as a sorbent for the removal of Cd 2+ and Zn2+ in
aqueous solution. An orthogonal array L16 was used for the chemical modification of this
biomass and to study the effect of AT modification process on the metal uptake of BPA.
The variables used in this experimental design were the mass-volume ratio of BPA and
AT solution, the tartaric acid concentration, the chemical modification temperature and
the contact time. Removal experiments were performed using mono-component
solutions of these metals at 30 °C, 150 rpm and 24 h. Results indicate that the maximum
-----------------*Autor de correspondencia: [email protected]
858
metal uptake for the BPA in its natural form was 5.60 mg/g for Cd 2+ and 2.59 mg/g
for Zn2+. On the other hand, the modified BPA showed metal uptakes of 10.93 and
7.41 mg/g for Cd2+ and Zn2+, respectively. In general, heavy metal uptakes using
BPA modified with AT increased until 195 and 286 % for Cd 2+ and Zn2+. In
conclusion, our results indicated that the BPA is a low cost and promising sorbent
for the removal of Cd2+ and Zn2+ ions in aqueous solutions.
Key words: heavy metals, persea americana, pollution water, sorption.
Introducción
La contaminación del agua por metales pesados es considerado un problema
ambiental relevante ya que puede causar diversos trastornos y enfermedades a
los seres vivos y humanos tales como daños al riñón, cáncer pulmonar, malestar
gastrointestinal, entre otros (Srivastava et al., 2009; Asberry et al., 2014).
Actualmente se dispone de diversos métodos de remoción de metales pesados,
entre los que se encuentran el intercambio iónico, ósmosis inversa, precipitación
química y la sorción (Sprynskyy et al., 2006; Wan et al., 2008; Miretzky et al.,
2010). El proceso de sorción presenta diversas ventajas operativas y ecónomicas
entre las que se encuentra una alta eficiencia de remoción, fácil operación y bajo
costo (Juang et al., 2000; Gupta et al., 2012; Bernardo et al., 2013). En los últimos
años se han buscado y utilizado diversos materiales lignocelulósicos como
precursores de sorbentes para la remoción de metales pesados. Estos materiales
incluyen al bagazo de azúcar, cáscara de arroz, cáscara de granada, cáscara de
plátano, cáscara de almendra, entre otros (Yeneneh et al., 2011; Asberry et al.,
2014).
Diversos estudios han demostrado que la capacidad de remoción de metales
pesados de un material lignocelulósico es mayor para un material modificado con
agentes químicos en comparación con la obtenida para el material en su forma
QDWXUDOâüLEDQHWDO3HKOLYDQHWDO9HOD]TXH]-Jimenez et al., 2013).
Considerando lo anterior, en este proyecto la biomasa de la Persea Americana fue
utilizada como sorbente para la remoción de metales pesados. Se estudió el
desempeño de este residuo lignocelulósico, en su forma natural y modificado con
ácido tartárico, para establecer sus propiedades de remoción de metales pesados.
Materiales y metodología
En primera instancia, se realizó la limpieza y reducción de tamaño de la BPA.
Posteriormente, esta biomasa fue lavada con agua desionizada a 150 rpm y 30 °C,
y secada a 50 °C durante 24 h. Se realizó la reducción de tamaño y un proceso de
tamizado con la finalidad de homogeneizar el tamaño de la partícula del residuo.
Se realizó la modificación química de la BPA con ácido tartárico. Para esta etapa
se utilizó un diseño de experimentos basado en un arreglo ortogonal L16 donde se
estudió el efecto de las variables de impregnación de la BPA sobre la capacidad
de remoción de metales pesados en el agua. La variable de respuesta fue la
capacidad de remoción de metales pesados del material modificado. En la Tabla 1
859
se muestra el diseño experimental utilizado, el cual incluye los niveles de las
variables estudiadas.
TABLA 1. Arreglo ortogonal L16 utilizado para la modificación de la BPA
empleando ácido tartárico
Factores
Experimento
Concentración,
M
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
0.5
0.5
0.5
0.5
1
1
1
1
1.5
1.5
1.5
1.5
2
2
2
2
Relación
masa/volumen,
g/mL
1/5
1/10
1/15
1/20
1/5
1/10
1/15
1/20
1/5
1/10
1/15
1/20
1/5
1/10
1/15
1/20
Temperatura,
°C
60
90
120
150
90
60
150
120
120
150
60
90
150
120
90
60
Tiempo de
contacto,
h
1
2
3
8
3
8
1
2
8
3
2
1
2
1
8
3
Se realizaron estudios de equilibrio de sorción de metales pesados utilizando
la BPA en su forma natural y modificada con ácido tartárico. Los experimentos se
realizaron a 150 rpm, 30 °C, pH 5, con una relación masa-volumen de 0.01 g/mL y
un tiempo de equilibrio de 24 h. Inicialmente, se realizaron las isotermas de
sorción para los metales Cd2+ y Zn2+ con la BPA en su forma natural. Una vez
alcanzadas las 24 h de contacto solución ± sorbente, la BPA se separó de la
solución metálica y las concentraciones de los metales se cuantificaron en un
espectro de absorción atómica. Para evaluar la capacidad de remoción de la BPA
modificada con AT se utilizaron soluciones de Cd 2+ y Zn2+ con una concentración
de 250 ppm, 30 °C, pH 5, relación masa-volumen de 0.01 g/mL y un tiempo de
equilibrio de 24 h.
860
Resultados
En la Figura 1 se muestran las isotermas de sorción para los iones metálicos de
Cd2+ y Zn2+ utilizando la BPA en su forma natural. La capacidad de remoción de
este material varía entre 1.47 - 5.60 mg/g y 0.98 - 2.59 mg/g para el Cd2+ y Zn2+,
respectivamente. Con respecto a los experimentos de remoción utilizando la BPA
modificada con AT, en la Figura 2a se muestra la capacidad de remoción para el
Cd2+ utilizando las 16 muestras modificadas y la capacidad de remoción de la BPA
HQ VX IRUPD QDWXUDO /D PXHVWUD GHQRPLQDGD ³´ UHSUHVHQWD HO PDWHULDO VLQ
modificar, el cual tiene una capacidad de remoción máxima de 5.60 mg/g. Se
puede observar que la capacidad de remoción del material modificado puede
incrementar hasta un 195% con respecto al material natural dependiendo de las
condiciones de modificación química utilizadas. La muestra número 9 presenta la
mayor capacidad de remoción del metal y corresponde a 10.93 mg/g. En la Figura
2b se muestra la capacidad de remoción para el Zn 2+ para la biomasa modificada
empleando el diseño experimental. La BPA sin modificar presenta una capacidad
de remoción máxima de 2.59 mg/g mientras que la muestra número 9 presenta la
mayor capacidad de remoción que es 7.41 mg/g.
6
2.5
b)
a)
5
2.0
Q (mg/g)
Q (mg/g)
4
3
1.5
1.0
2
0.5
1
0.0
0
0
0
50
100
150
200
250
50
100
150
200
Ce (mg/L)
Ce (mg/L)
FIGURE 1. Isotermas de sorción para a) Cd2+ y b) Zn2+ utilizando BPA en su forma
natural.
861
a)
8
b)
7
10
6
8
Q (mg/g)
Q (mg/g)
5
6
4
4
3
2
2
1
0
0
0
1
2
3
4
5
6
7
8
9 10 11 12 13 14 15 16
Número de Muestra
0
1
2
3
4
5
6
7
8
9 10 11 12 13 14 15 16
Número de Muestra
FIGURE 2. Resultados obtenidos para la capacidad de remoción de a) Cd 2+ y
b) Zn2+ empleando BPA en su forma natural y modificado con ácido tartárico.
Conclusión
De acuerdo a los resultados obtenidos, se observa que la capacidad de remoción
de metales pesados utilizando la BPA modificada con AT puede aumentar hasta
un 195 y un 286 % para el Cd2+ y el Zn2+, respectivamente, con respecto al
material en su forma natural. Estos resultados demuestran que el proceso de
modificación química permite incrementar sustancialmente las propiedades de
sorción de esta biomasa. Esta biomasa BPA puede ser una alternativa viable, de
bajo costo y prometedora para su aplicación en la remoción de iones de Cd 2+ y
Zn2+ en soluciones acuosas.
Agradecimientos
Al Instituto Tecnológico de Aguascalientes y CONACYT por el apoyo otorgado
para la realización del presente estudio.
Referencias
Asberry H. B., Kuo C. Y., Gung C. H., Conte E. D., Suen S. Y. (2014). Characterization of water
bamboo husk biosorbents and their application in heavy metal ion trapping. Microchemical
Journal (113), 59 ± 63.
Bernardo M., Mendes S., Lapa N., Gonऊalves M., Mendes B., Pinto F., Lopes H., Fonseca I. (2013).
Removal of lead (Pb2+) from aqueous medium by using chars from co-pyrolysis. Journal of
Colloid and Interface Science (409), 158 ± 165.
Gupta V. K., Ganjali M.R., Nayak A., Bhushan B., Agarwal S. (2012). Enhanced heavy metals
removal and recovery by mesoporous adsorbent prepared from waste rubber tire. Chemical
Engineering Journal (197), 330 ± 342.
Juang R. S., Shiau R. C. (2000). Metal removal from aqueous solutions using chitosan-enhanced
membrane filtration. Journal of Membrane Science (165), 159 ± 167.
Miretzky P., Fernandez Cirelli A. (2010). Cr(VI) and Cr(III) removal from aqueous solution by raw
and modified lignocellulosic materials: A review. Journal of Hazardous Materials (180), 1 ±
19.
862
3HKOLYDQ ( $OWXQ 7 3DUOD\LFL ù 0RGLILHG EDUOH\ VWUDZ DV D SRWHQWLDO ELRVRUEHQW IRU
removal of copper ions from aqueous solution. Food Chemistry (135), 2229 ± 2234.
6SU\QVN\\0%XV]HZVNL%7HU]\N$31DPLHãQLN-6WXG\RIWKHVHOHFWLRQPHFKDQLVP
of heavy metal (Pb2+, Cu2+, Ni2+, and Cd2+) adsorption on clinoptilolite. Journal of Colloid
and Interface Science (304), 21 ± 28.
âüLEDQ 0 .ODãQMD 0 âNUELü % $GVRUSWLRQ RI FRSSHU LRQV IURPV ZDWHU E\ PRGLILHG
agricultural by products. Desalination (229), 170 ± 180.
Srivastava V. C., Mall I. D., Mishra I. M. (2009). Competitive adsorption of cadmium(II) and nickel(II)
metal ions from aqueous solution onto rice husk ash. Chemical Engineering and Processing
(48), 370 ± 379.
Velazquez-Jimenez L. H., Pavlick A., Rangel-Mendez J. R. (2013). Chemical characterization of
raw and treated agave bagasse and its potential as adsorbent of metal cations from wáter.
Industrial Crops and Products (43), 200 ± 206.
Wan Ngah W.S., Hanafiah M.A.K.M. (2008). Removal of heavy metal ions from wastewater by
chemicall modified plant wastes as adsorbents: A review. Bioresource Technology (99), 3935
± 3948.
Yeneneh A. M., Maitra S, Eldemerdash U. (2011). Study on Biosorption of Heavy Metals by
Modified Lignocellulosic Waste. Journal of Applied Sciences (11), 3555 ± 3562.
863
CHAPTER 7.19. CADMIUM REMOVAL CAPACITY FROM WATER AND ITS
EFFECT ON Schoenoplectus pungens
Carmen Méndez-Hurtado (1); Edith Flores-Tavizón* (1);
Juan Pedro Flores-Marguez (2); Sergio Saúl Solís (1);
Laura Yáñez-Espinosa (3)
IIT-UACJ Ciudad Juárez Chihuahua, México
ICB-UACJ Ciudad Juárez Chihuahua, México
IIDZ-UASLP San Luis Potosí, San Luis Potosí, México ABSTRACT
The constant anthropogenic activities significantly increased the concentration of
heavy metals, such as cadmium (Cd), in human ecosystems. As a result of
activities such as mining, industry, use of fossil fuels or household waste,
application of phosphate fertilizer and waste incineration, Cd has been pushed in
the atmosphere, giving it the way to travel great distances to be subsequently
deposited on the ground or water. The presence of high concentrations of Cd in the
environment has a negative effect on the metabolism of plants, with the generation
of reactive oxygen species and blocking essential functional. Schoenoplectus
pungens is a species of flowering plant in the sedge family known by the common
name Common Three-Square, is used for the stabilization or recovery of metals
from contaminated areas; their roots grow easily in the presence of abundant water
so it is an excellent choice for soil stabilization in wetlands and clean sites. Also
meets the requirements to be used in hydroponics techniques because it grows
easily without soil and is tolerant to pH acid characteristics. In this work we
determined the capacity for Cd removal from hydroponic solution at concentrations
of 0.2, 0.5, 1.0, 3.0 and 5.0 ppm by the Schoenoplectus pungens species, as well
as damage caused by metal on vegetative structure, chlorophyll content and
functional groups present. The observed removal of Cd from the solutions by the
plant was more than 90%, being it accumulated in the plant. Finally, Cd affects
chlorophyll content which is inversely proportional to the concentration of the metal, because Cd in their ionic form replaces to magnesium on the chlorophyll molecule;
while the decreasing the functional groups is because the Cd bind to these groups
or is substituted to other elements.
Keywords: cadmium; Schoenoplectus pungens; removal capacity; water
-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐-‐ *Author for correspondence: [email protected]
864
Introduction
Cadmium (Cd) is a little heavy metal abundant naturally in the earth's crust,
however in recent decades has greatly increased its concentration, as a result of
mining and industrial activity, since cadmium is considered a by-product of mining,
smelting and refining of zinc and to a lesser extent of lead and copper (RodríguezSerrano et al 2008). As a result of activities such as mining, industry, use of fossil
fuels or household waste, application of phosphate fertilizer and waste incineration, Cd has been pushed in the atmosphere, giving it the way to travel great distances
to be subsequently deposited on the ground or water. (ATSDR, 2012). In the air is
as oxide, chloride or sulphate as particles or vapors from high temperature
processes. They can also move through the soil, but their mobility depends on
several factors such as pH and organic matter, which vary according to the local
environment.
Generally, Cd binds strongly to organic matter which remains immobile in soil
and be taken up by plants, thus entering the food chain. The present Cd in water
exists in the hydrated form as ion or associated with other inorganic or organic
substances ionic complex. Soluble forms are mobilized in water and insoluble
forms are immobile and deposited in the sediment where they will be adsorbed
(ATSDR, 2012; Tejada 2010). Currently, the presence of high concentrations of cadmium in the environment
having a negative effect on the metabolism of plants, affecting mainly due to their
viability as altered mechanisms: generation of reactive oxygen species, blocking
essential functional groups in biomolecules. The Cd as a member of the class B
transition metals, has high affinity for COOH, OH, NH2, SH groups mainly indole
and iminazol (Anfinsen et al., 1991; Polle and Schutzendubel, 2003).
The reduction in the growth of the radicle seedlings and root length in adult
plants (Ferjani, 2005; Pomponi et al., 2006; Méndez 2010, Liu et al., 2013), as well
as inhibition of stomatal opening (Barcelo et al., 1986), fotosíntesis (Pietrini et al,
2003; Drazkiewicz et al., 2003) and the synthesis of chlorophyll (Drazic y
Mihailovic, 2005, Quin et al., 2013), was also reported inhibition of pollen
germination and pollen tube growth (Xiong y Peng, 2001), carotenoids decrease (Raí
et al., 2005) and the transpiration rate (Xiong y Peng, 2001, Liu et al., 2013), accelerating cellular senescence (Di Toppi y Gabbrielli, 1999; Benavides et al.,
2005; Méndez-Hurtado el al., 2013), and chlorosis (Prasad, 2003; Méndez 2010,
Liu et al., 2013; Quin et al., 2013). In addition to the interference with the taking,
transportation and use of various micronutrients such as Fe, Mn and Zn (Di Toppi y
Gabbrielli, 1999; Liu et al., 2013), and DNA polymorphism (Liu et al; 2005).
Schoenoplectus pungens is a species that has been used for the stabilization
or recovery of degraded or disturbed areas, added to this, the roots grow easily in
the presence of abundant water so it is an excellent choice for soil stabilization in
wetlands and riparian sites. It complies with the requirements to be used in
hydroponics techniques, because it is a plant that grows easily without soil besides
being tolerant to acid pH (Canché, 2012).
Therefore, the objectives to this work were: 1) determine the reduction of
functional groups present in the cells of the species Schoenoplectus pungens
using FTIR, 2) Evaluate the physiological damage caused by the Cd to the
865
species, 3) determine the removal of Cd by the species, and 4) Evaluate the
chlorophyll content.
Schoenoplectus pugens specimens were placed in hydroponics treatments
(growing technique in aqueous solutions). The roots are immersed in an aerated
Hoagland nutrient solution (Meraz, 2010), Cd were added for experimentation.
Experiments with five different concentrations of Cd (0, 0.2, 0.5, 1, 3, 5 ppm) were
made, concentrations found in natural environments in addition to being the limits
set by EPA and SEMARNAT (SEMARTNAT-NOM-127, 03, Razo 2006). After the
process of hydroponics, plants are harvested and dried to make the physiological
analysis, where took weight, radicle length because it indicates the ability to grow
and develop, and the ability to establish and grow for to thrive to stress.
Furthermore fragments of leaves were obtained at the beginning, middle and end
of the experiment, with the purpose of monitor the decrease of chlorophyll.
Fresh plant material was weighed, 0.5 g, and ground in a porcelain mortar
adding 2 ml of 80% acetone with the purpose of extract total chlorophyll. Analyzed
by UV-visible spectroscopy at 645, 652 and 663 nm. For the total amount of
chlorophyll per gram of the sample the following equations were used.
1) Total Chlorophyll (mg / L) = 27.8 * A652nm
2) mg chlorophyll = vol. capacity (L) * Total chlorophyll (mg / L) / 1L
3) mg chlorophyll / g sample = mg chlorophyll *1g sample / g sample
The presence of functional groups was also determined by Fourier transform
Infrared Spectroscopy (FTIR) Thermo Nicolet 6700, with the use of a zinc selenide
crystal. Scans of 1/32 seconds were performed by IR technique to identify the
different organic compounds Schoenoplectus pungens species possesses, as
carboxylic acids, amines, amides, aldehydes, sulfur groups, etc., with which the Cd
could interact by binding to its structure and to facilitate entry to the plant. This due
to it has been reported that the Cd has a high affinity for sulfhydryl groups, as well
as by amino groups (Parson et al., 2013). To quantify the ability of Cd removal by the species, seedlings were
harvested and analyzed chemically. Then proceeded to dry the plants at 45 ° C for
48 hours. Were placed 0.5 g of the portion of the plant (stem, root, leaves) in flasks
to HDPT digestions were made in a microwave (CEM MARS XPRESS). This
procedure is done in triplicate. And it proceeded to acid digestion made. At the end
of the acid digestion, the products were analyzed by atomic absorption
spectrophotometer, obtaining the concentrations of Cd in seedlings. To quantify the
removal of Cd in water, was made after treatment of hydroponics, which lasted a
period of 15 days established, proceeded to collect water samples. Samples were
took in the following way: during the first 5 days were taken daily 50 ml of
hydroponic solution and then just sample was taken on days 7, 9, 11, 13 and 15.
After taking water samples, the solid adhered to the container was removed to
866
have clean sample residues. After 50 ml were taken from each sample and were
put into glass beakers of 250 ml. They were then placed 3 ml of reagent grade
nitric acid HNO3 Fisher. Immediately were placed in warming rack type Thermoline
2200-65 °C, they care that the temperature did not rise to prevent the loss of metal.
The beakers glass were placed watch glass with water on top to prevent
evaporation and improve digestion. Under these conditions are maintained for 2
hours and then 5 ml of HNO3 was added for 5 hours more.
In Table 1 shows the ratio of fresh weight, height and root elongation in the
different treatments hydroponics. Data at the beginning and end of the experiment
are observed. Observing that, the most lethal Cd concentrations for the species
were 3 and 5 ppm.
TABLE 1. Physiological responses to Cd
0 day
15 days
Concentration
of Cd (ppm)
Plant height
(cm)
SD
Weight (gr)
SD
0
0.2
0.5
1
3
5
0
0.2
0.5
1
3
5
68
69
71
70
70
69
83
78
76
73
70
70
0.0010
0.0012
0.0013
0.0020
0.0021
0.0022
0.0012
0.0015
0.0020
0.0014
0.0022
0.0018
203
205
204
200
204
206
456
389
380
259
216
187
0.0025
0.0038
0.0026
0.0024
0.0025
0.0024
0.0016
0.0016
0.0017
0.0023
0.0033
0.0038
Root
elongation
(cm)
6
6
7
6
7
6
15
13
12
10.5
9
7
SD
Stolons
numbers
SD
0.0012
0.0016
0.0019
0.0014
0.0023
0.0025
0.0010
0.0026
0.0031
0.0026
0.0028
0.0026
0
0
0
0
0
0
29
24
20
16
12
9
0
0
0
0
0
0
0.0024
0.0031
0.0016
0.0025
0.0032
0.0012
Notes: Effects of Cd on the physiological development of the species Schoenoplectus pungens.
Data at the beginning and end of the experiment in their respective concentrations and
physiological variables are shown.
Cd concentrations shown in Figure 1, was decreasing as advanced the day,
so that the lower concentrations were removed entirely from the water, while the
concentrations of 3 and 5 ppm decreased concentration such that the Cd would
have no toxic effects on the ground.
867
FIGURE 1. Cd removal per day and Cd decreasing in water were
demonstrated.
Cd removal by plant is demonstrated because was found Cd in the tissue of
Schoenoplectus pungens. The plant removes higher concentrations of Cd, 3 and 5
ppm, in water almost entirely.
The concentration of chlorophyll determination in leaves was performed. The
results are shown in Figure 2. The chlorophyll analysis extended over 15 days of
experimental treatment because although the roots seemed to be further
developed, most of the leaves of the highest concentrations almost completely
dried. This is why chlorophyll was measured at the beginning, middle and end of
the experiment. Can be seen that the toxic effect of Cd to the species
Schoenoplectus pungens were exposed, can be determined by a clear decrease in
the concentration of chlorophyll at 7 and 15 days.
FIGURE 2. Total concentration of chlorophyll in function of time.
Hydroponics experiment.
In the analysis Fourier Transform Infrared Spectroscopy (FTIR) in leaf / stem,
the appearance of functional groups such as amines and sulfurized groups was
868
observed, which shows that with increasing concentrations the peak intensity
decreases; it could observe that in the concentration of 5 ppm of Cd are reduced
drastically peaks. Also in the FTIR of root is shown a marked decrease of the
functional groups. The peaks in the Cd concentration of 5 ppm were decreasing.
Peaks of the amine group to treatment compared to the control are observed
clearly, shows all well-defined functional groups.
Conclusion
In this study, so far obtained results, show that the presence of Cd is affecting
species Schoenoplectus pungens; low Cd concentrations as 0.2, 0.5 and high Cd
concentrations as 1, 3 and 5 ppm, clearly shows symptoms of intoxications in
plants. The most damage on the medium, in addition to much smaller numbers
stolons in root because the root was dying, was at 5 ppm of Cd. The chlorophyll is
decreased due to substitution of magnesium (Mg) by Cd in the plant structure, lose
their green color. This symptom was most noticeable in concentrations 1, 3 and 5
ppm, 5 ppm being the most toxic concentration. Also the root elongation was more
marked at low concentrations as well as 0.2 and 0.5 and too for the control and this
is because these concentrations are permissible limits and are tolerated by living
organisms, for that reason 0.2 and 0.5 were less toxic to the plant. Regarding the
reduction of functional groups, FTIR analyzed in this is primary, because when the
Cd enters to the plant several molecules containing these functional groups bind to
mitigate the stress of the metal, so that the reduction of these functional groups
appears.
Acknowledgements
Scholarship CONACYT-221805, UACJ, Dr. Edith Flores Tavizón, Dr. Laura Yanez,
Dr. Juan Pedro Flores Margez.
References
Agency for Toxic Substances & Disease Registry. ToxFAQs for: Cadmium. (2012).
http://www.atsdr.cdc.gov. ATSDR
Barceló J., Poschenrieder C., Andreu I. & Gunse B. (1986). Cadmium-induced decrease of water
stress resistance in bush bean plants (Phaseolus vulgaris L. Contender). I. Effects on water
potential, relative water content and cell wall elasticity. Journal Plant Physiology 125, 17-25
Drazic G. & Mihailovic N. (2005). Modification of cadmium toxicity in soybean seedlings by salicylic
acid. Plant Science 168, 511-517.
Drazkiewicz M., Tukendorf A. & Baszynski T. (2003). Age-dependent response of maize leaf
segments to cadmium treatment: Effect on chlorophyll fluorescence and phytochelatin
accumulation. Journal of Plant Physiology 160, 247-254.
Di Toppi L. & Gabbrielli R. (1999). Response to cadmium in higher plants. Environmental and
Experimental Botany 41, 105- 130.
869
Duchenkov V., Kumar N., Motto H., Rakin L. (1995) Rhizofiltation: the use of plants to remove
heavy metals from aqueous stream. Environ. Sci. technol 29: 1239-1245
Liu M., Li P.J., Qi X.M., Zhou Q.X., Zheng L., Sun T.H. & Yang Y.S. (2005). DNA changes in barley
(Hordeum vulgare) seedlings induced by cadmium pollution using RAPD analysis.
Chemosphere 61, 158-167.
Méndez Hurtado C. A. (2010). Efecto de Cd disuelto en agua en la germinación de semillas y el
crecimiento de plántulas de Agave lechuguilla provenientes de zonas con diferente nivel de
contaminación por metales. Tesis de maestría del Instituto Potosino de Investigación
Científica y Tecnológica (IPICYT)
Polle A. & Schützendübel A. (2003). Heavy metal signalling in plants: linking cellular and organismic
responses. Topics in Current Genetics 4, 187- 214.
Pomponi M., Censi V., Di Girolamo V., De Paolis A., Di Toppi L., Aromolo R., Costantino P. &
Cardarelli M. (2006). Overexpression of Arabidopsis phytochelatin synthase in tobacco plants
enhances Cd2+ tolerance and accumulation but not translocation to the shoot. Planta 223,
180-190.
Quin Q, Wu H, Zhang Y, Feng Q, Tai P. (2013). CHARACTERIZATION OF CADMIUM 108 Cd
distribution and acumulation in Tagetes erecta L. seedlings: effects of split-root and of
remove xylem phoem Chremospher 93: 2284-228
Razo S.I. (2006). Identificación de áreas prioritarias de restauración de suelo contaminado por
arsénico y metales pesados en el sitio minero y metalúrgico de Villa de la Paz-Matehuala,
SLP (México). Tesis de doctorado en ciencias ambientales del Programa Multidisciplinario en
Ciencias Ambientales de la Universidad Autónoma de San Luis Potosí, UASLP.
Rodríguez-Serrano M., Martinez de la Casa N., Romero Puestas MC., Rio Legazpi LA., Sandalio
Gonzalez LM. (2008) Toxicidad del cadmio en plantas. Ecosistemas 17, 13-15
Xiong Z.T. & Peng Y.H. (2001). Response of pollen germination and tube growth to cadmium with
special reference to low concentration exposure. Ecotoxicology and Environmental Safety 48,
51-55.
Zhuang P., McBride M., Xia H., Li N., y Li Z. (2009). Health risk from heavy metals via consumption
of food crops in the vicinity of Dabaoshan mine, South China. Science of the Total
Environment, 407, 1551±1561
Notation
Cd
cm
FTIR
ml
ppm
Cadmium
centimeter
Fourier transform Infrared Spectroscopy
Milliliter
Parts per million
870
CHAPTER 7.20. SWINE WASTEWATER TREATED BY AN ANAEROBIC
SECUENTIAL SYSTEM WITH PULSED RECIRCULATION
Germán Giácoman Vallejos* (1); Ileana de los Ángeles González Canché (1);
María del Carmen Ponce Caballero (1)
(1) Universidad Autónoma de Yucatán (UADY), Facultad de Ingeniería, Mérida, Yucatán,
México.
ABSTRACT
The swine industry is one of the most important activities in the state of Yucatan,
Mexico. Currently, it generates 9,428.37 m3 / d of wastewater with high organic
charge and that is mainly disposed to a karstic soil (characteristic of this area of
Mexico) creating a more serious surface and groundwater pollution problem
because of the high permeability. As alternative for the treatment of this
wastewater, a pilot ASBR with four different configurations of recirculation was
evaluated. It was found a 81% efficiency of SCOD removal in the case of few
pulsations and 95% on the TCOD without recirculation. In the case of total
phosphorus it was obtained at average removal of 75% for the 4 cases evaluated.
From the results it is observed that the ASBR with recirculation is an alternative for
the treatment of swine wastewater.
Key words: anaerobic sequencing batch reactor (ASBR), pulsed recirculation,
swine wastewater
Introduction
In Yucatan, Mexico, the pig production is the third income generating activity in the
agricultural sector. The average production 83,000 tons of pork with a commercial
value of around 1.600 million of Mexican pesos situate the State nationally in the
fourth place in production and the first in productivity (PEDAPEY, 2007).
It has been reported that in Yucatan, there are 470 pig farms, among them
304 are not registered at the national ministry of livestock (SAGARPA). The total
stock of hogs in the state was estimated at 670.174 (85,371 breeding stock,
215,034 weaning and 369.769 fattening) which generated daily 3884.78 tons of
excreta, 9428.37 m3 / d of highly polluted wastewater and with an organic loading
of 443,133.39 kg / d measured as COD (Mendez et al, 2009).
From previous research it was found that in the region, wastewater
discharges are a major source of organic pollution of groundwater (Vasquez and
Zapata, 1993). Therefore, the wastewater generated by the swine production
disposed directly to the carstic soil as in Yucatan, represents a serous pollution
problem since the high permeability, make more vulnerable the surface and
groundwater (Graniel et al)
Swine wastewater contains high concentrations of organic matter (over 10000
mg TCOD/L), nutrients (over 1000 mg/L both nitrogen and Phosphor), pathogens
871
(over 1*10+9 NMP), heavy metals (Cu over 3 mg/L and Zn over 54 mg/L) and
suspended solids (over 15000 mg/L). As a result of its high organic matter, oxygen
consumption is increased and its depletion leads to a decrease of the ecosystem
diversity. The high nitrogen and phosphorus concentrations in this wastewater, can
lead to the eutrophication of aquatic ecosystems. Also, the presence of pathogens,
could lead to diseases dissemination.
To treat swine wastewater it has been used anaerobic digestion by different
type of reactors; one of them is the recently employed anaerobic sequential Batch
Reactor (ASBR), normally operated without recirculation. This reactor (ASBR),
unlike other reactors (upflow anaerobic sludge blanket, UASB and expanded
granular sludge bed, EGSB), works in cycles and not in continuous flow (Sung and
Dague, 1992). Each operating cycle is divided into four stages: (1) Food: the
influent is feed to the reactor; (2) Reaction: variable time step according to the
wastewater characteristics, in it occurs to a greater extent, the degradation of
organic matter; (3) Sedimentation: is for stop the agitation and to allow the
decantation of the biomass, letting a clarified effluent; and (4) Download: The
clarified effluent is removed from the reactor.
In the ASBR reactors the agitation is important not only to provide good
mixing and improve the mass transfer from the liquid-phase into the bulk of
biomass, but also to improve the solubilisation of the particulate (with high content
of organic matter), with positive benefits for consumption rate of organic matter
(Zaiat et al. 2004).
According to studies by Zaiat et al. 2004 and de Novaes et al 2010, the
stirring speed plays an important role in the solubilisation of suspended organic
material. There is acceleration in the degradation of the particulate masse
(expressed as COD) with increasing agitation speed that is probably due to the
greater speed of cutting of larger particles and the primary contact between the
suspended organic matter and extracellular enzymes. However, at 1100 RPM in
the laboratory scale experimentation the hydrodynamic behaviour influences the
performance of the reactor, probably caused by the turbulent flow of the liquid
phase, even causing that the system can be oxygenated and operated in anoxic
form (Zaiat et al. 2004). On the other hand, increasing agitation speed 40 to 80 rpm
in a pilot reactor 1 m3 ASBR resulted in destabilization of the system, probably due
to the rupture of the granules due to the high agitation (de Novaes et al 2010).
According to the experience of Ndon and Dague, 1997 and the obtained
Knowledge for the anaerobic process until the time, it was mentioned that usually it
could not achieve efficient removal of organic matter in wastewater with high
organic content, so that it can reach values of less than 1000 mg / L COD,
especially at room temperature and in a short TRH from 1 day to 12 days. For an
ASBR operation when the objective is to maximize biogas production is preferable
a long reaction phase instead of three short phases, obtaining in this form, higher
yield of gas as a result of a more complete degradation of organic matter (Ndegwa
et al . 2008).
In the work of Martin et al. 2006, Liu et al. 2006, Wu and Rodgers 2010 and
Gürtekin 2010, it was reported that in the anaerobic processes, exist phosphorus
precipitation and it represents a significant proportion of the removal or release of
total phosphorus. Wu and Rodgers 2010 reported more than 50% removal of
872
phosphorus achieved in an ASBR. There are extracellular polymeric substances
(EPS) that play an important role in phosphorus removal, as well as for the survival
of organisms accumulating phosphorus, to create cells in dense flocks and finally
for the sedimentation of activated sludge (Martin et to 2006,. Liu et al 2006). Also
the optimum temperature for the biological phosphorus removal (in the presence of
acetate and glucose fed in sequential anaerobic / anoxic batch reactor) was 20 ° C
(Gürtekin, 2010).
The composition of the wastewater has an important influence on the
properties of the granules into a sludge blanket reactor. In fact, the treatment of
certain types of substrates may result in the development of a poor sludge blanket
(Franco et al. 2002). However, the selection of wastewater to be treated is not a
free option; but rather, it is the generated product by the use of water in a process
applied to a production system.
One possible alternative for improving the properties of granular sludge is the
modification of the operating conditions, including environmental and hydraulic
parameters.
Pulsed processes have been used in many chemical engineering units to
improve the mass transfer rate, were first used in separation processes in order to
improve contact between the phases. At present, it has been developed and
adapted for application in various fields, including biochemical reactors (Franco et
al. 2002).
This study evaluated as an alternative for the treatment of swine wastewater
an ASBR with recirculation and under four different sequences of pulsed and rest.
It was considered that the inclusion, of a pulsed sequential intermittent recirculation
in the ASBR allow increasing the contact between the substrate and the
microorganisms through the promotion of short mixing cycles. These short mixing
cycles also minimizes the suspension of solids due to the release of trapped gas
bubbles contained therein. This in fact prevents the formation of flocculent solids
and instead promotes sedimentation of dense material generating a granular
sludge. So the aim of this study was to evaluate the increased removal of organic
matter and nutrients, considering the effect of recirculation in pulsed form in a
sequential anaerobic treatment system (ASBR) a pilot scale wastewater from pig
farms.
Metodology
The experimental unit was located at the production and research unit site of the
Technological Institute of Conkal, Yucatan, Mexico (N21°5'7.52", W89°32'17.99").
The treatment train consisted of mechanical separation of coarse solids through a
drum screen and Imhoff-type settler (Figure 1), which has a retention time of 24
hours. The secondary biological treatment consisted of a sequential batch reactor
(ASBR) (Figure 1).
The ASBR was constructed of a high-density polyethylene with a total
volumetric capacity of 15,000 L. It consisted of a cylindrical structure with a
diameter of 2.2 m and a height of 3.9 m. Initially, it was seeded with sludge
obtained from the settling tank sediments (from swine farm) and ruminal liquid
(from cattle) collected from the slaughter house at the Faculty of Veterinary
873
Medicine and Zootechnique at the Autonomous University of Yucatan
(N20°51'57.91", W89°37'27.57"). The acclimatization
period for the
microorganisms in the sludge was approximately four months. This ASBR was fed
over a time period of approximately 2 minutes, on a daily basis, with 500 L of
wastewater obtained from the settling tank located prior to ASBR process (Figure
1) using a submersible pump (Franklin 10S-CIM, 0.5 hp) at a flow rate of 320
L/min.
The experiments carried out in the ASBR employed a working volume of
11,250 L (having a space of 3,750 L to contain the generated biogas). The
hydraulic retention time (HRT) in this reactor was 22.5 days.
The feed and recirculation pumps that shown in Figure 1, as well as solenoid
valves which allow the inlet and outlet water in each process were controlled by an
automatic on-off programmable logic controller (PLC-Crouzet Millennium 3).
Effluent
FIGURE 1. Diagram of the train of treatment System
Reactor feeding and wasting was conducted over of a period of two hours on
a daily basis, in the first one-hour period the system was at rest to allow for
sedimentation to take place. The subsequent one-hour period allowed the wasting
and feeding of the system. Firstly the electro valve placed for the effluent of the
ASBR was opened and the reactor wasted (discharging only 500 L into a receptor
tank for treated wastewater). This operation took about five minutes and the electro
valve for the effluent of the ASBR was closed. Afterward, a ten minutes rest was
allowed prior to the feeding of the ASBR (500 L) (Figure 1). This feeding was
conducted in approximately 2 minutes, after which the system remained at rest for
the remainder of the second 1-hour period. When this routine was concluded the
874
daily cycles of discharge and feeding of the system was ended. Next, the process
of recirculation in the ASBR was initiated as explained below.
For this study, the operational conditions were varied according to the
configuration presented in Tables 1 and 2. For the anaerobic reactor (ASBR) the
configurations employed are described as: first, the recycle ratio was alternated
consistently providing 1 hour with recirculation and 1 hour without recirculation over
a period of 22 hours (Table 1). To complete the 24 hours cycle, 2 hours per day
were used to allow for sedimentation (settling phase) and for the discharge and
feeding of ASBR system to be carried out as previously described. Under these
operational conditions, the ASBR was subjected to 11 recirculation cycles (1 hour)
daily (TWR) and 13 cycles without recirculation (TWOR) (Table 2). The second and
third configurations differed from the first simply in the time delay between
recirculation events (Table 1). The fourth configuration corresponded to 0 hours of
recirculation (22 hours rest). During the recirculation time (1 hour) of the ASBR, the
flow was pulsed. For that, one submersible recirculation pump (Franklin 9S-CIM,
0.4 hp) was intermittently operated with a flow of 200 L/min at intervals of 3
minutes (3 minutes with recirculation and 3 minutes rest). Hence, for each hour of
recirculation the pump was activated 10 times.
The water quality parameters monitored during ASBR treatment were: total
chemical oxygen demand (TCOD) and soluble chemical oxygen demand (SCOD)
and total phosphorus (TP). The samples were collected at specific monitoring
points including the influent and effluent of the ASBR. For each configuration
tested, 2 L samples were taken over a 6 day monitoring period, following 21 days
of stabilization. For transport to laboratory, the samples were preserved with
sulfuric acid (2 ml x L) and kept at a temperature of 4°C. They were analyzed at the
laboratory the same day of collection.
Since the data obtained naturally exhibited a high degree of variability, a oneway analysis of variance (ANOVA) at a 95% significance level was applied to
determine significant differences among the parameter means for all
configurations. The statistical analyses were conducted employing a software
program Statgraphics Centurion XVI.I.
TABLE 10. Configuration of the pulsed recirculation
Configuration
1
2
3
4
Recirculation time
(h)
1
1
1
0
Rest time
(h)
1
2
3
24
TABLE 11. Cycle of recirculation and rest time in the ASBR
Configuration
Daily Cycle
Ts
Tc
875
Pulsed
Recirculation
cycles (TWR)
Rest Cycles
(TWOR)
(h)
1
11
10
3
2
8
14
2
3
6
15
3
4
0
24
24
Time cycle (Tc), Time for the sedimentation at the end of daily cycle (Ts).
(h)
24
24
24
24
Result and discusion
Removal Efficiency of TCOD [%]
Removal Efficiency of SCOD [%]
Removal efficiencies of TCOD were on average greater than 88% and up to 94%,
and in the case of TSCOD was greater than 59% and up to 79% (Figure 2). In the
case of the SCOD, the more efficient configuration was reported when pulsed
recirculation was not applied and this was mainly due to the long rest time (or time
for sedimentation) and the low presence of suspended material in the effluent;
unlike the SCOD had the highest efficiency corresponded to the third configuration,
which consisted of 1 hour recirculation and 3 hours of rest. As noted above,
agitation had an important role in the ASBR reactors, it was fast enough to stir the
mixture but not so excessive as to cause unwanted aeration process or
defragmentation of the solids granules, which cause a resuspension of solids
giving contribution to the TCOD. This was demonstrated in the results (table 3).
When there is a greater agitation than a lower SCOD removal efficiency and when
there was no agitation the process efficiency of the SCOD was decreased.
Rest Time [h]
a)
Rest Time [h]
b)
FIGURE 2. a) Removal efficiency of TCOD. b) Removal efficiency of SCOD
From the Fig. 3 it is observed, that the higher removal efficiency (81%) for
total phosphorus (PT) was obtained in the first configuration (1 hour recirculation, 1
hour rest). In the case of phosphorous the removal efficiency achieved for all the
configurations showed high values (higher than 66%, even with a temperature
averaging operation of 31 ° C). This could be caused in part by the precipitation of
phosphorus and possibly by the presence of extracellular polymer substances.
876
Removal Efficiency of PT [%]
Rest Time [h]
FIGURE 3. Removal efficiency of PT
Table 3 shows the means of the inputs and outputs of each configuration, it
can be seen that the TCOD of the outputs of each configuration exceeds the value
of 1000 mg / L. This value is expected for a ASBR which operates at room
temperature and has short hydraulic residence times, but in this case, the
residence time was long enough (22,5 days) and this effect might be due to the
nature of the wastewater being treated. In the case of waste of animal manure, the
substrate contains matter difficult to degrade biologically, recalcitrant compounds
as well as components that are trapped in the solid fraction, like fatty acids (volatile
or short chain and the long chain) (Ndegwa et al 2008).
TABLE 12. Averages of the inputs and outputs of the parameters to the
different operating conditions
Configuration
TCOD
mg/L
Affluent
Effluent
SCOD
mg/L
Affluent
Effluent
1
9,892
1,178
%
Removal
88.09
2
15,135
1,171
92.26
2,689
709
3
23,964
2,059
91.41
4,998
1,035
4
25,206
1,348
94.65
1,929
781
ANOVA
p-value
0.2337
2,795
1,059
%
Removal
62.11
PT
mg/L
Affluent
Effluent
885
170
%
Removal
80.79
73.63
943
211
77.62
79.29
1,609
542
66.31
59.51
1,097
242
77.94
0.0069
0.2928
From the applied statistical analysis it can be seen that in the case of TCOD
there in not statistical difference in the removal percentages since the obtained
value was 0.2337 which is higher than 0.05. In the case of SCOD the found pvalue was of 0.006 which is lower than 0.05; this pointed out that the removal
efficiencies of SCOD showed significant differences, making evident the effect of
the pulsed agitation in the removal of such parameter. For the removal efficiency of
phosphorous there is no significant difference since the p-value found was of
877
0.2928 which is higher than 0.05. This corroborates that the pulsations in the
recirculation have no effect in the phosphorous removal.
Conclusion
There is a good removal efficiency of TCOD in all cases of the configurations, with
values greater than 88%. With respect to the SCOD can be concluded that the best
configuration was 1 hour pulsed recirculation and 3 hours of rest, which in turn, that
the least hours stirring has the positive effect in the removal efficiency; it could be
because than the frequently the recalculating occurs, it is possibly from the
operational scheme that it will be favoring a aeration inlets into the reactor and the
biomass stress increased, with the result of a lower removal efficiencies in the
system.
Good removal of both TCOD and SCOD were obtained, but for the output values of
the TCOD not reach values below 1000 mg / L and for SCOD, these are not
descended from 700 mg / L.
Total phosphorus showed good removal efficiencies greater than 66% between
different configurations.
The results make clear that ASBR with pulsated recirculation is an alternative for
the treatment of swine wastewater.
References
de Novaes L., Saratt B., Rodrigues J., Ratusznei S., de Moraes D., Ribeiro R. Zaiat M., Foresti
E.(2010). Effect of impeller type and agitation on the performance of pilot scale ASBR and
AnSBBR applied to sanitary wastewater treatment. J Environ Manage Vol.91 pp. 1647-56.
Franco A., Roca E., Lema J. (2002). Improvement of the properties of granular sludge in UASB
reactors, by flow pulsation. (Q ³9,, WDOOHU \ VLPSRVLR ODWLQRDPHULFDQR VREUH GLJHVWLyQ
anaeURELD´)(0,6&$,QWHUQDWLRQDOZDWHUDVVRFLDWLRQ<XFDWiQ0p[LFR
Graniel E., Sánchez I., Giácoman G., Ponce M. and Sauri M. (2009). Evaluación del cambio del
agua subterránea por metales pesados provenientes del exbasurero de Mérida, Yucatán.
Actas INAGEQ. Vol. 15(1) pp. 24-29.
Gürtekin, E. (2010). The effects of temperature on biological phosphorus removal in
anaerobic/anoxic sequencing batch reactor. e-Journal of New World Sciences Academy
(NWSA). Vol. 5 pp. 48-54.
Liu, Ya-nan; Xue, Gang; Yu, Shui-li; Zhao, Fang-bo. (2006). Role of extracellular exopolymers on
biological phosphorus removal. Journal of Environmental. Vol. 18 pp. 670-674.
Martin HG, Ivanova N, Kunin V, Warnecke F, Barry KW, McHardy AC, Yeates C, He S, Salamov
AA, Szeto E, Dalin E, Putnam NH, Shapiro HJ, Pangilinan JL, Rigoutsos I, Kyrpides NC,
Blackall LL, McMahon KD, Hugenholtz P (2006) Metagenomic analysis of two enhanced
biological phosphorus removal (EBPR) sludge communities. Nature biotechnology Vol. 24
pp. 1263-1269.
Mendez N., Castillo E., Vazquez E., Magaña A (2009). Manejo sanitario de granjas porcinas y
avícolas en Yucatán. Ediciones de la Universidad Autónoma de Yucatán. Mérida, Yucatán,
México.
878
Ndegwa P., Hamilton D., Lalman J. Cumba H., (2008). Effects of cycle-frequency and temperature
on the performance of anaerobic sequencing batch reactors (ASBRs) treating swine waste
Bioresource Technology, Vol. 99 pp. 1972-1980.
Ndon U., Dague R.(1997) Effects of temperature and hydraulic retention time on anaerobic
sequencing batch reactor treatment of low-strength wastewater.Water Research, Vol.
31 pp. 2455-2466.
Programa Estatal de Desarrollo Agropecuario y Pesquero del Estado de Yucatán 2007 (PEDAPEY,
2007).
Sung, S., Dague, R. (1992). Fundamental Principles of the Anaerobic Sequencing Batch Reactor
Process. Industrial Waste Conference. Proccedings . 47th Purdue University. West Lafayette,
Indiana.
Vazquez E., Zapata A. (1993). Calidad bacteriologica del agua de consume del puesto de sisal,
Yucatán. Boletín académico, (no 22). Universidad Autónoma de Yucatán. México. Facultad
de Ingeniería, pp. 13-21
Wu G., Rodgers M. (2010). Nutrient removal, microbial community and sludge settlement in
anaerobic/aerobic sequencing batch reactors without enhanced biological phosphorus
removal. Water Science And Technology. Vol. 61 (10), pp. 2433-41.
Zaiat M., Pinho SC, Ratusznei SM, Rodrigues JA, Foresti E,. (2004). Influence of the agitation rate
on the treatment of partially soluble wastewater in anaerobic sequencing batch biofilm
reactor. Water Research. Vol. 38. pp. 4117±412.
879
CHAPTER 7.21. CONTRIBUTION TO THE ASSESSMENT OF THE HEALTH
STATUS OF THE COAST OF SIDI IFNI: BIOMONITORING OF THE IMPACT OF
WASTEWATER
Mohamad ABBASSI, Abderrazak KAAYA*, Ali BANAOUI, Asma EL KHOU,
Maryem NADIR, Latifa LEFRERE
Aquatic Systems Laboratory (AQUAMAR), Faculty of Sciences ± Ibn Zohr University, BP.8106,
Agadir, Morocco
ABSTRACT
The present study, constitute a contribution to the assessment of the marine
ecosystem health in the town of Sidi Ifni (south of Morocco), in which not yet
studied was accomplished and some issues still receive wastewater.
Analysis set in this study was achieved the objectives set namely: environmental
characterization of the site by physicochemical approach and evaluation of the
degree of chemic pollution by biological approach.
For the chemical approach, analysis of physicochemical parameters of marine
water (Temperature (T°C), pH, conductivity and salinity) was do. Our results, at
both sites studied; show a real ecosystem disturbance of Sidi Ifni. This led us, in
the present conditions, to strengthen coastal surveillance of the city and
neighboring sites and to develop effective prevention against pollution. In addition,
think long term, to treat all wastewater from the city.
Preliminary measurements of biochemical indicators: Acetylcholinesterase,
Glutation S-transeferase, Catalase and Malondialehyde (AChE, GST, CAT and
MDA) in the Mediterranean mussel (Mytilus galloprovincialis) collected from the
two sites on the coast of Sidi Ifni, showed that these activities are measurable and
inducible by pollution existing in the study sites. It would now be useful to
characterize the best expressed activities and study their induction in situ, their
qualitative and quantitative correlation with contaminants in this environment.
Seasonal variations of the measured parameters show that it is imperative to take
into account such variations in the development and validation of biomarkers.
Key words: acetylcholinesterase, biomarker, catalase, glutation S-transeferase,
malondialehyde, marine ecosystem, Morocco, Mytilus galloprovincialis, sidi Ifni
880
4th International Symposium on Environmental Biotechnology and
Engineering
September 9-12, 2014
Cinvestav, Mexico City, Mexico
Introduction
With a coastline of 3,500 Km, extending along the Mediterranean and the Atlantic,
Morocco is known for its fisheries resources and touristic beaches. This ecosystem
plays an important socioeconomical role which is often accompanied by a demographic
and urban significant growth. Although, several types of pollutants are discharged into
the seawater without any treatment in a number of areas along these coasts.
As part of the efforts of our country for the implementation of a strategy of
continuous monitoring and vigilance on the quality and health of the marine environment
and its resources, our laboratory contributes by developing biochemical parameters
called biomarkers and considered as indicators of pollution and water quality (Narbonne
et al 1993; Linvingstone 1993). Such diagnostic and prognostic early-warming tests and
complementary to chemical analysis, offer the potential of specificity, sensitivity and
application to a wide range of organisms and for discrimining water contamination over
broad geographic regions. Many biomarkers are actually subject of several international
research programs of pollution monitoring (Gabrieldes 1997, Narbonne et al., 1999)
In previous studies, in wich assays of acetylcholinesterase (AChE), glutathione Stranferase (GST), catalase (CAT) and malondialdehyde (MDA) were developed and
used as biomarkers of pollution in Agadir bay, we have demonstrated that marine
organisms (like Mytilus galloprovincialis, Perna perna, Donax trunculus, Nereis
deversicolors) living at the sites receiving wastewater were significantly affected
(Moukrim et al. 1997, Najimi et al. 1997, Kaaya et al. 1999, Moukrim et al. 2004, Ait Alla
et al. 2006). The study of the biology of these organisms showed also many
perturbations in the reproductive cycle and growth (Id Halla et al. 1997, Lagbouri 1997).
In ordrer to generalize this biological approach and evaluate the health state of
some marine ecosystems of south Atlantic Moroccan coasts not yet well studied, we
have established a research program focused on the study of the four biomarkers,
mentioned above, in the Mediterranean mussel (Mytilus galloprovincialis) collected from
the two sites on the coast of Sidi Ifni region (south of Morocco) never studied.
Our preliminary study was conducted on samples of standardized size (30 to 50 mm) of
Mytilus galloprovincialis collected in two sites representative of the coastline of Sidi Ifni
(Figure 1): i) Mirleft, located at 30 Km to the north of Sidi Ifni and far from any source of
pollution (MIR) is considered as reference site, and ii) Cheikh Sidi Ali Ifni located at the
entrance of the Sidi Ifini city and receives untreated waste waters of this city (CHK).
FIGURE 1. Map of the sites sampled in coastline of Sidi Ifni [CHK: Cheikh Sidi Ali
Ifni (Polluted site), MIR: Mirleft (Reference site)]
Mussels collected monthly (between June and November 2013 for our preliminary
study) were washed in fresh seawater and rapidly transported to laboratory and frozen
at ± 30°C until analysis. All the preparation procedures were conducted at 4°C. Soft
tissue (whole animal) were collected, thawed, washed with cold 100 mM Tris buffer (pH
7.4), weighed and homogenized in three volumes (w/v) of the same buffer with an Ultra
Turax homogenizer. Homogenates were centrifuged at 9000g for 30 minutes and the
resulting supernatant (post-mitochondrial fraction or S9) frozen (- 30°C) until use.
The AChE activity, used as biomarker of neurotoxicity inhibited by the presence of
pesticides such as organophosphorous compounds, carbamates and various heavy
metals or PAHs in mussels (Bocquené et al., 1993; Mora et al. 1999), was determined
by the method described by Ellman et al. (1961) using acetylthiocholine iodide (ATC) as
substrate. Spectrophotometric measurement was performed at 412 nm every 15
seconds for 2 minutes at 25°C. The reaction mixture containing 1.05 ml of 100 mM Tris
buffer (pH 7.4), 50 µl DTNB (80 mM mM in the assay), 50 µl of acetylthiochoilne (ACT
45 mM in the assay) and 50 µl of S9. Enzyme activity was expressed as nmoles of
882
acetylthiocholine/min/mg of S9 protein using the molar coefficient of extinction 13.6 ×
103 M±1× cm±1.
GST activity, used as biomarker of pollution exposure to various foreign
compounds like PAH, PCB Phenobarbital (Buhler and Williams 1988, Suteau et al.
1988, Fitzpatrick et al. 1997), was measured according to Habig et al. (1974) using 1chloro-2-4-dinitrobenzene (CDNB) as a substrate. This latter is often preferred choice
when total GST is being measured and is recommended for determination of GST
activities (Habig et al. 1974). Assay was carried out in a reaction mixture containing 1 ml
of 100 mM Tris buffer (pH 7.4), 50 µl CDNB (1 mM in the assay), 50 µl of reduced
glutathione (GSH 1mM in the assay) and 50 µl of S9. Enzyme activity was determined
by monitoring changes in absorbency at 340 nm for 2 minutes at 25°C. GST activities
are expressed as nmoles of dinitrophenyl glutathione (produced by conjugation of
CDNB et GSH) per minute per milligram of S9 protein using the molar coefficient of
extinction 9.6 mM±1× cm±1.
Assay of CAT activity, used as a biomarker of the oxidative stresses induced by a
wide range of contaminants, including organic xenobiotics, heavy metals and PAHs
(Livingstone 2001; Sureda et al. 2011), was accomplished according the method of Aebi
(1985) which quantify the loss of H2O2 per minute at 240 nm in a reaction mixture
containing 1 ml of potassium phosphate buffer (pH 7.4), 1 ml of H2O2 and 50 µl of S9.
The activity of CAT was expressed in nmoles of H2O2 transformed per minute per
milligram of S9 protein using the molar coefficient of extinction 40 M±1× cm±1.
The determination of the MDA, used as marker of membrane phospholipid
oxidation through lipid peroxidation (Avery et al. 1996, Charissou et al. 2004), was
estimated in terms of thiobarbituric acid reactive species (TBARS), with the use of 1, 1,
3, 3-treaethyloxypropane as standard. The reaction was assessed at 532 nm using TBA
reagent as described by Sunderman (1985). MDA content was expressed as mg of
MDA/mg S9 protein.
The protein content in samples was evaluated according the method of Lowry et
al. (1951) using BSA as standard.
Data were expressed as mean ± standard error (SD). The statistical significance of
WKH GLIIHUHQFHV EHWZHHQ VDPSOHV ZDV GHWHUPLQHG E\ WKH ³W´ WHVW XVLQJ WKH 6WDWLVWLFDl
software. A p value of less 0.05 was considered as statistically significant.
During the sampling period (June to November 2013), ACHE activity measured in
Mytilus galloprovincialis (Figure 2) showed that this biomarker of neurotoxicity was
inhibited in mussels populations living in polluted site (CHK) versus those sampled in
Mirleft considered as reference site. The amplitude of the inhibition is more significantly
marked in september and november (with a percent inhibition of 13,85% and 31,69% in
october and November respectively). Such changes would certainly be related to the
state of pollution of the CHK site that receives domestic wastewater and agricultural
discharge without any prior treatment of contaminants that would be responsible for the
inhibition of AChE. Indeed, many studies have demonstrated that ACHE activities were
inhibited in the presence of some contaminants like pesticides linked to agricultural
activities and metals (Galgani & Bocquéné 1988, Scaps & Borot 2000). AChE has a
883
fundamental role in the nervous system of both vertebrates and invertebrates, and its
inhibition is considered a typical effect of organophosphate and carbamate pesticides
[38]. Similar effects may be caused by other factors which are known to modulate this
enzymatic activity, including trace metals (Labrot et al. 1996, Amiard-Triquet et al.
1998). The observed inhibition of AChE activities may be attributed to the presence of
contaminants in the environment. The same inhibitions were observed in Mytilus
galloprovincialis and Perna perna mussels (Najimi et al. 1997) living in stations
receiving wastewater and contaminated by PAHs (Azdi et al. 2006), metals such as Fe,
Zn, Cd, and Cu (Banaoui et al. 2004) and pesticides (Agnaou et al. 2014) in Agadir Bay.
FIGURE 2. Acethylcolinesterase activity (ACHE) in Mytilus galloprovincialis
collected the coastline of Sidi Ifni [CHK: Cheikh Sidi Ali Ifni (Polluted site), MIR:
Mirleft (Reference site)].
The ACHE activity seems to follow a seasonal profile. Higher levels of ACHE
activity were observed in mussels collected in the two sites during the summer months
than such collected at the autumn. Previous studies in our laboratory have shown
seasonal variation for this enzyme activity in Mytilus galloprovincialis and Perna perna,
which could be related to different levels of cholinergic system activation during the
reproductive cycle of mussels (Najimi et al. 1997).
Concerning GST, the present study shows a markedly higher enzyme activity in
mussels living in CHK site versus Mirleft site (Figure 3). The amplitude of the induction
is more significantly marked in summer months (with a percent induction of 208,02%,
138,07% and 221,52% in june, july and august respectively). Results like ours have
already obtained the Mytilus galloprovincialis and Perna perna in Moroccan coasts
(Agadir bay) (Kaaya et al. 1999). Several other studies have described a similar
relationship between environmental pollution and GST activity in mussels and other
organisms in many marine ecosystems (Burgeot et al. 1996, Bainy et al. 2000; Cheung
et al., 2001, 2002; Gowland et al. 2002, Lau & Wong 2003, Manduzio et al. 2004).
Indeed, it is known that GST activity is a phase II enzyme involved in the detoxification
metabolism of lipophilic organic contaminants. This enzyme catalyzes the conjugation of
various organic electrophilic compounds making it easily extractable. Its induction will
884
seem to be, in our opinion, an adaptive response to altered environment by wastewater
in CHK site.
Figure 3: Glutathione-Stransferase activity (GST) in Mytilus galloprovincialis
collected the coastline of Sidi Ifni [CHK: Cheikh Sidi Ali Ifni (Polluted site), MIR:
Mirleft (Reference site)].
Otherwise and like ACHE, GST activity seems present a seasonal evolution and
variation in presence of pollution. Higher enzyme activity was obtained in summer.
Seasonal variations of this activity were already described in Mytilus galloprovincialis
and Perna perna in Moroccan coasts (Agadir bay) (Kaaya et al. 1999) and others
organisms in different regions (Wilhelm Filho et al. 2001, Sladjan et al. 2010).
Figure 4 show that changes in catalase activity during the six months of study
recorded high values in CHK site versus Mirleft. The induction of this enzyme activity in
mussels living in polluted site was significantly marked in the beginning of summer (with
a percent induction of 50%, and 48,632% in june and july respectively) and the end of
autumn (with a percent induction of 57,73% in november). Results obtained in our study
are in general agreement with induction linked to pollution described in Mytilus.
galloprovincialis mussels in Bizerte lagoon (Tunisia) by Dellali et al. (2001) and in the
coast of Casablanca (El Jourmi et al. 2014). Such variations are due to various
environmental stimuli that may induce prooxidative processes in organisms which are
often associated with biochemical and histological alterations in mollusks (Nicholson,
2001; Chelomin et al., 2005; Manduzio et al., 2005). Indeed, CAT activity is considered
as primary enzyme in the antioxidant defense system of organisms is the earliest
antioxidant enzymes induced (Capello et al. 2013) against higher H2O2 generation
directly or indirectly induced by contaminants present the CHK site which received
domestic wastewater and agricultural discharge without any treatment.
The seasonal trend of CAT activity obtained in our study confirmed the results
described in Mytilus galloprovincialis living in Basque Estuaries (Bay of Biscay) (Orbea
885
et al 1999) and in the Saronikos Gulf of Greece (Vlahogianni et al. 2007) and in
Ruditapes decussatus in Tunisian coastal areas (Banni et al. 2009, Kamel et a. 2014).
FIGURE 4. Catalase activity (CAT) in Mytilus galloprovincialis collected the
coastline of Sidi Ifni [CHK: Cheikh Sidi Ali Ifni (Polluted site), MIR: Mirleft
(Reference site)]. Data were expressed as mean ± standard error (SD).
The monthly monitoring of MDA levels shows high values in Mytilus
galloprovincialis collected in the polluted site (CHK) during the period of our study
(Figure 5). The pourcentage of induction is comparable during the period of study and
including between 13 and 30%. Such variation linked to pollution has also been
reported by El Jourmi et al. (2012) in Perna perna along the moroccan atlantic coast
(big Casablanca) and by Kamel at al. (2014) in Mytilus galloprovincialis sampled in
Bizerte lagoon (Tunisia). Indeed, measurement of MDA is widely used as an indicator of
lipid peroxidation (Wheatley 2000). Several other studies have described that MDA
levels may be positively correlated with the level of certain pollutants. Indeed, the
capacity for heavy metals and organic compounds to induce lipo-peroxidation was
reported by Livingstone et al (1990), Viarengo et al, (1990), Narbonne et al. (1991) and
Labrot et al (1996).
886
FIGURE 5. Malondialdehyde level (MDA) in Mytilus galloprovincialis collected the
coastline of Sidi Ifni [CHK: Cheikh Sidi Ali Ifni (Polluted site), MIR: Mirleft (Reference
site)].
An apparent and seasonal evolution of MDA level was obtained in our study. This
kind of results was described in Perna viridis living in sea beach of Bambolim (India)
(Verlecar et al. 2007) and translate the relationship between reproductive cycle and
oxidative stress described by Filho et al. (2001).
Conclusion
The aim of this work is to initiate a program of research about the study of the health
state of the coastline of Sidi Ifni (South of Morocco) not yet studied and to test as
biomarkers of pollution a battery of biochemical parameters already validated in large
scale in marine ecosystems.
The result obtained in our preliminary work, despite its limitations in time and space,
indicates a significant influence of pollution on all parameters tested (ACHE, GST, CAT
and MDA) in Mytilus galloprovincialis. The biomarkers responses obtained during our
period of study (June to November 2013) shown clearly the presence of different
contaminants in CHK site which receives untreated waste waters and validate the large
application of biomarkers studied in Moroccan marine ecosystems in a sentinel
organisms, Mytilus galloprovincialis. Our result indicates also a significant influence of season on all biomarkers tested. This
variation must be correlated to the complex interactions between biomarkers and
exogenous and endogenous factors. These results oblige us to take these variations
into account in the development and validation of biomarkers as biological approach
which well be our useful tool for monitoring in an integrated program of environmental
measurements in Moroccan marine ecosystems. Acknowledgements
887
This work was funded by the Ministry of Higher Education, Scientific Research and
Professional Training (Morocco) We would like to thank Pr Ihya Ait Ichou (Dean of the Faculty of Sciences of Agadir) for his technical help. References
Agnaou M., A. Ait Alla, M. Ouassas, Lh. Bazzi, Z. El Alami, A. Moukrim (2014). Assessment of
organochlorine pesticides contamination of Oued Souss estuary (South of Morocco): Seasonal
variability in sediment and a detritivore annelid Neries diversicolor. J. Mater. Environ. Sci. 5 (2) :
581-586.
Ait Alla A, Mouneyrac C, Durou C, Moukrim A, Pellerin J (2006). Tolerance and biomarkers as useful
tools for assessing environmental quality in the Oued Souss estuary (Bay of Agadir, Morocco).
Comp Biochem Physiol C. Comp Biochem Phys C 143(1):23-29.
Amiard-Triquet C., Altman S., Amiard,J.C., Ballandufrançais,C., Baumard P., Budzinski,H., Crouzet, C.,
Garrigues,P., His E., Jeantet A. Y., Menasria R., Mora P., Mouneyrac C., Narbonne J. F., Pavillon
J. F., (1998). Fate and effects of micropollutantes in the Gironde estuary, France: a
multidisciplinary approach. Hydrobiologia, 373/374: 259-279.
Avery E L, Dunstan RH, Nell JA, (1996). The detection of pollutant impact in marine environments:
condition index, oxidative DNA damage, and their associations with metal bioaccumulation in the
Sydney rock oyster Saccostrea commercialis. Arch Environ Contam Toxicol 31: 192±198.
Azdi M., Moukrim A., Burgeot Thierry, Budzinski H., Chiffoleau Jean-Francois, Kaaya A., Zekhnini A.,
Narbonne J. F., Guarrigues Ph. (2006). Hydrocarbon pollution along Moroccan coasts and BPH
activity in the mussel Perna perna. Polycyclic Aromatic Compounds (1040-6638) (Taylor & Francis
Ltd), Vol. 26 , N. 4: 265-282.
Bainy A.C.D., Almeida E.A., Muller I.C., Ventura E.C., Medeiros I.D. (2000). Biochemical responses in
farmed mussel Perna perna transplanted to contaminated sites on Santa Catarina Island, SC
Brazil. Mar. Environ. Res. 50: 411-416.
Banaoui A, Chiffoleau JF, Moukrim A, Azdi M, Kaaya A, Auger D, Rozuel E (2004). Trace metal
distribution in mussel Perna perna along the Moroccan coast. Mar Pollut Bull 48:378-402.
Banni M., Bouraoui Z., Ghedira J., Clearandeau C., Jebali J., Boussetta H. (2009). Seasonal variation of
oxidative stress biomarkers in clams Ruditapes decussatus sampled from Tunisian coastal. Environ
Monit Assess.,155 (1-4):119-28.
Bocquené G, Galgani F, Burgeot T, Le Déan L , Truquet P (1993). Acetylcholinesterase levels in marine
organism along french coasts. Mar Poll Bull 26: 101-106.
Buhler D.R. & Williams D.E. (1988). The role of biotransformation in the toxicity of chemicals. Aquat.
Toxicol., 11: .19-28.
Burgeot T., Bocquene G., Porte C., Dimeet J., Santella R.M., GarciadelaParra L.M., Pihol-Leszkowicz A.,
Raoux C., Galgani F. (1996). Bioindicators of pollutant exposure in the north western
Mediterranean Sea. Mar. Ecol. Prog. Ser. 131:125-141.
Cappello T., Maisano M., D'Agata A., Natalotto A., Mauceri A., Fasulo S. (2013). Effects of environmental
pollution in caged mussels (Mytilus galloprovincialis). Mar Environ Res. 91:52-60.
Charissou AM, Cossu-Leguille C, Vasseur P (2004). Relationship between two oxidative stress
biomarkers, malondialdehyde and 8-oxo-7,8-dihydro- 2'-deoxyguanosine, in the freshwater bivalve
Unio tumidus. Sci Tot Environ 322: 109±122.
Chelomin V.P., Zakhartsev M.V., Kurilenko A.V., Belcheva N.N. (2005). An in vitro study of the effect of
reactive oxygen species on subcellular distribution of deposited cadmium in digestive gland of
mussel Crenomytilus grayanus. Aquat. Toxicol. 73: 181-189.
Cheung C.C.C., Zheng G.J., Li A.M.Y., Richardson B.J., Lam P.K.S. (2001). Relation- ships between
tissue concentrations of polycyclic aromatic hydrocarbons and antioxidative responses of marine
mussels, Perna viridis. Aquat. Toxicol. 52: 189±203.
888
Cheung C.C.C., Zheng G.J., Lam P.K.S., Richardson B.J. (2002). Relationships between tissue
concentrations of chlorinated hydrocarbons (polychlorinated biphenyls and chlorinated pesticides)
and antioxidative responses of marine mussels, Perna viridis. Mar. Pollut. Bull. 45: 181±191.
El Jourmi L., Amine A., Boutaleb N., Labbardi H., Lazar S., El Antri S. (2012). Oxidative stress biomarkers
in mussels sampled from four sites along the Moroccan atlantic coast (big casablanca). Scientific
Study & Research - Chemistry & Chemical Engineering, Biotechnology, Food Industry, 13 (3): 325
± 331
El Jourmi L., Amine A., Boutaleb N., Abouakil N., Lazar S., El Antri S. (2014). Multimarker approach
analysis in the brown mussel to evaluate the anthropogenic stress: A preliminary study. J. Mater.
Environ. Sci. 5 (5): 1326-1331.
Dellali M., Gnassia-Barelli M., Romeo M., Aissa P. (2001) The use of acetylcholinesterase activity in
Ruditapes decussatus and Mytilus galloprovincialis in the biomonitoring of Bizerta lagoon. Comp.
Biochem. Physiol. C 130: 227-235.
Filho D. W., Tribess T., Gaspari C., Claudio F. D., Torres M. A., Magalhaes A. R. M. (2001). Seasonal
changes in antioxidant defenses of the digestive gland of the brown mussel Perna perna.
Aquaculture, 203: 149-158.
)LW]SDWULFN 3- 2¶+DOORUDQ - 6KHHKan D. & Walsh A.R. (1997). Assessment of a glutathione Stransferase and related proteins in the gill and digestive gland of Mytilus edulis (L.), as potential
organic pollution biomarkers. Biomarkers, 2: 51-56.
Gabrieldes G. (Ed.) (1997). The MED POL Biomonitoring Program Concerning the Effects of Pollutants
on Marine Organisms Along the Mediterranean Coasts. UNEP (OCA)/MED WG. 132/3. Athens.
Galgani F. & Bocquéné G. (1988). A method for routine detection of organophosphates and carbamates
in sea water. Environ. Technol. Lett., 10: 311-322.
Gowland B.T.G., Mcintosh A.D., Davies I.M., Moffat C.F., Webster L. (2002). Implications from a field
study regarding the relationship between polycyclic aromatic hydrocarbons and glutathione-Stransferase activity in mussels. Mar. Environ. Res. 54 : 231-235.
Habig W.H., Pabst M.J. & Jakoby, W.B. (1974). Glutathione S-transferases: The first enzymatic step in
mercapturic acid formation. J. Biol. Chem., 249: 7130-7139.
Id Halla M., Bouhaimi A., Zekhnini A., Narbonne JF., Mathieu M. & Moukrim A. (1997). Etude du cycle de
reproduction de deux moules Perna perna (Linné 1785) et Mytilus galloprovincialis (Lamarck 1819)
dans la baie d'Agadir, (Sud du Maroc). Haliotis, 26: 51-62.
Kaaya A., Najimi S., Ribera D., Narbonne J. F., Moukrim A. (1999). Characterization of glutathione Stransferases (GST) activities in Perna perna and Mytilus galloprovincialis used as a biomarker of
pollution in the Agadir Marine Bay (South of Morocco). Bull. Environ. Contam. Toxicol.62, 623±629.
Kamal N., Burgeot Th, Banni M., Chalghaf M., Devin S. Minier Ch., Boussetta H. (2014). Effects of
increasing temperatures on biomarker responses and accumulation of hazardous substances in
rope mussels (Mytilus galloprovincialis) from Bizerte lagoon. Environ Sc Poll Res, 21 (9) : 6108.
Labrot F. Ribera D, Saint Denis M., Narbonne J. F. (1996). In vitro and in vivo studies of potential
biomarkers of lead and uranium contamination: lipid peroxidation, acetylcholinesterase, catalase
and glutathione peroxidase activities in three non-mammalian species. Biomarkers, Vol. 1, No. 1:
21-28.
Lagbouri A. (1997). Etude de la biologie de Donax trunculus dans la baie d'Agadir et réponse de trois
biomarqueurs de pollution (Acétylcholinestérase, peroxydation lipidique et glutathion S transférase). 'LSO{PH G¶(WXGHV 6XSpULHXUHV )DFXOWp GHV 6FLHQFHV 8QLYHUVLWp ,EQ =RKU ,
Agadir.
Lau P.S. & Wong H.L. (2003). Effect of size, tissue parts and location on six biochemical markers in the
green-lipped mussel, Perna viridis. Mar. Pollut. Bull. 46 : 1563-1572.
Livingstone D.R (1990) Cytochrome P450 and oxidative metabolism in inverterbrates. Biochem. Soc.
Trans., 18(1): 15-19.
Livingstone DR (2001). Contaminant-stimulated reactive oxygen species production and oxidative
damage in aquatic organisms. Mar Pollut Bull 42:656±66
Lowry O.H., Roseborough N.J., Farrand A.L. & Randall R.J. (1951). Protein measurement with the folin
phenol reagent. J. Biol. Chem. 193: 265-275.
Manduzio H., Monsinjon T., Galap C. Leboulenger F., Rocher B. (2004). Seasonal variations in
antioxidant defenses in blue mussels Mytilus edulis collected from a polluted area: major
889
contributions in gills of an inducible isoform of Cu/Zn-superoxide dismutase and of glutathione-S
transferase. Aquat. Toxicol. 70: 83±93.
Manduzio H., Rocher B., Durand F., Galap C., Leboulenger F. (2005). The point about oxidative stress in
molluscs. Invertebrate Survival J. 2: 91-104.
Mora P, Fournier D, Narbonne JF (1999). Cholinesterases from the marine mussels Mytilus
galloprovincialis LmK. and Mytillus edulis from the freshwater bivalve Corbicula fluminea Muller.
Comp Bioche Physiol C122: 353± 36
Moukrim A, Bouhaimi A., Lagbouri A., Kaaya A., Zekhnini A (1997). Health state of Agadir bay evaluated
by biomarkers of pollution. Proceedings of International Conference on Water Problems in the
Mediterranean Countries, 17-21 November 1997, Nicosia-North, Cyprus.
Moukrim A, F. El Hamidi, A. Lagbouri, A. Kaaya, A. Zekhnini, A. Bouhaimi, and J. F. Narbonne (2004).
Study of Donax trunculus as a Sentinel Species for Environmental Monitoring of Sandy
Beaches on Moroccan Coasts; Bull. Environ. Contam. Toxicol.73, 674-681
Najimi S., Bouhaimi A., Daubeze M., Zekhnini A., Pellerin J., Narbonne J. F. & Moukrim, A. (1997). Use of
acetylcholinesterase in Perna perna and Mytilus galloprovincialis as a biomarker of pollution in
Agadir Marine Bay (South of Morocco). Bull. Environ. Contam. Toxicol., 58: 901-908.
Narbonne J.F., Garrigues P., Ribera D., Raoux C., Mathieu A., Lemaire P., Salaün J.P. & Lafaurie M.
(1991). Mixed-function oxygenase enzymes as tools for pollution monitoring : field studies on the
frensh coast of the mediterranean sea. Comp. Biochem. Physiol., 100C(1/2): 37-42.
Narbonne, J.F., Daubeze, M., Clerandeau, C., Garrigues, P. (1999). Scale of classification based on
biochemical markers in mussels: application to pollution monitoringin European coasts.
Biomarkers, 4: 415-424.
Nicholson S. (2001). Ecocytological and toxicological responses to copper in Perna viridis (L.) (Bivalvia:
Mytilidae) haemocyte lysosomalmembranes. Chemosphere 45 : 399-407.
Orbea A., Marigo´mez I., Fernandez C., Tarazona J. V., Cancio I., M. P. Cajaraville M. P. (1999).
Structure of peroxisomes and activity of the marker enzyme catalase in digestive epithelial cells in
relation to PAH content of mussels from two Basque estuaries (Bay of Biscay): Seasonal and sitespecific variations. Arch. Environ. Contam. Toxicol., 3:, 158-166 (1999)
Scaps P.& Borot, O. (2000). Acetylcholinesterase activity of the polychaete Nereis diversicolor: effects of
temperature and salinity. Comp. Biochem. Phys. C 125: 377-383.
Sladjan Z., Pavlovic Slavica S., Borkovic Mitic, Tijana B., Radovanovic Branka R., Perendija Svetlana G.,
Despotovic Jelena P., Gavric, Zorica S., Saicic (2010). Seasonal Variations of the Activity of
Antioxidant Defense Enzymes in the Red Mullet (Mullus barbatus l.) from the Adriatic Sea. Mar
Drugs; 8(3): 413±428.
Sureda A, Box A, Tejada S, Blanco A, Caixach J, Deudero S (2011). Biochemical responses of Mytilus
galloprovincialis as biomarkers of acute environmental pollution caused by the Don Pedro oil spill
(Eivissa Island, Spain). Aquat Toxicol 101: 540-549.
Suteau PM, Daubeze M, Migaud ML, Narbonne JF (1988). PAH-metabolizing enzymes in whole mussels
as biochemical tests for chemical pollution monitoring. Mar. Ecol. Prog. Ser, 46: 45-49.
Verlecar X. N., Jena K. B., Chainy G. B. (2007). Biochemical markers of oxidative stress in Perna viridis
exposed to mercury and temperature. Chem Biol Interact., May 1;167 (3): 219-26.
Viarengo A., Canesi L., Pertica M., Poli G., Moore M, N. & Orunesu M. (1990): Heavymetal effects on lipid
peroxidation in the tissues of Mytilus galloprovincialis Lam. Comp. Biochem. Phys. C, 97 (1): 37-42.
Vlahogianni T., Dassenakis M., Scoullos M. J., Valavanidis A. (2007). Integrated use of biomarkers
(superoxide dismutase, catalase and lipid peroxidation) in mussels Mytilus galloprovincialis for
DVVHVVLQJ KHDY\ PHWDOV¶ SROOXWLRQ LQ FRDVWDO DUHDV IURP WKH 6DURQLNRV *XOI RI *UHHFH Mar Poll
Bull, 54: 1361-1371.
Wheatley R. A. (2000). Some recent trends in the analytical chemistry of lipid peroxidation. Trends in
Analytical Chemistry, 19 (10): 617-628
Wilhelm Filho D, Torres M. A., Tribess T. B., Pedrosa R. C., Soares C. H. L. (2001). Influence of season
and pollution on the antioxidant defenses of the cichlid fish acará (Geophagus brasiliensis).
Antioxidant defenses in Geophagus. Brazilian. Journal of Medical and Biological Research, 34:
719-726.
890
Notation
ACHE: Acetylcholinestersae
CAT: Catalase
CHK: Cheikh Sidi Ali Ifni (Polluted site)
GST; Glutathion S-transferase
MDA: Malondialdehyde
MIR: Mirleft (Reference site)
891
CHAPTER 7.22. REMOCION DE METALES PESADOS UTILIZANDO SISTEMAS
INMOVILIZADOS DE Spirulina máxima
A.R. Domínguez-Bocanegra
División de química
[email protected]
y
Bioquímica
Tecnológico
de
Estudios
Superiores
de
Ecatepec
RESUMEN
En el presente trabajo se estudió la capacidad que presenta la biomasa de Spirulina
maxima (Arthrospira) inmovilizada en alginato de sodio y agar-agar para crecer,
adsorber y remover cadmio y plomo presentes en aguas residuales. Los experimentos
se llevaron a cabo en matraces Erlenmeyer de 1000 mL de capacidad total, con 800
mL de medio de cultivo Zarrouk modificado y 100 g de pellets con y sin células de S.
maxima en fase exponencial de crecimiento, adicionando simultáneamente y de forma
paralela los metales cadmio y plomo a 5ppm, 10ppm, 20ppm y 50 ppm a partir de
soluciones patrón de acetato de cadmio (CH3COO)2 Cd·2H2O y carbonato de plomo
PbCO3 respectivamente. Los matraces se incubaron a temperatura ambiente (28
±2ºC), aireación continua 0.5 vvm, agitación 100 rpm y fotoperiodo natural (luz solar).
Para evaluar la capacidad de adsorción de la biomasa de S. maxima y de los soportes
(alginato de sodio y agar-agar) se tomaron muestras durante la primera hora cada 10
min, en la segunda hora cada 20 minutos y durante la tercera hora cada 30 minutos; a
partir de la cuarta hora se tomaron muestras cada 60 minutos hasta completar 8 h,
posteriormente cada 24 h. 6H FXDQWLILFR SHVR VHFR FRQFHQWUDFLyQ GH FORURILOD ³D´ \
concentración de metal utilizando un espectrofotómetro de Absorción atómica marca
Varian. Los resultados obtenidos indican que la máxima adsorción de los metales en la
biomasa inmovilizada de Spirulina maxima se llevó a cabo durante la primera
obteniendo una remoción del 84% en presencia de 20 ppmPb +2 en alginato de sodio y
un 79% con agar-agar, mientras que en presencia de 20ppmCd2 los porcentajes de
remoción fueron menores obtenido un 24% con alginato de sodio y 19.7% en agaragar.
La capacidad máxima de biosorción (qmáx) de S. maxima, bajo estas
condiciones, para cadmio fue de qmax 0.85 g de Cd/g biomasa con alginato de sodio y
de qmax 0.56 mg de Cd/g biomasa; en agar-agar mientras que para plomo fue de qmax
11.61mgPb2+/g biomasa en alginato de sodio y de qmax de 1.5 mgPb2+/g biomasa en
agar-agar.
Introducción
A lo largo de las últimas décadas nuestro planeta se ha encontrado expuesto a
innumerables fuentes de contaminación lo cual ha llevado a la comunidad científica a
desarrollar investigaciones que permitan controlar y eliminar los agentes que causan
dicha contaminación, uno de los temas más preocupantes es el de la contaminación del
agua ya que sin esta la vida en nuestro planeta no sería posible. Los compuestos más
tóxicos de las aguas contaminadas son los metales pesados, por sus características
xenobíticas y recalcitrantes; para tratar esta grave situación existen técnicas de
892
purificación para este vital liquido; desafortunadamente las tecnologías que se
requieren para estos procedimientos son altamente costosas, la bioquímica ofrece una
amplia gama de alternativas para llevar a cabo una biorremediación del agua, es decir
la utilización de microorganismos capaces de remover una importante cantidad de
contaminantes. La bioadsorción es una técnica que utiliza biomasa viva, muerta o
inmovilizada para la remoción y recuperación de metales pesados a partir de
disoluciones acuosas, y se puede considerar un método alterno a los métodos
convencionales para el tratamiento de efluentes contaminados. La naturaleza nos
brinda una inmensa variedad de microorganismos que se adaptan con facilidad a
diversas condiciones ambientales tal es el caso de las cianobacterias, de entre las
cuales s. máxima, se distingue por la composición de su pared celular, su actividad
fotosintética, óptimo desarrollo en sistemas inmovilizados, capacidad para captar
tolerar y utilizar ciertos metales pesados lo cual la convierte en una excelente opción
para el tratamiento de aguas residuales contaminadas por metales pesados. El
objetivo del presente estudio fue evaluar la capacidad de adsorción y crecimiento de
Spirulina maxima (Arthrospira) en sistemas inmovilizados para la remoción metales
pesados presentes en medio líquido.
Spirulina maxima (Arthrospira) aislada del río de los Remedios Ecatepec Estado de
México.
Inmovilización de la biomasa. La biomasa S. maxima se mezcló en relación 0.5:1 en
masa, con una solución acuosa de alginato de sodio a 3%. Posteriormente, por medio
de un gotero la mezcla se hizo gotear en un matraz que contenía una solución de CaCl
0.1 M. de esta manera se obtuvo la biomasa inmovilizada, en alginato de calcio, en
forme de pellets. Conjuntamente se prepararon pellets de agar-agar a 3% para ello se
utilizó una solución acuosa de agar- agar y se mezcló con la biomasa en una relación
0.5:1 en masa posteriormente la mezcla se vació en placas de petri una vez gelificada
la mezcla se formaron los pellets con ayuda de un horadador de un centímetro de
diámetro.
Biosorción de metales. En matraces Elernmeyer de 1000 mL, de capacidad total, se
colocan 800 ml de medio Zarrouk y 100g de pellets sin células para evaluar la
capacidad de adsorción del metal por el soporte utilizado (agar-agar) para la
inmovilización de S. máxima; se prepararon 100g de pellets con biomasa de S. máxima
en fase de crecimiento exponencial, simultáneamente se agrega respectivamente
Cadmio y Plomo a concentraciones de 5,10 y 20 ppm a partir de soluciones de acetato
de Cadmio (CH3COO)2 Cd 2H2O y Bicarbonato de plomo PbCO3 respectivamente. Los
matraces se incubaron a temperatura ambiente (28 ±2ºC), aireación 0.5 vvm, agitación
100 rpm, foto periodo 12hluz (1192 luxes) /12h oscuridad. Después de agregado el
metal se tomó muestra del sobrenadante, al inicio de la cinética t o, y durante la primera
hora cada 10min, en la segunda hora cada 20 minutos, en la tercera hora cada 30
minutos y a partir de la cuarta hora cada 60 minutos hasta las 8 h y posteriormente se
893
tomaron cada 24 h. Todos los experimentos de absorción se se realizaron por triplicado
y de forme independiente.
Finalmente la cuantificación de metales pesados se realiza por medio de
Absorción atómica en un espectrofotómetro de la marca Varian. Se ajustaron los datos
experimentales a los modelos matemáticos de Freundlich y Langmuir (2) y (3).
q
KC e
1
n
(2)
Donde q es la capacidad de absorción (mg/g); Ce es la concentración de
equilibrio (mg/L); es el parámetro de la ecuación (L/mg) relacionado con la afinidad del
bioadsorbente por los iones metálicos y n es el parámetro de la ecuación relacionado
con la intensidad de la adsorción.
( q max bC e
(3)
qt
1 bC e
Donde q es la capacidad de adsorción (mg/g); Ce es la concentración de equilibrio
(mg/L); qmax es la capacidad máxima de adsorción (mg/g) y b el parámetro de la
ecuación (L/mg).
Resultados y discusión.
La figura 1 muestra el crecimiento de Spirulina maxima dentro de los pellets de
alginato de sodio (1a) y en los pellets de agar-agar (1b) en presencia de metales
pesados.
1a
FIGURAS 1a) Pellets de alginato de sodio con
agar.agar con S. maxima
1b
S. maxima
1b) Pellets de
Biosorción de cadmio por S. maxima en alginato de sodio. En seguida se muestra
la concentración de cadmio captada por la biomasa S. maxima inmovilizada en alginato
de sodio durante la cinética de adsorción. Este proceso se dividió en dos fases; una
rápida y otra más lenta, la primera se llevó a cabo durante las primeras 8 horas donde
se obtuvo la mayor adsorción del metal (fig. 2). Se observa que a 20 ppm el cadmio
captado fue de 4.8 mgL-1, mientras que para 10 y 5 ppm fue de 6.42 mgL -1 y 5 mgL-1
respectivamente. En la segunda etapa se ve claramente como S. maxima capta el
894
cadmio más lentamente y casi permanece constante. La ventaja de utilizar a este
microorganismo en sistemas inmovilizados es que con esta técnica se optimiza el
proceso de eliminación y recuperación de metales pesados (Ilangovan et al. 1998).
FIGURA 2. Cd2+ captado durante las primeras 8 horas por
S. maxima
inmovilizada en alginato de sodio a temperatura ambiente 28 ±2ºC, aireación 0.5
vvm, agitación 100 rpm, fotoperiodo natural.
Biosorción de Cadmio por S. maxima en agar-agar. En la figura 3 se muestra la
biosorción de cadmio por S. maxima inmovilizada en pellets de agar-agar; se observa
que la concentración captada a 20 ppmCa+2 fue del 20.4%, para 10 ppmCa+2 de
59.6% y finalmente para 5 ppm Ca+2 se logró captar el 100% de la concentración
inicial por pellets de agar-agar con biomasa de S. maxima, esto nuevamente se llevó a
cabo en dos etapas; durante las primeras 8 horas se capta la mayor parte de metal y
en el tiempo restante, el proceso se alcanzó el equilibrio y se vuelve estable, por lo
tanto se observa que la biomasa de S. maxima es capaz de captar metales pesados a
bajas concertaciones. (Quintelas, C. 2009).
895
FIGURA 3. Cd2+ captado en las primeras 8 horas de remoción por S. maxima
inmovilizada en agar-agar a temperatura ambiente 28 ±2ºC, aireación 0.5 vvm,
agitación 100 rpm, fotoperiodo natural.
Biosorción de plomo por S. maxima en alginato de sodio. En las figura 4, se
muestra la variación de la concentración de plomo captada por biomasa de S. maxima
inmovilizada en pellets de alginato de sodio con diferentes concentraciones de plomo:
en presencia de 20 ppm Pb+2 la concentración captada al final del proceso fue de 16.8
mgL-1, mientras que para 10 ppm Pb+2 la concentración final captada fue de 7.9mgL-1;
por ultimo para el caso de 5ppm Pb+2 la concentración captada durante todo el
proceso fue de 5 mgL-1. Este comportamiento confirma que la biomasa de S. maxima
tiene la capacidad de captar metales pesados a bajas concentraciones sin afectar su
crecimiento normal, incluso es capaz de utilizar al metal como un mecanismo alterno
para su mejor desarrollo (Hong et al., 2005).
896
FIGURA 4. Pb2+ captado en las primeras 8 horas de remoción por S. maxima
inmovilizada en alginato de sodio a temperatura ambiente (28 ±2ºC), aireación 0.5
vvm, agitación 100 rpm, fotoperiodo natural.
Biosorción de plomo por S. maxima en agar-agar. En las figura 5, se muestra la
variación de la concentración de plomo captada por pellets de agar-agar con biomasa
de S. maxima; se observó que la mayor adsorción sucedió durante las primeras horas
del proceso alcanzando el equilibrio después de 8 horas, teniendo una concentración
final de 15.9 mgL-1 para 20ppm Pb2, mientras que para 10 ppm Pb2 fue de 7.1 mgL-1 y
finalmente para 5ppm Pb2 la concentración captada fue de 4.5X10 mgL-1; los
resultados confirman que la cantidad residual de metal desciende cuanto mayor es el
tiempo de contacto, metal-biomasa inmovilizada, no obstante conforme el soporte y la
biomasa se van saturando de metal la velocidad de biosorción disminuye lo cual
confirma el comportamiento del sistema descrito anteriormente (Sánchez et al., 2008).
897
FIGURA 5. Pb2+ captado en las primeras 8 horas de remoción por S. maxima
inmovilizada en agar-agar a temperatura ambiente (28 ±2ºC), aireación 0.5 vvm,
agitación 100 rpm, fotoperiodo natural.
Porcentaje de metal adsorbido por S. maxima. En la figura 6 se muestran los
porcentajes de remoción de cadmio por biomasa de S. maxima inmovilizada en agaragar, donde se observa claramente que el porcentaje de metal captado es
directamente proporcional a la concentración adicionada alcanzando un 98% de
remoción con 5 ppm Cd+2, mientras que para 10 y 20ppm Cd+2 los porcentajes de
remoción alcanzados fueron de 59% y 19% respectivamente, nótese que estos valores
fueron menores que los obtenidos en los sistema de S. maxima inmovilizada con
alginato de sodio, donde se obtuvieron valores de 100% para 5ppm, 64% y 24% para
10 y 20ppm respectivamente. También se puede observar los porcentajes removidos
de Pb2+ captado por S. máxima en sistemas inmovilizados de alginato de sodio; para
el caso de 5ppm Pb+2 el porcentaje removido fue de 100%, mientras que para 10 y 20
ppm Pb+2 fue de 79% y 83% respectivamente, se puede observar que estos valores
fueron mayores que los resultados obtenidos en los inmovilizados de agar-agar en
donde el porcentaje captado para 5ppm Pb+2 fue de 89%, mientras que para 10 y 20
ppm Pb+2 el porcentaje fue de 71% y 79% respectivamente; este comportamiento
sugiere que S. maxima en sistemas inmovilizados de alginato de sodio presenta mayor
porcentaje de remoción de metal que S. maxima en sistemas inmovilizados de agaragar.
898
FIGURA 6. Porcentaje de remoción de plomo y cadmio por S. maxima
inmovilizada en pellets de alginato de sodio y agar-agar a temperatura ambiente
(28 ±2ºC), aireación continua 0.5 vvm, agitación 100 rpm, fotoperiodo natural.
Conclusiones.
Spirulina maxima inmovilizada en ambos soportes resulto ser un buen biadsorbente de
metales pesados como el cadmio y el plomo en medio acuoso.
El soporte de alginato de sodio presento mayor afinidad por los iones metálicos de
plomo y de cadmio, que el soporte de agar-agar.
Referencias
Ahmet, C., Tamer, A., Sibel, T. and Ozge, T. (2006). Biosorption characteristics of Bacillus sp. ATS-2
immobilized in silica gel for removal of Pb(II) Journal of Hazardous Materials B136: 317±323.
Bayramoglu, G., Bektas, S. and Arica, M. Y. (2006). Biosorption of heavy metal ions on immobilized
white-rot fungus Trametes versicolor. J Hazard Mater, 101:285±300.
Cañizares, V. (2010). Biosorción de metales pesados mediante el uso de biomasa microbiana. Revista
Latinoamericana de Microbiología, 42:131-143
Chen, H. and Pan, S. (2005). Bioremediation potential of Spirulina: toxicity and biosorption studies of
lead. Journal of Zhejiang University SCIENCE ISSN, 6B(3): 171-74.
Converti, A. Lod, A. Solisio, C. Soletto, D. Del Borghi, M. and Carvalho, J. C. M. (2006). Spirulina
platensis Biomass as adsorvent for copper removal. Ciencia y tecnologia alimentaria, 5(2): 85-88.
Doshi, H. Seth, C. Ray A. And Kothari, I. (2008). Bioaccumulation of Heavy Metals by Green Algae.
Curr Microbiol, 56:246±55.
Hong, S., Haldin, C., Lawson, N., Weinstein, B., Dawid, I. and Hukriede, N. (2005).The zebrafish
kohtalo/trap230 gene is required for the development of the brain, neural crest, and pronephric
kidney, 05, 1214-26.
Ilangovan, K., Canizares-Villanueva, S. Gonzalez Moreno and Voltolina, D. (1998). Effect of Cadmium
and Zinc on Respiration and Photosynthesis in Suspended and Immobilized Cultures of Chlorella
vulgaris and Scenedesmus acutus. Bull. Environ. Contam. Toxicol, 60: 936±43.
Sánchez, T. E., Garza, G. M., Almaguer, C. V., Sáenz, T. , Liñán, M. A. (2008). Estudio cinético e
isotermas de adsorción de Ni (II) y Zn (II) utilizando biomasa del alga Chlorella sp. Inmovilizada.
CIENCIA UANL. 10(2): 168-76.
899
Quintelas, C., Rocha, Z., Silva B., Fonseca, B., Figueiredo H. and Tavares T. (2009). Removal of Cd (II),
Cr(VI), Fe(III) and Ni(II) from aqueous solutions by an E. coli biofilm supported on kaolin. Chemical
Engineering Journal, 149: 319±24.
900
CHAPTER 7.23. MARCADORES FENÓLICOS DE ESTRÉS EN PLANTAS IN VITRO
DE Vallisneria americana CON ENRIQUECIMIENTO Y DESBALANCE
AMBIENTAL N/C
Patricia Cruz-Cerino (1), Violeta Ruiz-Carrera*(1); Miguel A.Salcedo-Meza(1); Alberto
J. Sánchez (1); Carlos A. Ledesma Escobar (2); Victor J. Robles-Olvera (3)
(1) Centro de Investigación para la Conservación y Aprovechamiento de Recursos Tropicales (CICART).
Universidad Juárez Autónoma de Tabasco, Villahermosa, Tabasco. México.
(2) Campus Rabanales, Universidad de Córdoba, Córdoba. España.
(3) Unidad de Investigación y Desarrollo en Alimentos (UNIDA). Intituto Tecnológico de Veracruz.
Laboratorio de Enzimología. Veracruz, Veracruz. México.
RESUMEN
En las lagunas costeras del continente americano la angiosperma sumergida
Vallisneria americana y los servicios ambientales asociados se encuentran
amenazados por el cambio global en los ciclos biogeoquímicos de N y C. En
aproximación experimental se analizó en las hojas de V. americana el contenido de
flavanoles, flavonoides y de compuestos fenólicos para definir si son marcadores
metabólicos de estrés por efecto bivariado del incremento de N y del CO 2 en el agua.
Con esta intención fue evaluado en plantas enraizadas in vitro el desbalance ambiental
N/C (N-NO3 y C-HCO3) en relaciones de 80 a 483. Las relaciones N/C se diseñaron en
la columna de agua con la concentración de nutrientes de 12.5, 25 y 50%, del medio de
cultivo MM, la relación 48 a 72 de N/P y 0.15 a 0.6 mM de C-HCO3. Los cultivos en
condición estática permanecieron en ambiente controlado hasta 75 días. Las plantas
murieron en 50% de concentración de nutrientes. Los extractos fueron obtenidos en
80% etanol en la biomasa fresca y congelada de hojas viables. Los contenidos de
flavanoles (método Vainillina-acidificada) y flavonoides (método Cloruro de aluminio)
fueron estimados con el estándar (±) Catequina (Ct) y en fenólicos totales (método
Folin-Ciocalteu) fue Ácido Gálico (AG). En las determinaciones se usaron técnicas de
adición de estándar o evaporación. La linearidad de los tres analitos presentó
coeficientes de regresión y correlación muy altos (r y r2>0.99; p<0.0001). En el
contenido promedio de fenólicos totales (0.36 ± 0.074 mg Eq AG/g) predominaron los
flavonoides (0.33 ± 0.12 mg Eq Ct/g) y fue escaso el contenido de flavanoles. La
relación N/C afectó el contenido de flavonoides (p<0.015). Los flavonoides fueron
marcadores de estrés en V. americana por la variación ambiental N/C que provocó el
enriquecimiento de N.
Palabras clave: angiosperma sumergida, estrés
enriquecimiento de nutrientes, flavonoides, fenólicos
in
vitro,
desbalance
N/C,
---------------*Author for correspondence: [email protected]
ABSTRACT
901
In the coastal lagoons of the American continent the angiosperm submerged Vallisneria
americana and associated environmental services are threatened by the global change
in the N and C biogeochemical cycles. In experimental approach was analyzed in the
leaves of V. americana the contents of flavanols, flavonoids and phenolic compounds to
define if are metabolic markers of stress by effect bivariated of increase of N of CO2 in
the water. With this intention in the water column were exposed plants rooted in vitro the
unbalance environmental N/C (N-NO3 y C-HCO3) in relations N/C of 80 to 483. The
relations N/C are designed with increases in the concentration of nutrients in 12.5, 25
and 50% the medium culture MM, the relations 48 to 72 of N/P and of 0.15 to 0.6 Mm of
C-HCO3.The cultures remained in environment controlled up 75 days. The plants died in
concentration of nutrients from 50%. The extracts were obtained in 80% of ethanol in
the biomass fresh and frozen leaves viable. The contents of flavanols (Vanillin-acidified
method) and flavonoids (Aluminum Chloride method) were estimated with standard of
(±) Catechin (Ct) and in phenolics total (Folin-Ciocalteu method) was Gallic acid (GA).
In the determinations were used techniques of addition of standard or evaporation. The
linearity of the three analytes presented r and r2 very high relation (>0.99;
p<0.0001).The linearity of the three analytes presented very high regression and
correlation coefficients (r y r2>0.99; p<0.0001). In the average content of total phenolics
(0.36 ± 0.074 mg Eq AG/g) predominated the flavonoids (0.33 ± 0.12 mg Eq Ct/g) and
was low flavanol content. The relation N/C affected the content of flavonoids (p<0.015).
The flavonoids were markers of stress in Vallisneria americana by the variation
environmental N/C that causing the enrichment of N.
Key words: angiosperm submerged, stress in vitro, unbalance N/C, enrichment of
nutrients, flavonoids, total phenolics
Introducción
En los ecosistemas limnéticos de Centro y Norte América la angiosperma sumergida
enraizada Vallisneria americana ha presentado marcada declinación en sus
poblaciones por el enriquecimiento de nutrientes o eutrofización, específicamente N y P
(Lauer et al. 2011, Morris et al. 2003a y 2003b, Ruiz-Carrera y Sánchez 2012a). El
efecto de estrés por nutrientes ha sido registrado letal o crónico en el género de
Vallisneria en diferentes regiones ecobiográficas (Morris et al. 2003b). Esta tendencia
global alerta sobre el impacto colateral que el cambio climático global ejerce en la
vegetación acuática derivado del incremento de CO2 ambiental en el agua (Alexandre
et al. 2012, Väisänen et al. 2013, Yan et al. 2006). El interés por preservar las
poblaciones de V. americana radica en las funciones ecológicas vitales que ejercen en
la salud ambiental de los ecosistemas limnéticos tropicales vulnerables al incremento
gradual de nutrientes (Ruiz-Carrera y Sánchez 2012b). En consecuencia, resulta
imperativo desarrollar conocimientos y métodos de evaluación experimental en estrés
abiótico y bioquímica ecológica (Larkum et al. 2006), en particular de V. americana, por
la necesidad de guiar el conocimiento a soluciones biotecnológicas para el manejo y
rehabilitación de los servicios ambientales de los ecosistemas acuáticos.
902
La enorme variedad e importancia de los metabolitos secundarios fenólicos en
diferentes tipos de estrés se ha reconocido ampliamente en las plantas terrestres y
algas marinas (Fallovo et al. 2011). En las angiospermas acuáticas este conocimiento
es escaso y poco figuran las especies sumergidas limnéticas (Bauer et al. 2009,
Larkum et al. 2006, Lodge et al. 1998, Kannan et al. 2010a). Las angiospermas
sumergidas son monocotiledóneas y dicotiledóneas herbáceas que cumplen su ciclo
biológico vegetativo y reproductivo bajo el agua o emergen durante la reproducción (Lot
y Novelo 2004). Las monocotiledóneas sumergidas presentan mayor marginación
científica que las dicotiledóneas, y ambas respecto a la vegetación flotante y emergente
y de otras macrófitas.
La comparación de las estructuras de flavonoides de los pastos marinos Enhalus
acoroides, Halophila stipulacea, Posidonia oceanica y Thalassia hemprichii sugirió que
algunos podrían ser utilizados como marcadores quimiotaxonómicos para diferenciar
entre los géneros de estas especies marinas (Qi et al. 2012). También la filogenia
molecular con base en la concentración de algunos metabolitos fenólicos se ha
determinado en los morfotipos de los géneros Enhalus, Thalassia y Halophila de la
familia Hidrocharitaceae (Attaway et al. 1970, Heglmeier y Zidom 2010, Qi et al. 2012,
Lason et al. 2012).
En su papel de defensa primaria los metabolitos fenólicos incluyen elevados
potenciales antioxidantes y la inactivación de radicales libres (Soobratte et al. 2005). En
el mecanismo de defensa han sido involucrados compuestos fenólicos como ácidos
carboxílicos fenólicos (no flavanólicos) y los del grupo de flavonoides (flavanólicos)
hidroxilados tales como flavonas, flavonoles, flavanoles, leucocianidinas, antocianinas y
procianidinas, así como los isoflavonoides (Matkowski y Wosniak 2005, Treutter 2006).
Específicamente los flavonoides son los metabolitos fenólicos más involucrados en la
resistencia de la planta contra el estrés ambiental abiótico (Ostrofsky y Zettler 1986). La
función adjudicada a los metabolitos fenólicos es la protección contra la oxidación
excesiva causada por radiación ultravioleta (UV), oxidantes químicos, ataque de
patógenos y de otros tipos de estrés, como la deficiencia de nitrógeno (Stewart et al.
2001). Aunque la función protectora de los flavonoides al estrés por luz ultravioleta (UVB) es la más estudiada (Treutter 2006). En las angiospermas dulceacuícolas su función
comienza a ser reconocida, por ejemplo algunos compuestos polifenólicos como
ésteres de flavonoles se han identificado como metabolitos de defensa importantes
(Kliebenstein 2004).
El estudio de los compuestos fenólicos totales, en grupo e individuales, o en
clases y subclases como biomarcadores de la calidad del medio ambiente ha
incrementado (Bauer et al. 2009, Sieg y Kubanek 2013, Ferrat et al. 2003, Ashraf et al.
2009). En particular los flavonoides son investigados ya que tienen el potencial de ser
confiables (Kaundun et al. 1998, Cannac et al. 2006, Ashraf et al. 2009). Sin embargo,
por la multifuncionalidad de los flavonoides y la heterogeneidad del ecosistema a
menudo se complica la interpretación de los resultados experimentales de campo
(Treutter 2006).
Reiteradamente se ha relacionado el incremento de la concentración de
flavonoides con la limitación de nitrógeno y fósforo, empero las funciones en respuesta
a la limitación y en especial al enriquecimiento de nutrientes en las plantas es incipiente
y oscuro (Stewart et al. 2001, Scheible et al. 2004, Misson et al. 2005, Lea et al. 2007,
903
Morcuende et al. 2007, Müller et al. 2007). Aún más, la respuesta al estrés abiótico en
macrófitas puede presentar divergencias metabólicas en referencia a la capacidad que
presentan de asimilar nutrientes por hojas y raíces debido a su condición sumergida.
Por lo tanto, el estrés ambiental por exceso de nutrientes y CO 2 disuelto en las plantas
sumergidas enraizadas requiere resolución de varias interrogantes: a) ¿los flavonoides
son un mecanismo de defensa por incrementos de N y C acuático?, b) ¿influye en la
biosíntesis de flavonoides el balance ambiental N/C? y c) ¿representa el contenido de
flavonoides un biomarcador potencial de estrés ante el incremento de nutrientes?.
El sistema experimental in vitro representa una oportunidad de determinar sí los
flavonoides u otros compuestos fenólicos funcionan como biomarcadores de estrés en
la monocotiledónea Vallisneria americana con la variación de los nutrientes (N y P) y el
carbono inorgánico disuelto (CID). El CID aestá constituido por CO2 libre (= H2CO3 +
CO2 disuelto), el HCO3- y CO3-2, que son especies inorgánicas de C interconvertibles,
presentes en aguas naturales (Margalef 2005). En este contexto, la investigación
experimental reduccionista usando técnicas de cultivo in vitro ofrece ventajas por su
pequeña escala y por el control microbiótico y de los factores ambientales (Davey y
Anthony 2010).
La búsqueda de biomarcadores de estrés para descifrar la respuesta de defensa
de V. americana planteó la hipótesis que el desbalance ambiental N/C desencadena
efectos de estrés que modifican el contenido total de compuestos fenólicos,
especialmente de flavonoides. Lo anterior porque el metabolismo de N mantiene
interacciones importantes con diferentes vías metabólicas como: 1) la fotosíntesis, 2) la
interconversión de glutamina y asparagina, vía asparagina sintetasa que balancea el
metabolismo C/N dentro de la planta, 3) la asimilación de novo del NH4, y 4) la
exportación, almacenaje en la vacuola para biosíntesis de proteínas y de metabolitos
secundarios, como los flavonoides, entre otras (Ruiz-Carrera y Sánchez 2012b).
El objetivo de esta investigación fue determinar el efecto del desbalance N/C en la
columna de agua en el incremento foliar de flavanoles, flavonoides y fenólicos totales
para predecir si funcionan como biomarcadores de estrés por enriquecimiento de
nutrientes y el incremento de carbono inorgánico en el agua.
Especie y recolección de semillas. La monocotiledónea Vallisneria americana Michx
(Hidrocharitaceae) es una planta angiosperma, enraizada, dioica y perenne de aguas
limnéticas o ligeramente salobres (Novelo 2006, Korschgen et al. 1997). La planta
presenta vida sumergida, hojas lineales en forma de cinta y flexibles que se extienden
hasta 2 m o más. Las hojas surgen de pequeños tallos laterales erectos llamados
brotes cortos o rizoma ortotrópico. La colonización ocurre por semillas y estolones
(rizoma plagiotrópico). Las semillas se producen en vainas que emergen en la
superficie del agua. En este estudio la procedencia de semillas fue el área de la
Reserva de la Biósfera Pantanos de Centla ( ¶´1 ¶´:). La vaina
fue disectada con escalpelo y las semillas almacenadas en refrigeración con agua
purificada.
904
Cultivo inicial de plantas. Las semillas de Vallisneria americana se desinfectaron y
cultivaron en medio bifásico estéril siguiendo el protocolo de Ruiz-Carrera y Sánchez
(2012a). Las plantas desarrollaron: a) sumergidas en la columna de agua de 4 cm y b)
enraizadas en la fase de soporte de 4% de agar-agar de 1 cm. Ambas fases fueron
preparadas de agua semidura artificial. La unidad de cultivo fue de vidrio de capacidad
de 300 mL (6.5 cm de alto x 5.5 cm de ancho) con tapas de policarbonato (Magenta ®).
El tiempo de cultivo para regenerar la planta enraizada fue de dos meses y medio. Las
plantas desarrollaron en ambiente de luz tenue, fotoperiodo de 16/8 luz: oscuridad y
30º C.
Ensayo experimental. En los cultivos in vitro bifásicos de plantas enraizadas se
evaluaron nueve relaciones N/C, en el rango de 80 a 483. Las relaciones N/C se
diseñaron con incrementos en la columna de agua de la concentración de nutrientes de
12.5, 25 y 50% de medio de cultivo MM, la relación molar 48 a 72 de N/P y 0.15 a 0.6
mM de C-HCO3 (Tabla 1). El MM fue el de MS (Murashige y Skoog 1962) modificado
por Ruiz-Carrera y Sánchez (2012b) con el NaHCO3 incluido en la concentración de
nutrientes. La relación N/C se modificó con la pesada directa de KNO3 y NaHCO3 en la
solución de cultivo. Las plantas permanecieron en ambiente controlado hasta 75 días.
El cultivo fue intermitente y estático. La columna de agua de las plantas in vitro fue
renovada cada 15 días con el medio de cultivo experimental.
TABLA 1. Diseño experimental del medio de cultivo de Vallisneria americana con
variación en la concentración de nutrientes (CN) en la relación ambiental N/C (NNO3 y C-HCO3).
Factores
Concentración
nutrientes (MM)
Niveles
de
Concentración de N
(mM)
Concentración de C
(mM)
Relación molar N/C (NNO3/C-HCO3)
12.5%
25%
50%
48, 60, 72
48, 60, 72
48, 60, 72
0.15
0.3
0.6
160, 200, 240
80,100,120
322,
483
402,
Análisis de flavanoles, flavonoides y fenólicos totales
Procesamiento de muestra y extracto. Los extractos fueron obtenidos en 80% etanol
en la biomasa fresca congelada (bf) de hojas viables. Las hojas de plantas in vitro
fueron separadas de las raíces con un bisturí realizando el corte al ras del gel. Las
hojas íntegras permanecieron congeladas tres meses a -5 ºC antes de obtener el
extracto. Las muestras congeladas se pesaron en balanza analítica y se molieron en
mortero. El extracto foliar se obtuvo con 20 mL de etanol al 80% y fue sonicado durante
cinco minutos. El sobrenadante se filtró en papel filtro Whatman 40. El extracto crudo
se almacenó en el congelador al abrigo de la luz.
905
Reactivos
Reactivo de Folin-Ciocalteu (Hycel), Na2CO3 anhidro (Baker), Ácido gálico (3,4,5Trihydroxybenzoic
acid,
Sigma),
reactivo
de
Vainillina
(4-Hydroxy-3methoxybenzaldehyde, Sigma, H2SO4 (Merck), (±) Catequina hidratada (Sigma),
NaNO2 (Baker), AlCl3.6H2O (GoldenBell), NaOH (Baker), Etanol absoluto anhidro
(Baker).
Determinaciones analíticas. Las determinaciones de flavanoles, flavonoides y de
fenólicos totales fueron espectrofotométricas. El análisis de flavanoles se basó en el
método de la vainillina-acidificada en ácido sulfúrico (Dvorakova et al. 2008) y el de
flavonoides en el método del complejo flavonoide-aluminio (Bakar et al. 2009). En los
dos casos se modificaron los métodos con la técnica de estándar añadido a la muestra
usando (±) Catequina (Ct) para mejorar la detección del analito (Quattrocchi et al.
1992). La absorbancia de las muestras se midió en espectrofotómetro (Thermo
Scientific Genesys 6) a 500 nm y los resultados se expresaron como mg de
equivalentes de (±) Catequina (Ct) por gramo de biomasa fresca (mg ECt/g bf). Las
curvas de calibración de flavanoles y flavonoides se construyeron con el estándar (±)
Catequina (Ct). La concentración del estándar (mg/mL) fue de 0.01 y el rango lineal
ȝJP/de 0.1 a 1 para los flavanoles; y en las mismas unidades fue de 1 con rango
lineal de 0.125 a 1.25 en flavonoides. Para analizar el contenido de fenólicos totales se
usó el método de Folin-Ciocalteu (Ledesma 2011), el espectrofotómetro Beckman en
longitud de onda 765 nm, estándar de Ácido Gálico (AG) de PJP/\DȝJP/HQ
el rango lineal. En este caso el extracto foliar fue concentrado 40 veces a 40ºC y
centrifugado a 14,000 rpm durante 10 min (Centrifugue 5810 R). El contenido de
flavanoles, flavonoides y fenólicos se expresó en mg de equivalentes del estándar por
gramo de biomasa fresca (mg std/g bf).
Análisis estadísticos. Los promedios de flavanoles, flavonoides y fenólicos totales se
obtuvieron de tres repeticiones por muestra. Los datos fueron analizados en normalidad
(Prueba Kolmogorov-Smirnov) y homogeneidad (Prueba Cochran) (Zar 1999). El
análisis de varianza fue no paramétrico de Kruskall Wallis. La significancia estadística
(p) fue p<0.05. El coeficiente r y r2 y la p de las curvas de calibración fue obtenido del
modelo de regresión lineal (y= a+ mx) (Zar 1999). Todos los análisis y gráficos se
realizaron en el programa Statistica 8.0 (Stat Soft, Inc).
Resultados
Resistencia al estrés. Las plantas cultivadas en concentración de nutrientes de 50%
murieron. En las relaciones de N/C (N-NO3 y C-HCO3) de 160-483 resistieron el estrés
presentando fenotipos de hojas viables y fotosintéticas (Figura 1).
906
50 %
N/C
25%
12.5 %
a
b
c
d
e
f
48
60
72
FIGURA 1. Fenotipos foliares de Vallisneria americana con desbalance en la
relación N/C a 75 días en cultivo bifásico. Orden concentración de nutrientes
(derecha-izquierda): 12.5% (a,c,e) y 25% (b,d,f). Orden en relación N/C (arriba-abajo):
48 (a,b), 60 (c,d) y 72 (e,f).
Contenido de flavanoles, flavonoides y fenólicos totales. El contenido promedio
de flavanoles, flavonoides y fenólicos totales de las hojas in vitro se presenta en la
Tabla 2. El peso de la biomasa foliar osciló entre 12-55 mg (n=38). El contenido de
flavanoles representó el 6.1% de los flavonoides y el de flavonoides el 92% de los
compuestos fenólicos totales.
TABLA 2. Contenido promedio de flavanoles flavonoides y fenólicos totales en
biomasa fresca de hojas de Vallisneria americana
Compuestos
Flavanoles (mg EC/g)
Flavonoides (mg EC/g)
Fenólicos (mg EAG/g)
Regresiones lineales. La Figura 2 muestra la dinámica lineal de la regresión obtenida
en las curvas de calibración de flavanoles, flavonoides y fenólicos totales. El ajuste
lineal de los analitos fue excelente con valores de r y r2 que excedieron el 0.99 en
significancia estadística elevada (p=0.0001).
1,0
a
Absorbancia (500 nm)
0,30
Absorbancia (500 nm)
Promedio ± D.S.
0.02 ± 0.002
0.33 ± 0.12
0.36 ± 0.074
0,24
0,18
0,12
0,06
y=0.0094+0.2893x
b
0,8
0,6
0,4
0,2
y=0.0002+0.7947x
x
0,00
0,2
0,4
0,6
0,8
µg ml-1 (± Catequina)
1,0
0,0
0,2
0,4
*x x
µg ml
0,6
-1
0,8
(± Catequina)
1,0
1,2
907
c
y=0.0093+0.0626x
x
xx
FIGURA 2. Curvas de calibración de flavanoles. (a) r=0.996, r2=0.993 y p=0.0001,
flavonoides (b) r=0.999, r2=0.999 y p=0.0001, y fenólicos totales(c) r=0.999, r2=0.998 y
p=0.0001.
Biomarcadores de estrés. La variación de N/C en la composición del medio de
cultivo no mostró cambios significativos en el contenido foliar de fenólicos totales
(F, p=0.365) y de flavanoles (KW, p=0.074); en cambio afectó en el contenido de
flavonoides (KW, p=0.015). El contenido de flavonoides fue más alto en N/C de
160 a 240 y contrastó con las de 322 a 483 que fueron diferentes y similares
(p>0.05). Las relaciones N/C de 160, 200 y 240, todas de la concentración de
nutrientes de 25% de MM, registraron promedios similares (Figura 3).
Contenido de flavonoides (µg/ g BF)
600
500
400
300
908
200
160
200
240
322
Relación N/C
402
483
FIGURA 3. Contenido promedio de flavonoides en biomasa de hojas de
Vallisneria americana con variación en la relación N/C en el medio de cultivo.
Promedio ± desviación estándar.
Discusión
La factibilidad de utilizar los contenidos de compuestos fenólicos totales y de la clase
flavonoides y subclase flavanoles como biomarcadores de estrés por el enriquecimiento
de N-NO3, acoplado al incremento C-HCO3, en las hojas de Vallisneria americana se
evaluó en aproximación in vitro como Su utilidad como posibles biomarcadores foliares
fue descartada en los ambientes que presentaron relaciones molares N/C de 80 a 120,
ya que las plantas murieron por estrés osmótico (Bautista 2013). En estas relaciones
críticas el nivel de enriquecimiento de nutrientes fue de 50%, mismo que ubica el límite
inferior oligotrófico a escala de ecosistema.
En concentraciones menores de 50% con variación N/C de 160 hasta 483 fue
cuantificado el contenido de fenólicos totales, flavonoides y flavanoles, sin embargo, la
variación N/C solo afectó el subgrupo de flavonoides. La producción de flavonoides fue
mayor en concentración de nutrientes de 25% que en 12.5% y fue notable en la
relación N/C de 240. En situaciones de estrés el incremento en la composición de los
fenoles de P. oceanica se ha atribuido al mecanismo de protección de la planta contra
las especies reactivas de oxígeno, presiones antropogénicas y competencias
interespecíficas (Castellano et al. 2012). Por lo tanto, el incremento de los flavonoides
podría estar ligado al estrés osmótico en la concentración salina de 25%.
El contenido de flavonoides en las relaciones N/C de 160 a 240, coincidió con
investigaciones que han señalado incremento de flavonoides determinados en
respuesta a la disminución de N en el medio de cultivo (Wang et al. 2012), aunque la
integración del C frecuentemente no ha sido citada. En constraste, la manipulación de
N y P ha sido usada con la intención de reducir los niveles de estos compuestos para
mejorar la calidad de las plantas con fines agronómicos. En las investigaciones el
argumento científico se ha basado en la fuerte presencia de N en los precursores
biosintéticos de flavonoides en la ruta metabólica del ácido shikímico. Sin embargo, los
efectos de la deficiencia de N y P en el nivel de los genes transcritos en la vía
metabólica del ácido shikímico son complejos e imprecisos (Lea et al. 2007). En
oposición, los genes de la vía de los flavonoides han mostrado recientemente
respuestas consistentes a la deficiencia de nutrientes.
El contenido de compuestos fenólicos foliares de V. americana en este estudio fue
mayor que los datos señalados en diferentes investigaciones; por ejemplo, Enhalus
acoroides y Halodule pinifolia registraron concentraciones de 0.323 y 1.080 mg/g Eqácido tánico. El contenido de flavonoides en Eichornia crassipes fue de 1.524 mg/g Eqrutina.
En las hojas de Vallisneria americana el contenido de flavonoides reveló
información científica importante para estimar el efecto del estrés por eutrofización.
Aunque los rizomas representan otra alternativa más con base en rangos elevados
909
registrados en el rizoma distal, intermedio y basal de Posidonia oceanica (Migliore et al.
2007). Las ventajas de usar el contenido de fenoles de rizoma, respecto a las hojas,
como marcadores de salud en el ecosistema fue recomendado por presentar ciclo de
vida amplio, menos fluctuaciones marcadas y para revelar la memoria de la expresión
ambiental.
Los flavonoides, y sus flavanoles, están representados por una familia química
compleja de diferentes formas isoméricas de monómeros y polímeros (Cheynier et al.
2000), por lo cual los problemas en el análisis han surgido debido a la diversidad
estructural, solubilidad e inestabilidad de los compuestos fenólicos. Los flavanoles
predominantes son: (+)-catequina), (-) epicatequina), (+)-galocatequina y (-)epigalocatequina y los siguientes ésteres de ácido gálico:(-galato) -epicatequina y (-) el
galato de epigalocatequina (Holman y Ilja 2000). En V. americana la presencia de
flavanoles en equivalentes de catequina fue analizada como parte de la composición
total de los flavonoides.
Una limitante en las determinaciones analíticas fue que la reducida cantidad de
biomasa foliar inherente a los experimentos in vitro, por consiguiente el reto en la
extracción adecuada de los compuestos fenólicos fue mayor (Herrera y Luque de
Castro 2004, Ledesma 2011). Además, en las determinaciones analíticas fue necesario
aplicar técnicas de evaporación y estándar añadido en los extractos etanólicos para
mejorar la detección espectrofotométrica (Quattrocchi 1992). Con estas modificaciones
analíticas el contenido de flavonoides registrado en las hojas de Vallisneria americana
fue de 92% respecto a los fenólicos totales. La evaporación probablemente influyó en la
solubilidad de los compuestos en el análisis de fenólicos totales y por ende afectó su
contenido. Tampoco se descartó la sobreestimación de flavonoides y flavanoles con la
técnica del estándar añadido.
En este estudio, la revelación de compuestos fenólicos totales, flavonoides y
flavanoles foliares de plantas in vitro en V. americana analizado con el incremento en
las relaciones molares N/C fue información novedosa y útil para enfocar su usó en
biomarcadores de estrés por enriquecimiento de nutrientes, aunque los contenidos
registrados son susceptibles de mejorar en investigaciones futuras con la optimización
del método de obtención del extracto y la cuantificación de compuestos particulares por
HPLC.
Conclusión
En dimensión in vitro la señal de estrés expresada con el incremento de flavonoides
foliares en plantas de V. americana fue revelada con cambios en la relación de N/C de
concentraciones de nutrientes oligosalinas. Por lo tanto es factible usar como
biomarcadores el contenido de flavonoides para diagnosticar el estado de salud de los
ecosistemas acuáticos limnéticos que se encuentran amenazados por el
enriquecimiento de nutrientes y el efecto del incremento de C acuático por el cambio
climático.
Agradecimientos
Los autores agradecemos el apoyo del Instituto Tecnológico de Veracruz en los análisis
químicos y a la UJAT por el financiamiento del proyecto PFICA-UJAT-2011-C07-30.
910
Referencias
Alexandre, A.; Silva, J.O.; Buapet, P.; Bjork, M.; Santos, R. (2012). Effects of CO2 enrichment on
photosynthesis, growth, and nitrogen metabolism of the seagrass Zostera noltii. Ecology and
Evolution 2 (10):2620-2630.
Attaway, D.H.; Parker, P. L.; Mears, J.A. (1970). Normal alkanes of five coastal spermatophytes.
University of Texas Marine Science 15:13-19.
Azcón-Bieto, J; Talón, M. (2008). Fundamentos de Fisiología Vegetal. Capítulo 29: Fisiología de las
plantas y el estrés (2 ed.) Interamericana-McGraw-Hill, Madrid 577-597 pp.
Ashraf, M. (2009). Biotechnological approach of improving plant salt tolerance using antioxidants as
markers. Biotechnology Advances 27:84-93.
Bakar, M.F.A.; Mohamed, M.; Rahmat, M.; Fry, F. (2009). Phytochemicals and antioxidant activity of
different parts of bambangan (Mangifera pajang) and tarap (Artocarpus odoratissimus). Food
Chemistry 113:479-483.
Bauer, N.; Blaschke, U.; Beutler, E.; Gross, E.M.; Jenett-Siems, K.; Siems, K.; Hilt, S. (2009). Seasonal
and interannual dynamics of polyphenols in Myriophyllum verticillatum and their allelopathic activity
on Anabaena variabilis. Aquatic Botany 91:110-116.
Bautista R J (2013). Bioensayos nutricionales para inducir in vitro el crecimiento estolonífero de la
angiosperma acuática Vallisneria americana. Tesis de licenciatura en Biología. Universidad Juárez
Autónoma de Tabasco. Villahermosa, Tabasco.
Cannac, M.; Ferrat, L.; Pergent-Martini, C.; Pergent, G.; Pasqualini, V. (2006). Effects of fish farming on
flavonoids in Posidonia oceanica. Science of the Total Environment. 370:91-98.
Castellano, G.; Tena, J.; Torrens, F. (2012). Classification of phenolic compounds by chemical structural
indicators and its relation to antioxidant properties of Posidonia oceanica (L.) Delile. MATCH
Communications in Mathematical and in Computer Chemistry 67:231-250.
Cheynier, V.; Remy, S.;Fulcrand, H. (2000). Mechanisms of anthocyanin and tannin changes during
winemaking and aging. In: The ASV 50th Anniversary Annual Meeting, Rautz, J., (Eds.), ASEV,
Davis, CA, 337-344.
Dvorakova, M.; Moreira, M.; Dostalek, P.; Skulilova, Z.; Guido, L.F.; Barros A.A. (2008). Characterization
of monomeric and oligomeric flavan-3-ols from barley and malt by liquid chromatography-ultraviolet
detection-electrospray ionization mass spectrometry. Journal of Chromatography A. 1189 (12):398-405.
Davey, M.R.; Anthony, P. (2010). Plant Cell Culture: Essential Methods, John Wiley y Sons, Ltd,
Chichester, UK.
Fallovo, C.; Schreiner, M.; Schwarz, D.; Colla, G.; Krumbein, A. (2011). Phytochemical changes Induced
by different nitrogen supply forms and radiation levels in two leafy Brassica species. Journal of
Agricultural and Food Chemistry 59 (8):4198-4207.
Ferrat, L.; Pergent-Martini, C.; Romeo, M. (2003). Assessment of the use of biomarkers in aquatic plants
for the evaluation of environmental quality: application to seagrasses. Aquatic Toxicology 65:187204.
Heglmeier, A.; Zidorn, C. (2010). Secondary metabolites of Posidonia oceanica (Posidoniaceae)
Biochemical Systematics and Ecology 38:964-970.
Hollman, PC.; H, Arts., Ilja, C.W. (2000). Flavonols, flavones and flavanols-nature, occurrence and dietary
burden. Journal of the Science of Food and Agriculture 80 (7):1081-1093.
Herrera, M.C.; Luque de Castro, M.D. (2004). Ultrasound-assisted extraction for the analysis of phenolic
compounds in strawberries. Analytical and Bioanalytical Chemistry 379 (7-8):1106-1112.
Kannan, R.R.R.; Arumugam, R.; Meenakshi, S. (2010a). Thin layer chromatography analysis of
antioxidant constituents from seagrasses of Gulf of Mannar Biosphere Reserve, South India.
International Journal of Chemical Technology Research (3):1526-1530.
Kaundun, S.S.; Lebreton, P.; Fady, B. (1998). Geographical variability of Pinus halepensis Mill. As
revealed by foliar flavonoids. Biochemical Systematics and Ecology. 26:83-96.
Kliebenstein, D.J. (2004). Secondary metabolites and plant/environment interactions: a view through
Arabidopsis thaliana tinged glasses. Plant Cell Environment 27:675-684.
911
Korschgen, C.E.; Green, W.L.; Kenow, K.P.; (1997). Effects of irradiance on growth and winter bud
production by Vallisneria americana and consequences to its abundance and distribution. Aquatic
Botany 58 (1): 1-9.
Lauer, N.; Yeagera, M.; Kahna, E.A.; Dobberfuhlb, D.R.; Rossa, C. (2011). The effects of short term
salinity exposure on the sublethal stress response of Vallisneria americana Michx
(Hidrocharitaceae). Aquatic Botany 95:207-213.
Lodge, D.M., Cronin, G.; van Donk, E.; Froelich, A.J. (1998). Impact of herbivory on plant standing crop:
comparisons among biomes, between vascular and nonvascular plants, and among freshwater
herbivore taxa. Springer New York.
Lason, G.R.: Dicke, M.; Hartley, S.E. (Eds.). (2012). The ecology of plant secondary metabolites: from
genes to global processes. Cambridge University Press.
Larkum, A.W.D; Orth, R.J.; Duarte, C.M. (2006). Seagrasses: Biology, Ecology, and Conservation.
Springer, Berlin 409.
Lea, U.S.; Slimestad, R.; Smedvig, P.; Lillo, C. (2007). Nitrogen deficiency enhances expression of
specific MYB and bHLH transcription factors and accumulation of end products in the flavonoid
pathway. Planta 225 (5):1245-1253.
Lot, A.; Novelo, A. (2004). Iconografía y estudio de plantas acuáticas de la ciudad de México y sus
alrededores. México, D.F: Universidad Nacional Autónoma de México.
Ledesma Escobar CA (2011). Efecto de la extracción sobre las características antioxidantes de los
extractos obtenidos de cortezas de limón persa Citrus latifolia Tanaka. Maestría en Ciencias en
Ingeniería Bioquímica. Instituto Tecnológico de Veracruz.
Migliore, L.; Rotini, A.; Randazzo, D.; Albanese, N.; Giallongo, A. (2007). Phenols content and 2-D
electrophoresis protein pattern: a promising tool to monitor Posidonia meadows health state. BMC
Ecology 7(1):6.
Murashige, T.; Skoog, F. (1962). A revised medium for rapid growth and bio assays with tobacco tissue
cultures. Physiologia Plantarum 15 (3):473-497.
Matkowski, A.; Wozniak, D. (2005). Plant phenolic metabolites as the free radical scavengers and
mutagenesis inhibitors. BMC Plant Biology 5: (Suppl 1):S23.
Margalef, R. (2005). Ecología. Ediciones Omega. 951 pp.
Morris, K.; Bailey, C.P.; Boon, PI.; Hughes, L. (2003a). Alternative stable states in the aquatic vegetation
of shallow urban lakes I. Effects of plants harvesting and low-level nutrient enrichment. Marine and
Freshwater Research, 54: 185-200.
Morris, K.; Bailey, CP.; Boon, PI.; Hughes, L. (2003b). Alternative stable states in the aquatic vegetation
of shallow urban lakes II. Catastropic loss of aquatic plants consequent to nutrient enrichment.
Marine and Freshwater Research, 54 (3):201-215.
Misson, J.; Raghothama, K.G.; Jain, A.; Block, M.A.; Blingny, R.; Ortet, P.; Creef, A.; Somerville, S.
Rolland, N.; Doumas, P.; Nacry, P.; Estrella, L.H.; Nussaume, L.; Thibaud, M.C. (2005). A genomewide transcriptional analysis using Arabidopsis thaliana affymetrix gene chips determined plant
responses to phosphate deprivation. Proceedings of the National Academy of Sciences of the
United States of America 102 (33):11934-11939.
Morcuende, R.; Bari, R.; Gibon, Y.; Zheng, W.; Pant, B.D.; Blansing, O.; Usadel, B.; Czechowski, T.;
Udvardi, MK.; Stitt, M.; Scheible, W.R. (2007). Genome-wide reprogramming of metabolism and
regulatory networks of Arabidopsis in response to phosphorus. Plant, Cell Environment 30 (1):85112.
Müller, R.; Morant, M.; Jarmer, H.; Nilsson, L.; Nilsen, T.H. (2007). Genome-Wide analysis of the
Arabidopsis leaf transcriptome reveals interaction of phosphate and sugar metabolism. Plant
Physiology 143 (1):156-171.
Novelo, A. (2006). Plantas acuáticas de la Reserva de la Biosfera Pantanos de Centla. Espacios
Naturales y Desarrollo Sustentable, A.C. Villahermosa, México. 260 pp.
Ostrofsky, M.L.; Zettler, E.R. (1986). Chemical defenses in aquatic plants. The Journal of Ecology.
74:279-287.
Qi, S.H.; Huang, L.S.; He, F.; Zhang, S.; Dong, J.D. (2012). Phytochemical and chemotaxonomic
investigation of seagrass Thalassia hemprichii (Ehrenb) Aschers (Hidrocharitaceae). Biochemical
Systematic and Ecology. 128-131.
Quattrocchi, O.A.; De Andrizzi, S.A.; Laba, R.F. (1992). Introducción a la HPLC: aplicación y práctica.
Buenos Aires: Artes Gráficas Farro.
912
Ruiz-Carrera, V.; Sánchez, A.J. (2012a). Estrategias experimentales y repoblación de angiospermas
sumergidas en un humedal fluvial en la zona costera del sureste de México in Sánchez A.J.;
Chiappa-Carrara X. (eds). Recursos acuáticos costeros del sureste: tendencias actuales en
investigación y del arte. CONCYTEY. RECORECOS, UNACAR, UJAT. ISBN 978-607-9060-06-0
(obra completa).
Ruiz-Carrera, V.; Sánchez, A.J. (2012b). Estrategias de propagación en Vallisneria americana:
experimentos in vitro en Vallisneria americana, Editorial Académica Española, ISBN978-3-659
059711.
Soobrattee M.A; Neergheen, V.S; Luximon-Ramma, A; Arouma, O.I; Bahorun, T. (2005). Phenolic as
potential antioxidant therapeutic agents: Mechanism and actions. Mutation Research/ Fundamental
and Molecular Mechanisms of Mutagenesis. 579:200-213.
Scheible, W.R.; Morcuende, R.; Czechowski, T.; Fritz, C.; Osuna, D.; Palacios-Rojas, N.; Schindelasch,
D.; Thimm, O.; Udvardi, MK.; Stitt, M. (2004). Genome-wide reprogramming of primary and
secondary metabolism, protein synthesis, cellular growth processes, and the regulatory
infrastructure of Arabidopsys in response to nitrogen. Plant Physiology 136:2483-2499.
Stewart, A.J.; Chapman, W.; Jenkins, G.I.; Graham, T.; Martin, T.; Crozier A. (2001). The effect of
nitrogen and phosphorus deficiency on flavonol accumulation in plant tissue. Plant, Cell and
Environment 24:1189-1197.
Sieg, R.D.; Kubanek, J. (2013). Chemical Ecology of Marine Angiosperms: Opportunities at the Interface
of Marine and Terrestrial Systems. Journal of Chemical Ecology, 39(6): 687-711.
Treutter, D. (2006). Significance of flavonoids in plant resistance: a review. Environmental Chemistry
Letters 4 (3):147-157.
Väisänen, M.; Martz, F.; Kaarlejärvi, E.; Julkunen-Tiitto, R.; Stark, S. (2013). Phenolic responses of
mountain crowberry (Empetrum nigrum ssp. hermaphroditum) to global climate change are
compound specific and depend on grazing by reindeer (Rangifer tarandus). Journal of Chemical
Ecology, 39, 1390-1399.
Wang, Y.; Gao, G.; Qin, B.; Wang, X. (2012). Ecophysiological and anatomical responses of Vallisneria
natans to nitrogen and phosphorus enrichment Knowledge and Management of Aquatic
Ecosystems (5):405.
Yan, X.; Yu, D.; Li, Y.K. (2006). The effects of elevated CO2 on clonal growth and nutrient content of
submerge plant Vallisneria spinulosa. Chemosphere 62 (4): 595-601.
Zar, J.H. (1999). Biostatistical analysis. Prentice-Hall Inc. Upper Saddle River. 663 pp
913
Section 8.
Solid Waste Management and Treatment
914
Page
Chapter 8.1. The electronic waste in Mexico: a review
Christian E. Hernández-Mendoza
916
Chapter 8.2. Biodegradation of plastic films by anaerobic digestion
Gabriela A. Santiago-Juárez; Alethia Vázquez-Morillas;
Maribel Velasco-Pérez; Margarita Beltrán-Villavicencio
934
Chapter 8.3. Biodegradation of Ecovio® compostable plastic by composting
and bioaugmentation
Xochitl Quecholac-Piña; Mariel A. García-Rivera;
Rosa M. Espinosa-Valdemar; Alethia Vázquez-Morillas;
Margarita Beltrán-Villavicencio
944
Chapter 8.4. Suelo y residuo como tema obligado en programas de
educación ambiental
Dante Camarillo-Ravelo; Diana G. Castro-Frontana;
Katia J. López-Mendieta
954
Chapter 8.5. Aprovechamiento de residuos de jitomate
(Lycopersicon esculentum) hidropónico como sustrato para la
producción del hongo seta (Pleurotus ostreatus)
Alma D. Sánchez-García; Rosa A. Guillen-Garcés;
Karla González-Velázquez
965
Chapter 8.6. Hydrocarbon elimination from residual sludge obtained in
a wash automotive center
Verónica Lavadores; Sergio E. Vigueras-Carmona;
Graciano Calva-Calva; Josefina Pérez-Vargas
977
Chapter 8.7. Application of a biological treatment, after Fenton treatment, in a
solid waste mature leachate
Marbella Amador-Cruz; Patricia E. Cruz-Ortega;
Rosa Salgado-Brito; Gabriel Pineda-Flores
983
Chapter 8.8. Impact of the spreading of urban waste on agricultural
soil bacterial communities in the periphery of Ouagadougou, Burkina Faso
Edmond Hien; Sabine Favre-Bonté; Salomon Bouda;
Aboubakar S. Ouattara; Dominique Masse; Sabine Houot; Sylvie Nazaret
990
915
CHAPTER 8.1. THE ELECTRONIC WASTE IN MEXICO: A REVIEW
Christian E. Hernández-Mendoza (1)*
(1) Universidad del Mar, Campus Puerto Ángel, Oaxaca, México.
ABSTRACT
Electronic waste (e-waste) contains heavy metals (e. g., lead, mercury) and other
hazardous materials with severe environmental and health risk associated with its
incorrect handling. The components of the electric and electronic equipment are
not hazardous while they are into its cases. However, once they get mixed with the
municipal waste they can be broken and the hazardous materials are released or
leachated causing negative impacts.
The e-waste production worldwide was 49 million tons in 2012 and it is
expected to rise 33% by 2017. In Mexico, the e-waste production increased 167%
from 1998 to 2006. In 2012 the e-waste production in the country was 1.138 million
tons with an average production per inhabitant of 9 kg/year. This accelerated
growth results in high e-waste quantities that cannot be treated as municipal solid
wastes.
In Mexico, the official studies are mainly focused on e-waste generation at the
north and central part of the country. While in the south part of the country there is
not official data about e-waste production. The e-waste management in Mexico is
as follows: 90% is stored (e. g., into houses, offices, storehouses), 4% is recycled,
and 1% is exported. The national law does not consider the e-waste as hazardous.
Instead, it is considered as special handling waste that is subject to waste
management plans for its adequate handling. Nonetheless, 5% of this waste is
disposed of with municipal waste. It is estimated that these wastes are responsible
for 70% of the toxic emissions in landfills. Although the Mexican Official Norm
NOM-161-SEMARNAT-2011 sets the criteria to classify the special wastes, it is not
specialized in e-waste management. Another factor affecting the e-waste
production and management is the import of e-waste. TVs and monitors are the
main e-wastes imported from USA to Mexico because most of the processing
facilities are located in Mexico and India. Mexico has some facilities authorized to
handle the e-waste but their capacity is limited.
In conclusion, there is not enough information about e-waste in Mexico and
the existing one is dispersed. The e-waste management is not properly done
although there are authorized processing facilities in the country. This may be due
to the inexistence of a specialized law on e-waste management. Furthermore, if the
e-waste production trend continues as at the present and no further actions are
taken, it is probably that Mexico could become the next e-waste dump of the world.
Key words: electronic waste, e-waste, e-scrap, WEEE, technological waste,
Mexico
916
Introduction
Nowadays is difficult to conceive our daily activities without the use of any electric
and electronic devices. Electric and Electronic Equipment (EEE) such as
computers, telephones, TV sets, music players, refrigerators, and washing
machines, among others, are part of our life style.
The electric and electronic industry (EEI) has been established in Mexico for
more than three decades (Pro México, 2010). Mexico is well positioned as
manufacturer and exporter country of consumer electronic equipment (Pro México,
2013). In 2012, the production of the EEI in Mexico was US$55,703 million dollars
and is expected to growth at an average annual rate of 4% within the years 20122020. There are more than 700 EEIs (Figure 1) distributed at the northern,
western, and central regions of the Mexico (Román, 2007; Pro México, 2013).
Chihuahua,
98 facilities
Baja California,
213 facilities
Nuevo León,
98 facilities
Tamaulipas,
62 facilities
Querétaro,
25 facilities
Jalisco,
77 facilities
Ciudad de Méx. and
Edo. Méx., 62 facilities
FIGURE 1. Distribution of EEE facilities established in Mexico.
The total production of EEE in Mexico increased 24% from 2011 to 2012
(US$78,117 million dollars) (Ojeda-Benitez et al., 2013). The EEI is the main
exporting sector in Mexico. In 2008, Mexico was the 2 nd world exporter of TV sets,
the 3rd world exporter of freezers, and the 4th world mobile telephones exporter
(Pro México, 2010). By 2012, Mexico was the 1st worldwide exporter of flat-screen
TV sets, the 5th exporter of PCs, and the 10th mobile telephones exporter in the
entire world (Pro Mexico, 2013). The main export destination of the EEE
manufactured in Mexico was USA (84%) followed by Canada (2.8%), the
Netherlands (2.8%), and China (1.2%) (Pro Mexico, 2013).
However, the fast technological development accelerates the substitution of
the EEE before reaching its end-of-life. People in USA change their mobile
telephones every 2 years while in Europe people keep them for 1.5 year (UNEP,
2006). Personal computers (PCs) lifespan was reduced from 10 years in 1999
917
(Culver, 2005) to 3-4 years by 2005 (Sánchez et al., 2008; SEPA, 2011). As result,
high quantities of waste electric and electronic equipment (WEEE) are being
generated each year.
WEEE represents 1±5% of the global municipal waste production (OECD,
2008; Jain, 2008; Oteng-Ababio, 2012). In Europe, e-waste amounts for 8% of total
municipal waste with an annual growth rate of 3±5% (Bandyopadhyay, 2010).
WEEE is composed by more than 1000 different materials classified as hazardous
and non-hazardous by local legislations. Its high volume and the impacts
associated with its inappropriate handling is a matter of concern for solid waste
professionals (Musson et al., 2000).
E-waste is considered as one of the most complex and challenging waste
streams due to its physical, chemical, and heterogeneity composition (Yla-Meella
et al., 2004; Kumar et al., 2005). The e-waste production worldwide by 2012 was
49 million tons and it is expected to growth 33% by 2017 (StEP, 2013). According
ZLWKWKH6W(386$ZDVWKHZRUOG¶VKLJKHVWH-waste producer (10.3 million
tons in 2012) followed by China (7.3 million tons).
While in the European Union member countries the prevention of WEEE
generation is of first priority (EU, 2002), in some developing countries the legal
basis for its management is still under development. This is the case of some Latin
American countries. At the present, only Costa Rica, Colombia, Peru, Argentina
and Chile have baseline studies available in Latin America (Boeni et al., 2008).
In the case of Mexico, there is a baseline study about e-waste generation
(Román, 2007), as well as some regional and specific studies about e-waste
production in the country (Acevedo et al., 2008; Román, 2009; Meraz, 2010; CruzSotelo et al., 2013; Estrada-Ayub and Kahhat, 2014). However, the existing
information is scarce and dispersed (Chávez, 2008; Gasca, 2013). Thus, this paper
presents a review of the available national and international information about ewaste generation and management in Mexico. In this paper the terms technological
waste, e-waste, e-scrap, and WEEE are used as synonyms.
What is e-waste?
Up-to-QRZWKHUHLVQRWD³VWDQGDUG´GHILQLWLRQRIZKDWH-waste is. The Organization
for Economic Cooperation and Development (OECD) defines electronic waste, or
e-ZDVWH DV ³DQ\ DSSOLDQFH XVLQJ DQ HOHFWULF SRZHU VXSSO\ that has reached its
end-of-OLIH´2(&'
According to the EU WEEE Directive 2002/96/EC (EU, 2002) and the
'LUHFWLYH(8(8WKH³:DVWH(OHFWULFDQG(OHFWURQLF(TXLSPHQWRU
:((( PHDQV HOHFWULFDO RU HOHFWURQLF HTXLSPHQW ZKLFK LV ZDVWH« including
components, subassemblies, and consumables which are part of the product at the
WLPH RI GLVFDUGLQJ´ 7KH :((( GLUHFWLYH FODVVLILHV :((( LQ FDWHJRULHV
indicated on Table 1.
TABLE 1. Classification of WEEE in the European Union.
918
Category
Examples
Large household appliances
Refrigerators, washing machines, dish washing
machines,
microwaves,
air
conditioner
appliances, etc.
Small household appliances
Vacuum cleaners, irons, toasters,
machines, electric knives, watches, etc.
IT
and
equipment
coffee
telecommunication Mainframes, minicomputers, personal and laptop
computers (CPU, mouse, screen, and keyboard
included), printers, facsimile, telephones, pocket
and desk calculators, etc.
Consumer equipment
Radio and TV sets, video cameras, video
recorders, audio amplifiers, etc.
Lighting equipment
Luminaries for fluorescent lamps with exception
of luminaries in households, low pressure sodium
lamps, etc.
Electrical and electronic tools
Drills, sewing machines, tools for welding,
soldering or similar use, etc.
Toys, leisure
equipment
and
sports Electric trains or car racing sets, video games,
computers for biking, diving, running, etc.
Medical
devices
(with Radiotherapy equipment, cardiology, dialysis,
exception of all implanted and pulmonary ventilators, analyzers, freezers,
infected products)
laboratory equipment for in-vitro diagnosis, etc.
Monitoring
instruments
and
Automatic dispensers
control Smoke
detectors,
heating
regulators,
thermostats, measuring, weighing or adjusting
appliances for household or as laboratory
equipment, etc.
Automatic dispensers for hot drinks, automatic
dispensers for hot/cold bottles/cans, automatic
dispensers for money, etc.
The Solving the E-waste Problem initiative defines e-ZDVWHDV³DWHUPXVHGWR
cover items of all types of electrical and electronic equipment (EEE) and its parts
WKDW KDYH EHHQ GLVFDUGHG E\ WKH RZQHU DV ZDVWH ZLWKRXW WKH LQWHQWLRQ RI UHXVH´
(StEP, 2014).
In Mexico, e-waste is called technological waste and is considered as special
handling waste. The Ley General para la Prevención y Gestión Integral de los
Residuos (General Law for Prevention and Integral Waste Management) defines
WKH VSHFLDOKDQGOLQJ ZDVWHDV WKH ³ZDVWH generated in productive processes, that
does not have the characteristics to be considered as hazardous or urban solid
ZDVWH RU WKDW DUH SURGXFHG E\ KLJK TXDQWLW\ XUEDQ VROLG ZDVWH JHQHUDWRUV´
(LGPGIR, 2003).
919
The
Mexican
Official
Norm
(NOM)
NOM-161-SEMARNAT-2011
(SEMARNAT, 2013) states the criteria to classify the special handling waste. This
NOM indicates that this waste is subjected to a waste management plan to ensure
that it will be handled in an environmentally sound way.
Technological wastes that require a management plan are (SEMARNAT,
2013): personal computers and its peripherals (desktops and laptops), cellular
telephones, cathode ray tube monitors (including TVs), liquid crystal and plasma
screens (including TVs), sound and video portable players, wires of electronic
equipment, printers, photocopiers, and multifunctional devices.
The main representative e-wastes in Mexico are TV sets, telephones (fixed
and mobile), computers, and audio and video players because of its participation in
the national market (Román, 2007). The evolution of the quantity of houses having
at least one TV set, computer and/or telephone is shown in Figure 2 (INEGI, 2013).
Barba-Gutiérrez et al. (2008) used the Life Cycle Assessment methodology to
compare the environmental impacts of washing machines, refrigerators, TV sets
and PCs. These authors found that TV sets and PCs have the most pollutant
components in comparison with washing machines and refrigerators.
Due to this, most of the Mexican studies related with e-waste generation and
management are related with TV sets, mobile telephones, and computers
(Sánchez et al., 2008; Cruz-Sotelo, 2013).
Millions of houses with at least one device
35
30
Houses with PC
Houses with TV
Houses with telephone
Houses with fixed telephone
Houses with mobile telephone
Houses with fixed and mobile telephones
25
20
15
10
5
0
2004
2006
2008
2010
2012
2014
Year
FIGURE 2. Evolution of the selected electronic devices in Mexico.
Composition of WEEE
The composition of WEEE is as diverse as the materials contained in EEE offered
in the market. This is due to the diversity and the fast evolution on the EEE design
and the enforcement of the current laws related with WEEE and some hazardous
920
materials contained in such devices. In weight basis, e-waste is mainly composed
by ferrous and non-ferrous metals, plastics and glass as illustrated in Figure 3.
The composition of both personal computers and cell phones may vary
according to the model and the equipment manufacturer. Nonetheless, on weight
basis, computers are mainly composed of glass, plastics, iron, and aluminum
(Figure 3a and 3b). While mobile telephones (Figure 3c) and TV sets (Figure 3d)
are composed of plastics, glass, and copper.
a)
b)
Copper
7%
Nickel
Zinc Laton 1%
Lead
2% 1%
6%
Zinc Laton Nickel
3% 1%
1%
Copper
7%
Others*
0.3%
Glass
25%
Others*
1%
Glass
26%
Aluminium
15%
Aluminium
14%
Plastics
23%
Iron
21%
c)
Plastics
24%
Iron
22%
d)
Lead
Zinc
1%
1%
Copper
4%
Nickel
1%
Laton
10%
Others*
3%
Glass
20%
Aluminium
2%
Iron
10%
Aluminium
10%
Iron
10%
Copper
3%
Plastics
23%
Others
0.3%
Lead
5%
Glass
57%
Plastics
40%
* Such as gold, silver, mercury, chromium
FIGURE 3. Weight-based composition of a) Desktop PC, b) Laptop PC, c)
Mobile telephone and d) TV set. Data obtained from Román (2007b).
The presence of lead, mercury, arsenic, cadmium, hexavalent chromium, and
brominated flame retardants in e-waste makes it hazardous. Although its quantity
is lower in comparison with those who are able to recycle, its potential effects on
human health and the environment have serious consequences.
E-waste generation
921
Before 2007 there was not official available information about e-waste production
in Mexico (Hernández-Mendoza and Villa, 2008). In 2007 the Mexico National
Ecology Institute (INE) and the Mexican Environmental Protection and Natural
Resources Secretariat (SEMARNAT) published a national baseline study. It was
focused on the five electronic devises considered as representatives in Mexico, i.e.
TV sets, PCs (desk and portable), audio players, and fixed and mobile telephones.
The study was carried out using the available data about electronic
equipment manufactured in Mexico, for export and self-consumption, and data
about import of electronic equipment to be sold in the country. This report
considered an annual dumping rate of 50% and assumed a defined lifespan time
for each EEE (Román, 2007). The quantity of the illegal imports of electronic
equipment was also estimated to have an idea of the e-waste production due to
this activity.
The baseline study estimated an e-waste production in Mexico within 150,000
and 260,000 tons by 2006 (Román, 2007). Considering the e-was production
estimations in this baseline study, in Mexico, the e-waste production increased
167% in 8 years (from 1998 to 2006). The preliminary balance showed that 55% of
the e-wastes are due to the legal and illegal imports of electronic devices for its sell
and use in Mexico. The other 45% of the e-wastes are from national manufactured
electronic equipment, and from recycled and refurbished products.
The relative low contribution of illegal importations (1.5%) to e-waste
production in the Mexico was attributed to the difference of price within the
products legally imported (Román, 2007). Due to the North American Free Trade
Agreement signed in 1994 by Canada, USA, and México, the cost difference
between legally and illegally imported products may vary from 15% to 20%
(Román, 2007). Furthermore, the electronic devices illegally imported to the
country lacks of warranty which becomes a disadvantage for the buyer.
This study also found that 2.4 million of computers were dismissed by 2006,
while for mobile telephones it was observed a growth rate of 500% in six years
(from 2000 to 2006). This was attributed to its increasing availability and cost
decrease, as well as its lifespan reduction (Román, 2007).
Nonetheless, as this baseline study was considered DV DQ ³LQLWLDO´
approximation, the author recommends carrying out further studies to confirm
these results (Román, 2007). Thus, the INE and the SEMARNAT carried out some
regional studies to further characterize the e-waste production in the country. To
achieve this, the country was divided in regions according to their relative
importance on EEE production and consumption.
The update of the Mexico national baseline study in 2010 showed an e-waste
production of 307,000 tons/year (Gavilán-García et al., 2012) while a recent study
has estimated an e-waste production in Mexico of 1.138 million tons with an
average production per inhabitant of 9 kg/year by 2012 (StEP, 2013). Such
increase may be due to the methodologies applied in each study to quantify ewaste generation.
922
E-waste production in the northeast region. The diagnostics on e-waste
generation in the northeast region of the country covered the states of Coahuila,
Nuevo León, and Tamaulipas (Acevedo et al., 2008). The relevance of this region
lies on the increase of the production of the industries located in this region and the
growth on electronic equipment consumption. Also, its proximity to the USA makes
possible the import of EEE and e-waste in a non-controlled way.
TABLE 2. Lifespan of some electronic devices in Monterrey, Nuevo León.
Maximum time of
Lifespan
Average
use before
considered by
Electronic device
lifespan
dumping
Román (2007)
(Years)
(Years)
(Years)
TV sets
5.9 ± 0.4
18
10
PCs
4.2 ± 1.3
12
5
Music players
3.4 ± 0.5
16
6
Fixed telephones
5.9 ± 0.6
20
6
Mobile telephones
1.6 ± 0.2
15
3
An analysis of the e-waste generation in the urban area of Monterrey, located
in the state of Nuevo León, allowed to determine the lifespan of the electronic
devices in this area (Table 2). In the case of the lifespan of TV sets, the
assumption taken by Román (2007) is almost twice the obtained in the study
carried out in Monterrey. In contrast, lifespan of PCs and fixed telephones are
almost the same as the ones considered in the national baseline study.
This finding suggests that the national e-waste production estimated by
Román (2007) may be underestimated. Moreover, this result advises the
necessary to identify the lifespan of the selected e-wastes, in the following regional
studies, to improve the quantification of e-waste production.
14000
TV sets
PCs
Sound players
Fixed telephones
Mobile telephones
12000
Waste (ton)
10000
8000
6000
4000
2000
0
Nuevo León
Coahuila
Tamaulipas
923
FIGURE 4. Distribution of e-waste in the northern region of México.
The results obtained from this analysis were used to estimate the e-waste
generation in the three states covered by the northeast region of the country.
Figure 4 shows the distribution of e-waste by the devices studied in the northeast
region. The e-waste generation in the state of Nuevo León was 20,835 tons,
Coahuila produced 12,665 tons and Tamaulipas generated 15,749 tons (Acevedo
et al., 2008). Considering the quantity of PCs dumped until 2007, the quantity of
lead and plastics disposed of in landfills and open dumps was 643 and 2,254 tons,
respectively (Acevedo et al., 2008).
Furthermore, there was 91,753 tons of e-waste that was not been disposed
yet in the state of Nuevo León since they were stored or near to reach its end-oflife (Acevedo et al., 2008). This also occurs in the state of Coahuila (55,774 tons)
and Tamaulipas (69,359 tons).
E-waste production in the northern border region. This zone is important due to
the relationship of Mexico with the USA. There is a high formal and informal
commercial activity in this region due to the relative low prices of the products
imported from USA. This study paid special attention on the cities of Tijuana,
located in the state of Baja California Norte, and Ciudad Juárez, located in the
state of Chihuahua. These cities have a big economic importance due to the
presence of the manufacturing factories including those that are related with
manufacturing of electronic equipment.
The average lifespan of the electronic devices in this region (Table 3) is within
the range of the lifespan observed in the urban zone of Monterrey. The potential ewaste production in the north border region is 31,777 tons (Román, 2009). Such
quantity is 2 - 3 times higher than the average e-waste potential production of the
country (9,702 - 15,882 tons/year) calculated by Román (2007).
TABLE 3. Lifespan of some electronic devices in Tijuana and Ciudad Juárez.
Average lifespan
Lifespan considered
(Years)
Electronic device
by Román (2007)
(Years)
Tijuana
Ciudad Juárez
TV sets
4.6
6.6
10
PCs
4.0
3.4
5
Sound players
5.0
5.9
6
Fixed telephones
4.8
6.1
6
Mobile telephones
2.8
2.4
3
This contrast was attributed to the informal imports of the electronic devices
from USA and its commerce in Mexico (Román, 2009). This contrast with the
924
balance of e-waste management obtained in the national baseline study and
highlights the importance of control the illegal imports of e-waste to Mexico.
E-waste production in the central region of Mexico. The next regional study
was focused on the central zone of Mexico specifically México (Mexico City) and
the urban region of the Estado de México. The high population density and its
intensive economic activity including the electronic industry make this region of
special interest for environmental authorities. Taking into account the results
obtained from the northern border region study it was estimated that the central
region of the country have a potential e-waste generation of 68,282 ± 94,203 tons
(Meraz, 2010).
7KLV TXDQWLW\ LV WZLFH WKH RQH UHSRUWHG LQ WKH FRXQWU\¶V EDVHOLQH VWXG\ ZKLFK
implies a growth rate of this waste stream of around 25% per year. Nonetheless,
this rate has to be considered carefully since there are some differences in the
assumptions taken in the baseline study, e.g. the average lifespan of the selected
electronic devices.
E-waste production in the western region. The quantification of e-waste
produced in the state of Jalisco, located in the western region, is a special case for
Mexico. Most of the electronic industries related with information technologies
industry, the manufacturing of PCs and its parts are located in the urban area of
*XDGDODMDUDLQWKHVWDWHRI-DOLVFR7KLVUHJLRQKDVEHHQFDOOHG³WKH6LOLFRQ9DOOH\
RI 0H[LFR DQG/DWLQ $PHULFD´6iQFKH] HWDO 0ROLQD PDLQO\ GXH WR
the intensive activity of such industries.
However, there is no official study about e-waste production in this region of
the country. Only scarce and dispersed information about e-waste generation in
Jalisco was found during this review. In 2010, local authorities of Guadalajara
together with other associations collected 51 tons of e-waste during a collection
campaign. By 2011, eight municipalities joined to this campaign and collected
around 100 tons.
Peña et al. (2013) reported that during a collection campaign it was possible
to collect 110 tons of e-waste. In this campaign participated 46 municipalities of the
state of Jalisco and covered 56% of the total state population. In 2014, with the
participation of 49 municipalities, it was collected 94 tons of e-waste in a campaign
covering 65% of the population of the state (Proyecto Ecovia, 2014).
These data provide an idea about e-waste production in this state of the
country and highlights the necessity to carry out an official study about e-waste
production and management in this region.
E-waste production in the remaining regions of Mexico. Much effort has been
done by the authorities and researchers to provide information about e-waste
production in Mexico. Nonetheless, up-to-now there is scarce information about ewaste production in the southern part of the country. Thus, it is necessary to carry
out studies on e-waste production in this region to have a better idea about the
production and management of e-waste in the Mexico.
925
E-waste management
Management of e-waste is a complex activity since it is affected by many factors
such as the available technology, cultural aspects, and environmental factors
(Estrada-Ayub and Kahhat, 2014). E-waste disposal methods include landfill and
open dumps, incineration, recycling, and storage.
Results obtained from the baseline study on e-waste generation in Mexico
suggested that the management of this waste stream was as follows: 90% was
stored, 5% was disposed of in landfills or open dumps, 4% was recycled, and 1%
was exported to the country of origin to be disposed/treated (Román, 2007).
Data obtained from the diagnostics of the northeast region of Mexico show
that 30% of the TV sets reaching their end-of-life were donated to friends, family
members or schools to extend their useful life time, 27% were sold for second
hand use, and 29% were disposed of in landfills or open dumps (Figure 4).
Most of the PCs out of use were sold (60%) for second hand use or for using
its parts as fixing parts. In the case of mobile telephones, 43% of them were lost.
So, it is probably that they were sold or used by people who found it. For all of the
studied devices the recycling rate was within 2% - 11% while the dumping rate
varied from 8% to 61%. The recycling rate of the e-waste generated in the central
region of Monterrey is 4%, which is in agreement with that reported by Román
(2007).
70
Donated
Dumped
Sold
Lost
Recycling
Sound players
Fixed
telephones
Unknown
60
Percentage (%)
50
40
30
20
10
0
TV sets
PCs
Mobile
telephones
FIGURE 4. E-waste management in the urban region of Monterrey.
926
b)
a)
Sold
11%
Stored
13%
Donated
34%
Stored
6%
Sold
12%
Dumped
35%
Dumped
42%
Donated
47%
FIGURE 5. E-waste management in a) Tijuana and b) Ciudad Juárez.
The diagnostics on e-waste production in the northern region of Mexico found
that 45% of the e-waste produced in Tijuana is donated to friends or family
members as well as selling them for second hand use, while 42% are dumped in
landfills or open dumps (Figure 5A). In Ciudad Juárez, most of the e-waste is
reused (59%) while a lower proportion is dumped (Figure 5B). Unfortunately, this
study does not proportionate data about recycling quantities despite of the
existence of infrastructure for this activity.
Landfill and open dumps leachates can transport toxic materials into ground
water and soil. According to a study conducted by Global Futures Foundation
(2001) e-waste are responsible for 70% of the toxic substances emitted in landfills.
Although most of the e-waste is disposed of in landfills (Darby and Obara, 2005;
Barba-Gutiérrez et al., 2008), no significant amounts of e-waste have been
observed in landfill facilities in Mexico (Román, 2007, 2009). This may be due to
the presence of informal recycling sector whose activity is stimulated by the
recovery of valuable materials (e.g. gold, copper, parts that can sell for reuse)
contained in e-waste.
The informal recycling sector is integrated by poor and marginalized social
people who resort to scavenging and waste picking for income and survival (Wilson
et al., 2006). Recyclable materials and used products taken out of the municipal
waste stream are sold on street markets or in other places. Presence of informal eZDVWH FROOHFWRUV RU VFDYHQJHUV DOVR NQRZQ DV ³SHSHQDGRUHV´ LQ 0H[LFR LV
recognized (Estrada-Ayub and Kahhat, 2014). Nevertheless, there is no
information about its importance and how they interact with formal stakeholders.
Some ways adopted worldwide to recover valuable materials by informal
recyclers are crushing of recovered devices and use of open burnings (EstradaAyub and Kahhat, 2014). Darby and Obara (2005) indicated that WEEE is
incinerated without any previous treatment. Incineration of e-waste can emit heavy
metals (e.g., lead, mercury, and cadmium), dioxins and furans to the air that are
toxic for both the environment and human health.
927
There are few cathode ray tube (CRT) processing facilities in the world. One
of these facilities is located in the city of Tijuana, Mexico. Due to the proximity of
Tijuana to the USA, more than 30% of used TV sets and monitors are exported to
this CRT processing facility (Duan et al., 2013). Also, PCs and mobile telephones,
which are out of use, are exported from USA to Mexico (Duan et al., 2013) to be
sold as second hand devices.
It has been observed that countries that allow importation of e-waste and
lacks of specialized norms and facilities for e-waste recycling have serious
problems with this waste stream. Cities that import e-waste tend to become
dumping sites because of the insufficient e-waste treatment capacity of the
installed facilities and the fast accumulation of e-waste imported (Sánchez et al.,
2008). This is the case of the city of Guiyu, China (Yang et al., 2008), South Africa
(Lombard and Widmer, 2005) and Senegal (Rochat et al., 2008). In India, the
hidden flows of e-waste to the authorized treatment facilities represented over 99%
(143,000 tonnes) of the WEEE to be treated (Greenpeace, 2008). Thus, if no legal
actions are taken to control the imports of e-waste to the country, it is possible that
Mexico could become the dumping site of the world (Sánchez et al., 2008; Varela,
2014).
Estrada-Ayub and Kahhat (2014) carried out a study to define the cultural,
VRFLDO DQG HFRQRPLFDO GHFLVLRQ IDFWRUV DIIHFWLQJ WKH SHRSOHV¶ FKRLFH WR GXPS RU
trade their used PCs in the northern region of Mexico. These authors divided the
stakeholders in three levels: micro, meso, and macro levels. Micro-level are
represented by the actions taken by the individuals, meso-level symbolizes the flow
of information within the organizations, and macro-level covers the characteristics
of the country.
According to these authors, technological changes are the most important
decision factor at macro-level while legislation was the most important factor at
meso-level. At micro-level, both perceived e-waste value and geographical location
DUHWKHPRVWLPSRUWDQWGHFLVLRQIDFWRUVWKDWLQIOXHQFHWKHSHRSOHV¶FKRLFHWRGXPS
or trade their computers.
Knowledge is another important factor with a high influence at different levels
(Estrada-Ayub and Kahhat, 2014). The diagnostics on e-waste in the northern
border region of Mexico investigated the knowledge of the people about hazards
related with the e-waste. A total of 384 personal interviews were done in each city.
It was found that 57% of the interviewed people in Tijuana and 72% of the people
in Ciudad Juárez did not consider that e-waste have environmental and health
hazards (Román, 2009).
In the same way, the diagnostics on e-waste production in the central zone of
Mexico investigated the knowledge about e-waste hazards on human health. From
the 1,000 interviewed people only 43% known that the e-waste can produce health
hazards, mainly due to the presence of hazardous waste in the EEE (Meraz,
2010). This agrees with a previous study carried out in 2006 in which a similar
tendency was observed (Hernández-Mendoza and Villa, 2008).
Legislation is the most important decision factor at the meso-level but its
impact depends on the existence of the adequate structure to support it (EstradaAyub and Kahhat, 2014). There are around 50 authorized facilities for reception
and recycling of e-waste in Mexico (Semarnat, 2010) whose distribution is shown
928
Figure 6. Although most of the computing related manufacturing industry installed
in Mexico is located in the state of Jalisco, there are only two authorized recycling
facilities in such region. Moreover, there are some states, such as Chiapas,
Campeche, Colima, and Nayarit, in which there are no authorized facilities for ewaste recycling.
The different regional diagnostics performed in Mexico tried to get information
about recycling capacities and technology used by the authorized industries in
Mexico. However, the authors could not obtain any valuable data mainly because
of the negative of chairmen to provide information considered as confidential. Only
few recycling industries indicated that they can recycle up to 95% of the e-waste
collected.
This agrees with LaDou and Lovegrove (2008) who indicated that it is
possible to recover 95% of the useful materials from a computer and 45% of the
materials from a CRT monitors. Mechanical processes for e-waste recycling (e.g.,
crushing, screening, magnetic and electrostatic separation of metals) can recover
8% of lead, 24% of tin, and more than 50% of the copper (LaDou and Lovegrove,
2008).
Recovery of used printer
cartridges to be recycled
Baja California,
2 facilities
Chihuahua,
1 facility
Coahuila,
1 facility
Nuevo León,
5 facilities
San Luis Potosí,
2 facilities
Guanajuato,
1 facility
Aguascalientes,
1 facility
Querétaro,
4 facilities
Jalisco,
2 facilities
Veracruz,
2 facilities
Michoacán,
1 facility
Estado de México,
4 facilities
Ciudad de México,
12 facilities
Morelos,
2 facilities
Tabasco,
3 facilities
FIGURE 6. Distribution of the reception and recycling sites for e-waste in
Mexico.
Conclusions
929
The available studies about e-was production and management are mainly focused
in the northern border, northwest and central regions of Mexico. The existing
information is dispersed and needs to be updated. Southern and occidental regions
of Mexico do not have official studies about e-waste production and management.
Only 2±11% of e-waste produced in the different regions of Mexico is recycled.
Although there are authorized facilities for e-waste treatment, it is neither known
the technologies nor the capacities to treat theme. E-waste importation to Mexico
may have a higher importance than the assumed in the baseline study. It has been
observed that both legal basis and lack of knowledge about the hazards associated
with the inadequate e-waste management have an important role in waste handling
by the population. If e-waste production and importation trend continues as at the
present and no further actions are taken, it is high probably that Mexico could
become the next e-waste dump of the world.
930
References
Acevedo J, Rivas ER, Carrillo O (2008). Diagnóstico regional de la generación de residuos
electrónicos al final de su vida útil en la Región Noreste de México. Centro de Calidad
Ambiental ± ITESM campus Monterrey, Instituto Nacional de Ecología y Secretaria del Medio
Ambiente y Recursos Naturales, 133 p.
Bandyopadhyay A (2010). Electronics waste management: Indian practices and guidelines.
International Journal of Energy and Environment;1:793-804.
Barba-Gutiérrez Y, Adenso-Díaz B, Hopp M (2008). An analysis of some environmental
consequences of European electrical and electronic waste regulation. Resources,
Conservation and Recycling; 52:481-495.
Boeni H, Silva U, Ott D (2008). E-waste recycling in Latin America: Overview, challenges and
potential. In: Mishra B, Ludwing C, Das S, Editors. Proceedings of the 2008 Global
Symposium on Recycling, Waste Treatment and Clean Technology, REWAS, October 12-15,
2008, Cancun, Mexico.
Chávez CR (2008). Basura, desechos, residuos. Simposio: Desarrollo de estadísticas del medio
ambiente: fuentes, alcances y usos. COLMEX-SEMARNAT-INEGI. October 20-21, 2008.
Mexico City, Mexico.
Cruz-Sotelo SE, Ojeda-Benítez S, Bovea MD, Santillán-Soto N, Favela-Ávila H, Aguilar WE (2013).
Hábitos y prácticas de consumo de teléfonos celulares en México y España. Revista
Internacional de Contaminación Ambiental; 29:33-41.
Culver J (2005). The life cycle of a CPU. Available on: http://www.cpushack.net/life-cycle-ofcpu.html; accessed on June 15, 2014.
Darby L, Obara L (2005). Household recycling behavior and attitude towards the disposal of small
electrical and electronic equipment. Resources, Conservation and Recycling; 44:17-35.
Duan H, Miller R, Gregory J, Kirchain R (2013). Quantitative characterization of domestic and
transboundary flows of used electronics. Analysis of generation, collection, and export in the
United States. Massachusetts Institute of Technology, 122p.
Estrada-Ayub J, Kahhat R (2014). Decision factors for e-waste in Northern Mexico: To waste or
trade. Resources Conservation and Recycling; 86:93-106.
EU (2002). Directive 2002/96/EC of the European Parliament and of the Council of 27 January
2003 on waste electrical and electronic equipment (WEEE) ± Joint declaration of the
European Parliament, the Council and the Commission relation to Article 9. Official Journal of
the European Union, p. 24-39.
EU (2012). Directive 2012/19/EU of the European Parliament and of the Council of 4 July 2012 on
waste electrical and electronic equipment (WEEE). Official Journal of the European Union, p.
38-71.
Gasca S (2013). El manejo de los residuos de manejo especial. Foro: La gestión de sitios
contaminados y residuos en México a 10 años de la publicación de la LGPGIR. Septiembre
4, 2013. Ciudad Universitaria, UNAM, Mexico City. Mexico.
Gavilán-García A, Alcántara-Concepción V, Gavilán-García I (2012). Análisis de ciclo de vida de las
alternativas para el manejo de residuos electrónicos en México. Proceedings of the 5th
Congress of the Mesoamerican Association of Ecotoxicology and Environmental Chemistry,
July 25-28, Aguascalientes, Mexico.
Global Futures Foundation (2001). Computers, e-waste, and product stewardship: Is California
ready for the Challenge? Report for the U. S. Environmental Protection Agency Region IX,
San Francisco California, USA.
Greenpeace (2008). Toxic Tech: Not in our Backyard. Uncovering the hidden flows of e-waste. The
Netherlands, 11p.
Hernández-Mendoza CE, Villa-Aguirre XC (2008). Los residuos electrónicos: un desafío para
México. Proceedings of the XVI Mexican National Congress on Sanitary Engineering and
Environmental Sciences, April 21-26, 2008, Mexico City, Mexico.
INEGI (2013). Módulo sobre Disponibilidad y Uso de Tecnologías de la Información en los Hogares.
Noviembre, 2013. México.
931
Jain A (2008). Global e-waste growth.Chapter one. In: Johri R (Ed.). E-waste: implications,
regulations, and management in India and current global best practices. Terri Press, India.
LaDou J, Lovegrove S (2008). Export of electronics equipment waste. International Journal of
Occupational and environmental Health; 14:1-10.
LGPGIR (2003). Ley General para la Prevención y Gestión Integral de los Residuos. Diario Oficial
de la Federación, 8 de octubre de 2003.
Lombard R, Widmer R (2005). E-waste assessment in South Africa, A case study of the Gauteng
province. EMPA ± Swiss Federal Laboratories for material Testing and Research,
Switzerland.
Meraz RL (2010). Diagnóstico de la generación de residuos electrónicos en la Zona Metropolitana
del Valle de México. Instituto Politécnico Nacional ± Centro Interdisciplinario de
Investigaciones y Estudios sobre Medio Ambiente y Desarrollo, Instituto Nacional de
Ecología y Secretaría del Medio Ambiente y Recursos Naturales, 81 p.
Molina C (2008). Jalisco, el Silicon Valley de Latinoamérica. BNamericas, Bussiness insight in Latin
America. 16 de Junio.
Musson SE, Jang Y-C, Townsend TG, Chung I-H (2000). Characterization of lead leachability from
cathode ray tubes using the toxicity characterization leaching procedure. Environmental
Science and Technology; 34:4376-81.
OECD (2001). Extended producer responsibility: a guidance manual for governments. OECD
Publishing. Paris, France.
OECD (2008). OECD Environmental Outlook to 2030.
Ojeda-Benitez S, Cruz-Sotelo SE, Velázquez L, Santillán-Soto N, Quintero M, García OR, Markus
W (2013). Electrical and electronic waste in Northwest Mexico. Journal of Environmental
Protection; 4:405-410.
Oteng-Ababio M (2012). Electronic Waste Management in Ghana ± Issues and Practices. In:
Curkovic S (Ed.). Sustainable Development ± Authoritative and Leading Edge Content for
Environmental Management. InTech press,
Pro México (2010). La industria eléctrico electrónica. México, 14p.
Pro México (2012). ¿Por qué manufacturar en México? México, 9p.
Pro México (2013). Industria electrónica 2013. México, 9p.
Proyecto Ecovia (2014). Resultados Electroacopio Jalisco 2014. Disponible en:
http://www.proyectoecovia.com/actividades_desglose.php?id=38; recuperado el 5 de Junio
de 2014.
Rochat D, Gassama CID, Kane C, Wone S (2008). Senegal: e-waste country assessment (2008).
Proceedings of the 19th Waste Management Conference of the IWMSA, October 6-10,
Durban, South Africa.
Román GJ (2007). Diagnóstico sobre la generación de residuos electrónicos en México. Instituto
Politécnico Nacional ± Centro Interdisciplinario de Investigaciones y Estudios sobre Medio
Ambiente y Desarrollo, Instituto Nacional de Ecología y Secretaría del Medio Ambiente y
Recursos Naturales, 122 p.
Román GJ (2007b). Desarrollo de un programa modelo para el manejo de residuos electrónicos en
México. Instituto Politécnico Nacional ± Centro Interdisciplinario de Investigaciones y
Estudios sobre Medio Ambiente y Desarrollo, Instituto Nacional de Ecología y Secretaría del
Medio Ambiente y Recursos Naturales, 61 p.
Román GJ (2009). Diagnóstico regional de residuos electrónicos en dos ciudades de la frontera
norte de México: Tijuana y Ciudad Juárez. Instituto Politécnico Nacional ± Centro
Interdisciplinario de Investigaciones y Estudios sobre Medio Ambiente y Desarrollo, Instituto
Nacional de Ecología y Secretaría del Medio Ambiente y Recursos Naturales, 111 p.
Sánchez M, Bonales J, Espinoza R (2008). Contaminación del medio ambiente en la región oriente
del estado de Michoacán por desechos electrónicos de equipo de cómputo obsoleto. Mundo
siglo XXI; 13:61-71.
SEMARNAT (2010). Directorio de centros de acopio de materiales provenientes de residuos en
México. Secretaría del Medio Ambiente y Recursos Naturales. México D. F., México.
SEMARNAT (2013). Norma Oficial Mexicana NOM-161-SEMARNAT-2011, Que establece los
criterios para clasificar a los Residuos de Manejo Especial y determinar cuáles están sujetos
a Planes de Manejo; el listado de los mismos, el procedimiento para la inclusión o exclusión
932
a dicho listado; así como los elementos y procedimientos para la formulación de los planes
de manejo. Diario Oficial de la Federación, 1 de Febrero de 2013.
StEP (2013). StEP E-Waste World Map. United Nations University/ Solving the E-waste Problem
initiative. Available on http://www.step-initiative.org/index.php/WorldMap.html; accessed on
June 11, 2014.
StEP (2014). One global definition of e-waste. Solving the E-waste Problem (StEP) White Paper.
United Nations University/Step Initiative 2014, 13 p.
SEPA (2011). Recycling and disposal of electronic waste. Health hazards and environmental
impacts. Swedish Environmental Protection Agency, Report 6417, Stockholm, Sweden.
UNEP (2006). Vital waste graphics 2. Volume 2. UNEP/GRID Arendal, The Basel Convention
Secretariat. Geneva, 44 p.
Varela, R (2014). Profepa decomisa residuos peligrosos. El Financiero. Disponible en
http://www.elfinanciero.com.mx/opinion/profepa-decomisa-residuos-peligrosos.html;
recuperado el 16 de abril de 2014.
Yang J, Lu B, Xu C (2008). WEEE flow and mitigation measures in China. Waste Management;
28:1589-1597.
Yla-Meella Y, Pongracz E, Keiski RL (2004). Recovery of waste electrical and electronic equipment
(WEEE) in Finland. In: Pongracz E, editor. Proceedings of the waste minimization and
resource use optimization conference, June 10, 2004, Oulu, Finland.
Notation
CRT
e-waste
EEE
EEI
EU
INE
NOM
OECD
PCs
StEP
SEMARNAT
USA
WEEE
cathode ray tube
electronic waste
electric and electronic equipment
electric and electronic industry
European Union
Mexico National Ecology Institute
Mexican Official Norm
Organization for Economic Cooperation and Development
personal computers
Solving the E-waste Problem initiative
Mexican Environmental Protection and Natural Resources
Secretary
United States of America
waste electric and electronic equipment
933
CHAPTER 8.2. BIODEGRADATION OF PLASTIC FILMS BY ANAEROBIC
DIGESTION
Santiago-Juárez Gabriela Araceli (1); Vázquez-Morillas Alethia (2);
Velasco-Pérez Maribel (3); Beltrán-Villavicencio Margarita (4)
(1) Maestría en Ciencias e Ingeniería Ambientales. Universidad Autónoma Metropolitana- Unidad
Azcapotzalco. Avenida San Pablo # 180, Colonia Reynosa Tamaulipas. C.P. 02200, México
D.F. email: [email protected]
(2) Departamento de Energía. Universidad Autónoma Metropolitana- Unidad Azcapotzalco. Ibidem.
email:[email protected]
(3) Ibidem. email: [email protected]
(4) Ibidem. email: [email protected]
ABSTRACT
Plastic films are widely used for their versatility and low cost. However, their high
rate of consumption, discard and stability in the environment, transform them in a
waste stream of complicated disposal and expensive treatment. In México City, the
local government has promoted the use of compostable and oxodegradable
plastics as a way to lessen the environmental impact of plastics. At present, it is
evaluating alternatives to allow the treatment of plastic films along with organic
waste. For that reason, this study assessed the degradability of oxodegradable
HDPE (oxidized and non- oxidized), conventional HDPE and compostable film
Ecovio® in anaerobic digestion conditions, following the guidelines established by
the ASTM D5511-11. The process was carried out in hermetic sealed reactors with
a volume of 1000 cm3. Each device contained cow manure as anaerobic inoculum
(187.5 cm3), dry grass as substrate, (150 cm3), water (562.5 cm3) and 30 plastic
strips (1x15 cm). The experiment was performed with a control device (without
plastic) and in triplicate for each plastic. Oxygen in devices was removed with
nitrogen at the beginning of the experiment. The anaerobic digestion process
lasted 40 days and was maintained at 52 ± 2 °C. Decrease of elongation at break
was measured at the beginning and the end of anaerobic digestion; it was used as
indicator of plastic films degradation. At the end of the experiment, elongation at
EUHDN GHFUHDVHG ı IRU (FRYLR ı IRU R[LGL]HG
R[RGHJUDGDEOH +'3( ı IRU XQR[LGL]HG R[RGHJUDGDEOH +'3(
38.8% IRU ZKLWH FRQYHQWLRQDO +'3( ı DQG IRU EHLJH FRQYHQWLRQDO
+'3(ı 7KHELRJDVSURGXFWLRQZDVDOVRXVHGDVLQGLFDWRURILQKLELWLRQRI
anaerobic digestion by plastic films. The greater amount of biogas was produced
by white conventional HDPE (5126.5 cm3), followed by beige conventional HDPE
(5054.5 cm3), unoxidized oxodegradable HDPE (4788 cm 3), control (3924.5 cm3),
Ecovio® (3850.67 cm3), oxodegradable oxidized HDPE (3593.33 cm 3). Thus, it
was concluded that the anaerobic digestion process produces some level of
degradation in the studied plastics, even if they have not been made to degrade in
absence of oxygen.
Key words: Biodigesters, Biogas, Biodegradable plastics, Municipal solid waste,
Plastic bags.
Introduction
934
Plastics are very versatile materials, given their characteristics as: low cost,
resistance to corrosion, thermal and electrical insulation, low weight and durability.
In Mexico in 2009, 20 million plastic bags made of polyethylene were used daily
(Alcántara, 2009; Sosa, 2002). In 2012, from a total of 42.1 million of tons of waste
generated in the country, 4.58 million of tons were plastic, this represent more than
10 % (INEGI, 2012; Conde, 2012; SEMARNAT-SNIARN, 2012).
The durability and stability of these materials, along with their accelerated rate
of consumption and waste has caused accumulation in terrestrial and aquatic
environments (Galbraith, 2012; Thompson et al., 2009; Sosa 2002). To control this
problema different cities have tried to control the increasing use of plastics.The use
of plastic bags has been banned in California in the United States, Toronto in
Canada, China, South Africa, the United Kingdom and some cities of India and
Australia (Ruíz, 2010; Mingquing, 2008).
In November 2010 the Legislative Assembly of Mexico City modified the Solid
Waste Act in order to reduce the environmental impacts of plastic. For example,
Article 6 Fraction XI explains the need to establish criteria, standards of
sustainable production and consumption of plastic products, ensuring a
degradation time shorter than the one of conventional polymers. In Article 6,
Fraction XII expressed the need to promote research and development of
technologies to prevent, control and reduce pollution by the use of plastic bags.
Articles 25 Fraction XI Bis and 26 Bis talk about the use of plastic bags that meet
the sustainability criteria and standards established in Article 6 Fraction XI (Ruíz,
2010).
One of the strategies employed to reduce the environmental impact of
plastics is the production of biodegradable plastics. Biodegradation is a natural
process in which organic chemicals degrade to simpler compounds, are
mineralized and return to the biological cycles of carbon, nitrogen and sulfur by the
action of living organisms and abiotic factors such as photo-degradation, oxidation
and hydrolysis (Amass et al., 1998; Chandra and Rugsti, 1998).
There are different kinds of biodegradable plastics, among the most
prominent may be mentioned the following (Scott et al., 2011; Scott, 2000; BPI,
2010):
x Inherently biodegradable plastics: they can be biodegraded in a variety of
environments, both in anaerobic and aerobic conditions. It is possible to
synthesize from fossil fuels (polycaprolactone) or renewable natural
resources (polyhydroxyalkanoates).
x Compostable plastics: they are degraded by hydrolytic process occurring
during the composting of organic solid waste. Examples include, polylactic
acid (synthesized from renewable sources) and commercial formulations as
Ecovio® and Ecoflex®, which are produced from petroleum or natural gas.
x Oxodegradable plastics: they are given that name because they degrade in
two stage, the first involving abiotic oxidation. These plastics are made from
the same resins than conventional plastics. Appearance, durability and
935
resistance are the same over its lifetime than that of conventional plastics.
The only difference is that the oxodegradable plastic contain about 1% of an
organic salt with transition metals (Co, Mn, Mg, Fe, or Zn), which facilitates
the oxidation and breakage of long chain polymers. As a result, they can
reach smaller sizes and being able to be consumed by the microorganisms
present in some environments.
In México City the degradable plastics that dominate the market are
oxodegradables and compostable. These materials are designed to degrade under
aerobic conditions, so ideally should be processed with the organic fraction of
PXQLFLSDO VROLG ZDVWH FRPSRVWLQJ SODQW ³%RUGR 3RQLHQWH´ 'XH WR FDSDFLW\
constraints and the time that the composting process takes, the local government
is evaluating the possibility of implementing other alternatives for the treatment of
organic material such as waste anaerobic digestion. In this context, this project
studies the behavior of the compostable and oxo-degradable plastic films under
anaerobic digestion conditions.
Five types of plastic films were evaluated: high density polyethylene with prodegradant additive (oxodegradable HDPE), high density polyethylene oxidized with
pro-degradant additive (oxidized oxodegradable HDPE), a compostable polylactic
acid formulation (Ecovio®) and conventional high density polyethylene
(conventional HDPE) similar in composition and only different in color, beige and
white. An experimental set with no plastic was used as a control.
Strips of each plastic film (1x15 cm) of interest were cut using a scalpel blade
to obtain clean cuts, as shown in Figure 1. A portion of samples of oxodegradable
HDPE was oxidized in accelerated weathering chamber for 2 days at 60 °C (see
Figure 2). Test were conducted to determine the percentage of elongation at break
of the plastic film strips, before and after the anaerobic digestion process, using a
XQLYHUVDO WHVWLQJ PDFKLQH ³/OR\G /) SOXV´ VHH )LJXUH 7KLV ZDV XVHG DV D
measure of degradation.
The substrate used for the anaerobic digestion process was grass. It was
introduced into the oven and dried for 2 days at 60 °C and subsequently triturated
ZLWKDPDQXDOJULQGHUWRREWDLQSDUWLFOHVVL]HRIPP)UHVKFRZGXQJREWDLQHG
a cowshed, was used as anaerobic inoculum. Dry grass (substrate) and inoculum
(cow dung) were characterized in triplicate before anaerobic digestion with analysis
of: organic matter, total nitrogen content, C/N ratio and pH of the mixture of water,
grass and manure. After the anaerobic digestion, a similar characterization was
performed to the digested phase.
936
FIGURE 1. Cut of strips of plastic films.
FIGURE 2. Oxidation of plastic films in the accelerated weathering chamber.
937
FIGURE 3. Testing machine Lloyd LF plus.
Semitransparent polypropylene containers were used as biodigesters for the
anaerobic digestion process. They had a wall thickness of about 8 mm, 1 liter
capacity, with screw cap and gasket that provides sealing. Cow dung was mixed
with water in ratio 1:3 (1 kg of manure: 3 liters of water). The mixture was stirred
until a homogeneous texture was obtained. With a 1000 ml beaker, 750 ml of the
mixture of manure and water were measured and poured into each biodigestor.
Following, 35 g of dry grass as primary substrate (Figure 4) and the
corresponding 30 strips of plastic were added to each digester (Figure 5). To
completely eliminate the presence of oxygen, nitrogen was blown through the
headspace of the biodigesters for 5 minutes and at the end of the purge, both the
outlet valve of gas and purge were closed.
The experiment was performed in triplicate for each plastic film (18 digesters
were operated simultaneously). Temperature was kept constant at 52 ± 2 °C in
temperature-controlled bath, and the degradation of the plastic was evaluated after
40 days of anaerobic digestion (Figure 6).
The biogas produced in the anaerobic digestion process was captured using
the collection system shown in ASTM D5511-11, and Tedlar ® bags for storing
biogas were used. Each bag was connected to a bottle that contained an acid
solution, the bag was pressed so that the content would be directed toward the
bottle. The displaced liquid from the bottle (equivalent to the volume of biogas in
the Tedlar® bag). It was collected in a 1000 ml graduated tube to quantify the
volume of biogas.
938
FIGURE 4. Dry grass added to the
biodigesters
FIGURE 5. Plastic films added to the
digesters
FIGURE 6. Anaerobic digestion experiment.
Degradation of plastic films. The plastic film with the lowest elongation at break
(i.e., higher degradation) after the anaerobic digestion process was Ecovio®,
followed by oxidized oxodegradable HDPE, white conventional HDPE, unoxidized
oxodegradable HDPE and beige conventional HDPE. Elongation at break
GHFUHDVHG ı IRU (FRYLR ı IRU R[LGL]HG R[RGHJUDGDEOH
+'3( ı IRU XQR[LGL]HG R[RGHJUDGDEOH +'3( IRU ZKLWH
FRQYHQWLRQDO +'3( ı DQG IRU EHLJH FRQYHQWLRQDO +'3( ı (Figure 7). While Ecovio® showed a high rate of biodegradation, the rest of the
939
plastics finished the process with values of elongation at break that show they are
still usable.
FIGURE 7. Elongation at break of plastic films before and after the anaerobic
digestion process.
Digestion process. The characterization of the substrates before and after the
digestion process is shown in table 1. The percentage of organic matter in all
treatments after anaerobic digestion was higher than the initial value and there
were no significant difference in the values of organic matter of different plastics
after anaerobic digestion. In the other hand, the initial values of the percentage of
total nitrogen to the substrate (dry grass) and inoculum (cow dung) are constraints
for specific type of waste (FAO, 2011; Fernández et al., 2009; Sosa, 2005). There
were no significant difference between the values of total nitrogen of the various
plastic studied.
It has been reported that suitable values of C/N for an appropriate anaerobic
digestion process, are between 25 and 35 (Buenrostro et al., 2000; Veyna, 2007).
Both initial and final results of C/N are suitable, which means that the process was
conducted with a proper balance of nutrients. There were no significant difference
between the C/N ratio of the various end digestate.
The pH at the start of the process was within the range of neutrality. The pH values
at the end of the process increased but remained within neutrality and no
acidification occurred, contrary to what has been reported in literature (Ocaña,
2011; López and López, 2009; Veyna, 2007; Martí, 2006; Monet, 2003; Verma,
2002).
940
TABLE 1.
digestion.
Before
anaerobic
digestion
After
anaerobic
digestion
Characterization of substrates before and after anaerobic
Substrate (dry
grass)
Inoculum (cow
dung)
Control
Ecovio®
Oxidized
oxodegradable
HDPE
Unoxidized
oxodegradable
Beige
conventional
HDPE
White
conventional
HDPE
Organic
matter
(%)
Total
nitrogen
content
(%)
Carbon/Nitrogen
87.26
1.74
29.09
pH
7.4
53.67
0.96
32.42
72.56
72.74
1.83
1.80
23.0
23.44
7.94
7.97
73.55
1.74
24.52
7.95
72.83
1.60
26.40
7.95
72.52
1.77
23.76
8.09
73.07
1.65
25.68
7.92
Biogas produced. All plastic film produced more biogas than the control, except
the oxidized oxodegradable HDPE, which produced 331.17 cm 3 less than the
control. Another treatment that had low biogas production was Ecovio®, which
produced only 73.83 cm3 more than the control.
The treatments that produced more biogas were those with conventional
HDPE plastic films. Beige conventional HDPE produced 1202 cm3 of biogas more
than the control and white conventional HDPE produced 1130 cm 3 more than the
control. These plastics showed more reproducible results as well. The total amount
of biogas produced by each treatment is shown in Figure 8.
941
FIGURE 8. Total biogas produced.
Conclusion
Initial parameters of organic matter content, nitrogen and pH were suitable for
starting the process of anaerobic digestion and the values obtained at the end of
the process reflect the stability of the process. The final values of organic matter,
nitrogen and pH indicate that the final digestate is suitable for use as fertilizer in
soil.
All treatments had biogas generation, so that it can be concluded that the
presence of plastic film does not inhibit the process of anaerobic digestion. The
treatment produced less biogas was oxidized oxo, the degradation process of this
plastic film could have released the pro oxidant additives and these have caused
sensitivity to microorganisms. However the anaerobic digestion process was not
inhibited.
All the plastics showed some degradation in the process, with Ecovio®
showing more significant changes. Including these amounts of compostable plastic
films, oxodegradables and even conventional HDPE in anaerobic digestion
processes contributes to their degradation, and there should be further analyzed as
a fore waste treatment option.
Acknowledgements
7KLV UHVHDUFK LV SDUW RI WKH SURMHFW ³(YDOXDFLyQ GH OD ELRGHJUDGDELOLGDG \
ecotoxicidad de plasticos oxodegradables y biodegradables en condiciones
DHURLDV \ DQDHURELDV´ IXQGHG E\ WKH )RQGR PL[WR &21$&<7 ± Gobierno del
Distrito Federal, FM/CONACYT-GDV/05/2013.
References
942
Alcántara, L. (2009). Bolsas de plástico envenenan al país. El Universal.
Amass W., Amass A. & Tighe B. (1998). A review of biodegradable polymers uses, current
developments in the synthesis and characterization of biodegradable polyesters, blends of
biodegradable polymers and recent advances in biodegradation studies. Polymer
International 89-144.
BPI, B. P. I. (2010). Background on Biodegradable Aditives.
Buenrostro, O., Cram, S., Bernache, G. & Bocco, G. (2000). La digestión anaerobia como
alternativa de tratamiento a los residuos sólidos orgánicos generados en los mercados
municipales. Revista Internacional de Contaminación Ambiental, 16, 19-26.
Conde, M. P. O. (2012). Presente futuro de la industria del plástico en México. In: PEMEX, P. M.
(ed.).
Chandra, R. & Rustgi, R. (1998). Biodegradable Polymers. Progress in Polymer Science, 63.
Galbraith, K. (2012). Should Plastic Bags Be Banned?. New York Times.
INEGI, I. N. D. E. Y. G. (2012). México de un vistazo. anual. 2012 ed.
López, M. C. & López, S. O. (2009). Diseño, construcción y puesta en operación de un biodigestor
anaerobio continuo para el laboratorio de ingeniería química de la facultad de ciencias
químicas de la universidad Veracruzana. Licenciatura, Universidad Veracruzana.
Martí, N. O. (2006). Phosphorus in anaerobic digestion process. Florida.
Mingqing, Y. (2008). Annual Review of Chinese Environmental Law Developments: 2008. In:
INSTITUTE, E. L. (ed.). Washington D. C. .
Monnet, F. (2003). An introduction to anaerobic digestion of organic wastes. Scotland: Remade
Scotland.
Ocaña, F. J. (2011). Biodigestor anaerobio de laboratorio. Licenciatura, Universidad Carlos III de
Madrid.
Ruíz, L. B. (2010). ¿Papel o plástico. ¿Cómo ves? . México: Universidad Nacional Autónoma de
México.
Scott, G. (2000). Green polymers. Polymer Degradation and Stability.
Scott, G., Lemaire, J., Jakubowich, I., Ojeda, T. & Packham, C. (2011). Executive summary on oxobiodegradable plastic. In: ASSOCIATION, O.-B. (ed.). Londres.
SEMARNAT-SNIARN (2012). Residuos Sólidos Urbanos. In: NATURALES, S. D. M. A. Y. R. N.-S.
N. D. I. A. Y. D. R. (ed.). México.
Sosa, M. A. (2002). Los plásticos: materiales a la medida. ¿Cómo ves? México: Universidad
Nacional Autónoma de México.
Thompson, R. C., Swan, S. H., Moore, C. J. & Saal, F. S. (2009). Our Plastic age. Philosophical
Transactions of The Royal Society
Verma, S. (2002). Anaerobic digestion of biodegradable organic in municipal solid wastes. Master,
Columbia University.
Veyna, S. C. (2007). Efecto de control de pH, temperatura y adición de nitrógeno sobre la digestión
anaerobia de residuos hortícolas. Maestría, Instituto Politécnico Nacional.
943
CHAPTER 8.3. BIODEGRADATION OF ECOVIO® COMPOSTABLE PLASTIC
BY COMPOSTING AND BIOAUGMENTATION
Xochitl Quecholac Piña*(1); Mariel Anel García Rivera (1);
Rosa María Espinosa Valdemar (1); Alethia Vázquez Morillas (1);
Margarita Beltrán Villavicencio (1)
(1) Universidad Autónoma Metropolitana, Unidad Azcapotzalco, México DF, México.
ABSTRACT
Biodegradable plastics have been developed as an option to diminish the
environmental impact of these materials at the end of their useful life. The aim of
this research was to assess the biodegradation of the compostable plastic film
Ecovio® in a bioaugmented composting system, mixed with organic waste. Two
200 l reactors were used to contain 50% of organic waste, 25% of mulch and 25%
of dry leaves, in weight basis. Bioaugmentation was performed by adding 4 kg of
brown sugar loaf and 1.4 kg of yeast in each reactor. One hundred Ecovio® strips
(1.5cm x 12 cm) were added to one of the reactors. The initial waste mixture was
characterized by measurement of pH, humidity, organic matter, ashes and
nitrogen. Temperature, volume and weight of the waste was monitored along the
process. Manual aeration of the mixtures was performed daily in the first 15 days
and every two weeks afterwards. The degradation of the plastics was evaluated
measuring its elongation at break. Plastic strips were sampled on days 10, 15, 20,
25, 45 and 57. In those samplings pH, humidity, organic matter, ashes and
nitrogen were evaluated in the compost. At the end of the process Ecovio® has
been degraded in 61.57% (mass basis), and it was not possible to recover
complete strips for the elongation test. The composts obtained in both reactors
complied with the parameters of organic matter and nitrogen set in the local
regulation for composting processes, NADF-020-AMBT-2011. No differences were
found between the reactors with and without plastic, so this material did not affect
the composting process.
Key words: bioaugmented, biodegradation, composting, plastic.
Introduction
Plastics are polymers of high molecular weight, formed by different combinations of
monomers, which form a variety of different materials (Eubeler, et al., 2009). Many
efforts were directed to increase the stability of plastics, but today some
applications require that plastics can be biodegraded in the environment (Imre &
Pukánszky, 2013). The first biodegradable plastics were created in the 80s as an
option to mitigate the end-of-life impacts of these materials, looking forward their
introduction in composting systems (Soroudi & Jakubowicz, 2013).
In general, composting can be done in piles or reactors. A reactor is a
process unit designed to phisically contain chemical reactions in an environment
944
with controlled operational parameters (humidity, aireation). Reactors can be fixed
or mobile. Mobile reactors make posible the continuous mixing of waste through
mechanical features, making the compostig process more homogeneous (Colomer
& Gallardo, 2007). One of the advantages of using a reactor to contain the
composting process is the possibility of having a better control, which allows the
continuous monitoring of the biodegradation (Tchobanouglos et al., 1994).
Composting
can
be
accelerated
by
bioaugmentation,
adding
microorganismos isolated from the same waste of from another source. This
addition can increase the conversión of organic substrates, reducing undesirable
reactions and the production of bad odours (Moreno & Moral, 2008). Another
possibility is the biostimulation, where a substance is added in order to increase
the activity of the endogenous microorganisms and to accelerate their metabolism.
The goal of this project is to asses the biodegradation of the compostable
plastic Ecovio® in a bioaugmented composting process and the effect of
introducing the plastic in the process. The quality of the resulting compost was
compared to the parameters set in the local standard NADF-020-AMBT-2011,
which sets the mínimum requirements for the production of compost from the
organic fraction of urban, agricultural, livestock and forestry waste, as well as the
mínimum quality parameters for the compost produced or distributed in Mexico
City.
Obtention of materials from the composting process. Vegetable waste was
obtained from the supply center in Atizapán de Zaragoza, Estado de México. Yard
waste, dry leaves and mulch were obtained from the Universidad Autónoma
Metropolitana, Unidad Azcapotzalco. Bioaugmentation was done by addition of a
PL[WXUH RI D VXJDU ORDI DQG ³0DJLGHO\´ \HDVW 7KH SODVWLF WHVWHG ZDV (FRYLR D
compostable film produced by Basf (Figure 1).
945
FIGURE 1. Materials used in the composting process. a) Food waste, b) dry
leaves, c) mulch, d) sugar loaf, e) levadura, f) compostable plastic Ecovio®.
Conditioning of materials. Food waste, dry leaves and mulch were used as
obtained. The sugar loaf was manually broken and it was disolved in wáter along
with the yeast. The plastic film was cut in strips (1.5 x 12 cm).
Set up of biorreactors. Two 200 L plastic bioreactors were used in this
experiment. Each of them was filled with a mixture containing food waste (50%),
mulch (25%) and dry leaves (25%, mass basis). The solution for bioaugmentation
for each reactor was prepared mixing 4 kg of sugar loaf, 1.4 kg of yeast and 4 L of
wáter. One hundred strips of Ecovio® plastic were added in one reactor, and the
other was used as a control. (Figure 2). Initial mass and volumen were recorded.
946
FIGURE 2. Mixing of the waste for the bioreactors. a) Mixing of food waste,
leaves and mulch, b) bioaugmentation and c) addition of plastic strips
Monitoring. Temperature, humidity, pH, organic matter, ashes, and nitrogen were
measured at days 0, 10, 15, 20, 25, 45 and 57. These analysis were done based
on the methods shown in Table 1, by triplicate.
TABLE 1. Phisicochemical analysis during the composting process
Parameter
Method
pH
NMX-AA-025-1984
Humidity
NMX-AA-016-1984
Fernández et at (2006)
Total nitrogen (micro kjendhal)
Organic matter
Espinosa, 2010
Ashes
NMX-AA-018-1984
C/N ratio
NMX-AA-067-1985
To promote aerobic conditions and an homogeneous mixture of the waste,
the content of the bioreactors was mixed and aereated manually, daily during the
first 15 days and every two weeks afterwards. (Figure 3).
947
FIGURE 3. Aireation and mixing of the waste contained in the bioreactors
Degradation of plastic films. The effect of the composting process in the plastic
films was assesed measuring their mechanical properties (elongation at break) in
an Universal testing machine Lloyd Instruments serie LF plus. Initial load was set
up at 2 N, the speed of the clamps was 150 mm/min and the distance between
them was 50 mm. In each sampling day 10 strips were collected, washed with
wáter to remove waste and dried with tissue paper, before measuring elongation at
break (Figure 4).
FIGURE 4. Sampling of plastic strips
Degradation of plastic. Elongation at break decreased along the process, as
expected. Figure 5 shows the absolute value of the elongation and the decrease in
this value. Elongation decreased in more tan 80% after 5 days, showing a quick
and easy consumptionf of the plastics by the microorganisms present in the
compost. For this kind of plastic, mechanical resistance is strongly related to
degradation. The phenomenon is the result of the combination of biotic and abiotic
factors.
In the last sampling, in day 57, it was not posible to recover complete strips to
test them. After sieving of the compost a 61.57% reduction in mass was measured
for Ecovio®.
948
Reduction in elongation at break (%)
100
Elongation at break (%)
600
80
500
400
60
300
40
200
20
100
0
0
0
10
15
20
25
45
Days
FIGURE 5. Biodegradation of plastics in the composting process
Reduction in elongation at
break (%)
700
Composting process. The initial mixture of organic waste showed characteristics
suitable to promote aerobic biodegradation, even if the C/N ratio was too high. A
high water content was measured, probably due to the presence of prickly pear
waste (Table 2).
TABLE 2. Characterization of the organic wastes mixture at the begining
Parameter
mixture at the begining
pH
6 ± 0.05
Humidity
77.72 ± 2.38 %
Total nitrogen (micro kjendhal)
0.62 ± 0.09 %
Organic matter
89.9 ± 1.2 %
Ashes
9.56 ± 1.01 %
C/N ratio
92.88 ± 1.68
In the fourth day of the process a microbial film (probably fungus) was
observed in both reactors (Figure 6). Water vapour was detected, produced by the
increase in temperature.
Volume loss was similar in both reactors. Final values reached 65.1% for the
compost containing Ecovio® and 57.47% for the control. Temperature showed a
typical profile for a composting process, with a mesophilic initial stage (T>45ºC). At
the end of this phase organic acids are produced and a mesophilic stage takes
place (T<45°C) that ends when temperature reaches atmospheric values. Figure 7
shows the average of temperature values along the process, with a máximum of
49°C for three days.
949
FIGURE 6. Microbial biofilm in the organic substrates
60
50
40
30
20
10
0
Temperature
Ecovio
0
20
Days
40
60
FIGURE 7. Temperature profile in the composting process
Humidity content moved between 65 and 57% (Figure 8), below the 70%
recomended as a maximum level to guarantee aerobic conditions (Bueno
Marquez, et al., 2011).
The pH affects directly the composting process due to their influence in the
microbial processes (Sundberg et al., 2004). Initial value for this experiment was 6,
and it increased as the composting advanced, due to the loss of organic acids and
production of ammonium by decomposition of proteins (Sánchez-Monedero, et al.,
2001). Final value was 9.2. It did not reach a neutral value, probably because the
lenght of the process did not allow the formation of humic substances (Bueno
950
Marquez, et al., 2011). No differences relating to pH were observed between the
reactors.
90
80
70
60
50
40
30
% Humidity
Ecovio
0
20
Days
40
60
FIGURE 8. Humidity profile in the biorreactors
As expected, nitrogen content increased due to the recycling of this element
produced by the death of microorganisms (Bueno Marquez, et al., 2011). Final
values were 2.08% for the reactor with plastics and 1.98% for the control. Organic
matter decreased from 89.99 % to 61.15 % for the control, and to 62.52 % for the
reactor with plastic. This is related to mineralization and the loss of carbón in the
way of gaseous compounds (Zucconi et al., 1987).
The content of ashes is related to the inorganic, non combustible substances
present in the substrates, such as mineral salts. Organic matter usually has a 530% content of ashes in dry basis (Ruiz, 2011). For this experiment at the end of
the process, they reached a 37 % value in mass, slightly higher tan the one
reported by Ruiz, 2011.
C/N ratio is the most relevan parameter in a composting process, as it affects
its speed and the rate of ammonium loss (Moreno y Moral, 2008). This value was
extremely high at the begining (92.8), due to the addition of sugar loaf and yeast.
By the end of the process it has reached a values between 15 and 21, in the limits
set by the NADF-020-AMBT-2011 and recommended by Moreno y Moral (2008),
as shown in Figure 9.
Table 3 shows the final characterization of the compost in both reactors,
compared to the limits set in the local regulation. The quality of the compost was
good, with a high content of organic matter and presence of nitrogen. C/N value is
slightly high, and pH is in the basic range, probably due to the kind of food waste
used in the experiment and the fact that it was concluded after only 57 days.
951
105
90
75
60
45
30
15
0
Control
C/N ratio
Ecovio
0
20
Days
40
60
FIGURE 9. C/N profile in the composting process
TABLE 3. Final characterization of the compost
Paramétros
Limits in NADFEcovio®
020-AMBT-2011
Humidity
25-45
54.35 ± 0.68
ph (1:9)
6.5-8
9.4 ± 0.05
Organic matter (%,
>20
62.5 ± 2.7
dry basis)
Nitrogen
(%, dry
1a4
2.08 ± 0.10
basis
C/N
<15
17.56 ± 0.57
Ashes (%, dry basis)
N.D
36.97 ± 2.23
N.D. Not determined
Control
58.11 ± 0.56
9.3 ± 0.05
61 ± 3.7
1.98 ± 0.04
17.40 ± 0.82
38.31 ± 3.14
Conclusion
Biodegradation of the compostable plastic film Ecovio® was achieved, as shown
by the fast decrease in mechanical properties (95% in 46 days) and reduction in
mass (61.57% in 57 days). Results indicate that this specific plastic can be
degraded effectively by composting.
The quality of the compost was not affected by the presence of plastic, and no
diferences were detected between the biorreactor containing plastic and the
control. Final characterization complies with most of the parameters set in the local
regulation NADF-020-AMBT-2011, and show a good agricultural quality
(Kiehl,1985). Bioagumentation was succesful, as it decreased the lenght of the
process in at least 30 days.
952
Acknowledgements
7KLV UHVHDUFK LV SDUW RI WKH SURMHFW ³Evaluación de la biodegradabilidad y
ecotoxicidad de plásticos oxodegradables y biodegradables en condiciones
DHURELDV \ DQDHURELDV´ IXQGHG E\ WKH )RQGR PL[WR &21$&<7 - Gobierno del
Distrito Federal, FM/CONACYT-GDV/05/2013
References
Bueno Marquez, P., Díaz Blanco, M. J., & Cabrera Capitán, F. (2011). Factores que afectan el
proceso de compostaje. In Compostaje (p. 570). Madrid: Ediciones Mundi-Prensa.
Colomer, M. J. F., & Gallardo, I. A. (2007). Tratamiento y gestión de residuos sólidos, México.
LIMUSA: Editorial Universidad Politécnica de Valencia.
Eubeler, J. P., Zok, S., Bernhard, M., & Knepper, T. P. (2009). Environmental biodegradation of
synthetic polymers I. Test methodologies and procedures. TrAC Trends in Analytical
Chemistry, 28(9), 1057±1072. doi:10.1016/j.trac.2009.06.007
GDF. (2012). NADF-020-AMBT-2012 que establece los requerimientos mínimos para la producción
de composta a partir de la fracción orgánica de los residuos sólidos urbanos, agrícolas,
pecuarios y forestales, así como las especificaciones mínimas de calidad de la composta.
Imre, B., & Pukánszky, B. (2013). Compatibilization in bio-based and biodegradable polymer
blends. European Polymer Journal, 49(6), 1215±1233. doi:10.1016/j.eurpolymj.2013.01.019
Moreno, C. J., & Moral, H. R. (2008). Compostaje. Madrid, España: Mundi-Prensa.
Ruiz, F. J. F. (2011). Ingeniería del compostaje. Texcoco, México: Universidad Autónoma de
Chapingo.
Sánchez-Monedero, M. A., Roig, A., Paredes, C., & Bernal, M. . (2001). Nitrogen transformation
during organic waste composting by the Rutgers system and its effects on pH, EC and
maturity of the composting mixtures. Biores. Technol, 78(3), 301±308.
Soroudi, A., & Jakubowicz, I. (2013). Recycling of bioplastics, their blends and biocomposites: A
review. European Polymer Journal, 49(10), 2839±2858. doi:10.1016/j.eurpolymj.2013.07.025
Sundberg, C., Smars, S., & Jonsson, H. (2004). Low pH as an inhibiting factor in the transition from
mesophilic to thermophilic phase in composting. Biores. Technol, 95(2), 145±150.
Tchobanouglos, G., Theisen, H., & Vigil, S. A. (1994). Gestión integral de residuos sólidos. Madrid,
España: Mc. Graw Hill.
Zucconi, F. D., & De Bertoldi, M. (1987). Compost specifications for the production and
characterization of compost from municipal solid waste.
953
CHAPTER 8.4. SUELO Y RESIDUO COMO TEMA OBLIGADO EN
PROGRAMAS DE EDUCACIÓN AMBIENTAL
Dante Camarillo-Ravelo, Diana G. Castro-Frontana, Katia J. López-Mendieta
ENCB-IPN, Departamento de Ingeniería en Sistemas Ambientales (ISA) México D.F., México.
RESUMEN
El plan de estudios con el que hasta el año 2013 opera la carrera de Ingeniería en
Sistemas Ambientales (ISA) de la Escuela Nacional de Ciencias Biológicas, del
IPN, data de 1998. Este plan requiere ser rediseñado de acuerdo a las
necesidades actuales y líneas de acción futuras que se vislumbran en las
diferentes ramas de la Ingeniería Ambiental.
Después de un análisis de referencias internas y externas, de la definición de
un nuevo perfil de egreso, misión y visión para los futuros ingenieros en Sistemas
Ambientales, se crearon tres líneas curriculares que responden a las tendencias
DFWXDOHV HQ OD IRUPDFLyQ GH ,QJHQLHURV $PELHQWDOHV (VWDV OtQHDV VRQ ³DJXD \
UHVLGXRVOtTXLGRV´³DLUHHQHUJtD\FDPELRFOLPiWLFR´\³VXHORV\UHVLGXRV´
El objetivo de este estudio es explicar los elementos que han sido tomados en
FXHQWD SDUD OD FUHDFLyQ GH OD OtQHD GH ³VXHORV \ UHVLGXRV´ /D FUHDFLyQ GH XQD
carrera que incluya el tema de suelos y residuos, está justificada por la existencia
en México del Programa Nacional de Remediación de Suelos Contaminados
(publicado a fines de 2010), así como por la existencia del Programa Nacional
para la Prevención y Manejo Integral de los Residuos, ambos competencia de la
Federación.
/D OtQHD FXUULFXODU GH ³VXHORV \ UHVLGXRV´ WLHQH FRPR FRPSHWHQFLD ILQDO
contribuir a la prevención, minimización y control de residuos (sólidos, especiales y
peligrosos), a través de acciones de minimización, separación en la fuente, reúso
y reciclaje de material, de energía y acciones de disposición final como última
opción. Además, propondrá la caracterización de residuos y suelos, de estrategias
de remediación de suelos contaminados mediante sistemas de gestión integral
técnicamente factibles, económicamente viables y socialmente aceptables, para
lograr la protección a la salud de la población y al ambiente. Dicha competencia a
su vez está asociada a saberes conceptuales, procedimentales y sociales. Para
alcanzar las competencia y los saberes asociadas a esta línea, se han propuesto
las siguientes unidades de aprendizaje: Muestreo y Caracterización de Suelos y
Residuos; Tratamientos físico-químicos de Suelos y Residuos; Tratamientos
Biológicos de Suelos y Residuos; Manejo Integral de Residuos Sólidos Urbanos;
Modelación; Manejo de Residuos Peligrosos, Remediación y Restauración, y
Temas Selectos de Suelos y Residuos. La información básica para el análisis fue
provista por el comité de Ingeniería en Sistemas Ambientales (ENCB-IPN, ISA)
954
Introducción
El plan de estudios con el que opera la carrera de Ingeniería en Sistemas
Ambientales (ISA) de la Escuela Nacional de Ciencias Biológicas, del Instituto
Politécnico Nacional (IPN), data de 1998. Este plan requiere ser rediseñado de
acuerdo a las necesidades actuales y líneas de acción futuras que se vislumbran
en las diferentes ramas de la Ingeniería Ambiental.
El Gobierno Federal a través del Plan Nacional de Desarrollo (PND) 20132018, establece cinco metas nacionales: México en paz, México incluyente,
México con educación de calidad, México próspero y México con responsabilidad
global; así como tres estrategias transversales, de observancia para todas las
dependencias y organismos: Democratizar la productividad, Gobierno cercano y
moderno y Perspectiva de género.
$SDUWLUGHODPHWDQDFLRQDO³0p[LFRFRQ(GXFDFLyQGH&DOLGDGÓD6HFUHWDUtD
de Educación Pública dirigió la elaboración del Programa Sectorial de Educación
(PSE) 2013-2018 y de todas las líneas de acción transversales que le
corresponden a dicho sector.
Con el propósito de elevar la calidad de la educación con equidad, el PSE se
fundamenta por un lado, en el artículo tercero Constitucional y en la Reforma
Educativa que modificó la Ley General de Educación y creó las leyes generales
del Servicio Profesional Docente y la del Instituto Nacional para la Evaluación de la
Educación. Dentro del PSE 2013-GHVWDFDHOVLJXLHQWHREMHWLYR³)RUWDOHFHUOD
calidad y pertinencia de la educación media superior, superior y formación para el
WUDEDMRDILQGHTXHFRQWULEX\DDOGHVDUUROORGH0p[LFR´
Para entender qué implica el nuevo plan de la carrera de Ingeniero en
Sistemas Ambientales es importante conocer los elementos del Modelo Educativo
Institucional (MEI, 2003) que han servido de base para la elaboración del Plan de
Estudios 2014, los cuales se detallan a continuación:
i) Centrado en el aprendizaje. Privilegia una formación que pone al estudiante en
el centro de la atención del proceso académico, considerándolo un individuo que
construye su propio conocimiento con el apoyo y la guía de sus profesores, diseña
y define sus propias trayectorias e intensidades de trabajo, dejando de lado la
concepción tradicional del estudiante como un ente abstracto, mero receptor de
conocimientos y de información.
ii) Promoción de una formación integral y de alta calidad científica, tecnológica y
humanística, y combine equilibradamente el desarrollo de conocimientos,
actitudes, habilidades y valores. Considera el desarrollo armónico de todas las
dimensiones del estudiante. Es decir, implica entre otros puntos, la formación en
los contenidos en cuanto a los conocimientos propios de la profesión, y de los
conocimientos básicos, aquellos que proporcionan las herramientas intelectuales
esenciales para el aprendizaje permanente y la resolución de problemas
complejos que requieren de la concurrencia de diversas disciplinas.
iii) Proporcionar una sólida formación y facilitar el aprendizaje autónomo. Se
refiere a la capacidad de aprender por sí mismo, primero con la guía de sus
profesores y luego sin ella, de manera que la institución se convierta en un
955
espacio al cual pueda recurrir el estudiante para actualizar sus conocimientos,
desarrollar y mejorar sus cualidades humanas.
iv) Expresar en procesos flexibles e innovadores, permitir el tránsito de los
estudiantes entre niveles educativos y cuente con múltiples espacios de relación
con el entorno. Esto significa que el modelo: reconoce las diferentes condiciones
económicas y sociales de los estudiantes, la diversidad de sus antecedentes,
capacidades y ritmos de aprendizaje, por lo que también la flexibilidad se expresa
en un currículum que ofrezca trayectorias formativas múltiples y opciones de
dedicación variable.
v) Formación bajo diferentes enfoques culturales para capacitar a los individuos en
su incorporación y desarrollo en un entorno internacional y multicultural. Incorpora
la internacionalización en la formación de los estudiantes, de manera tal que les
permita desarrollarse en un mundo multicultural. Significa también la asimilación
de la dimensión internacional a la esencia, identidad y cultura de la institución. Ello
requiere de disposición al cambio y a la transformación, programas flexibles y
normatividad que facilite el reconocimiento de créditos y la revalidación de
estudios realizados en otras instituciones educativas, y la participación en
programas y proyectos que desarrollen competencias, actitudes valores y
habilidades que formen a los estudiantes para su incorporación al entorno local,
nacional e internacional.
vi) Permitir que los egresados sean capaces de combinar la teoría y la práctica y
contribuyan al desarrollo sustentable de la nación. Privilegia los contenidos y
estrategias que combinen adecuadamente la formación teórica y la práctica,
concordante con los avances de la ciencia y la tecnología, la realidad del ejercicio
profesional, así como el conocimiento del entorno. De acuerdo con ello, el modelo
se constituye en el medio que proporciona oportunidades para que el estudiante
sea capaz de integrar correctamente los conocimientos, capaz de integrar
correctamente los conocimientos obtenidos en los distintos cursos o unidades de
aprendizaje, tanto en espacios curriculares de integración del conocimiento, como
en prácticas y servicios fuera de la institución. Requiere de una apropiada
selección de los contenidos curriculares. También se demanda atención
individualizada que reconozca que los estudiantes tienen antecedentes formativos
distintos, conocimientos, habilidades y capacidades diferentes que los
aprendizajes se logran en medidas y por vías diversas.
Con lo que respecta a la enseñanza de la ingeniería y el diseño de planes y
programas, hoy en día, la formación de un ingeniero se da en etapas sucesivas,
tanto en la escuela como en la práctica. Dos habilidades principales son funciones
centrales de un ingeniero: el diagnóstico y el diseño. Ambas tienen aspectos que
sólo se pueden aprender si el recién egresado pasa varios años de práctica
subordinada dentro de equipos de trabajo que incluyan ingenieros con
experiencia. Por lo tanto, la importancia y proporción del aprendizaje que ocurre
en la escuela y en la práctica son equiparables, aunque la naturaleza de lo que se
aprende en una y otra es distinta.
Resendiz (2008) hace una interesante reflexión sobre la importancia de la
ingeniería en el pasado y en el mundo actual y destaca los elementos relevantes y
alcances que debe tener la enseñanza de la ingeniería en las instituciones
956
académicas. Por un lado, parte del hecho de que la formación de un ingeniero
ocurre tanto en el ambiente escolarizado (indispensable en estos tiempos para el
aprendizaje de la teoría) como en la práctica (bajo la tutela de ingenieros con más
experiencia). Por lo tanto, no puede pretenderse que una institución educativa por
sí sola, dé al futuro ingeniero los dos tipos de saber que requerirá para el ejercicio
de su profesión: teoría y práctica. El autor concluye entonces que las condiciones
necesarias que deben darse en las instituciones académicas para que éstas
puedan cumplir con su rol y con los servicios que les corresponden en la
enseñanza de la ingeniería, son:
x ³Tomar en cuenta, al diseñar planes y programas de estudio, lo que en la
formación de los ingenieros compete a la escuela y lo que corresponde a la
SUiFWLFD SURIHVLRQDO´ /RV SURIHVRUHV HQFDUJDGRV GHO GLVHxR GH SODQHV \
programas deben ser capaces de discernir entre las exigencias que se le
imponen a las instituciones educativas para poder elaborar programas de
estudio pertinentes y efectivos puesto que: a) es imposible incluir en un
programa de estudios de cuatro o cinco años todo lo que un buen ingeniero
debe saber, b) algunos conocimientos se aprenden mejor en la práctica
profesional, c) un ingeniero recién graduado debe poseer un excelente
bagaje de conocimientos teóricos y capacidades básicas para aplicarlos, y
que en el futuro le permitan incrementar sus conocimientos durante la
experiencia práctica que a su vez le permitirá desarrollar su juicio
profesional.
x ³&RQRFHUODILORVRItD\ODVQRUPDVGHregulación nacional de la práctica de la
ingeniería, así como las peculiaridades de la misma en el ámbito
LQWHUQDFLRQDO´
x ³0DQWHQHUXQEDODQFHDSURSLDGRHQODSURSRUFLyQHQWUHSURIHVRUHVFRQSHUILO
y ocupación esencialmente académicos (docencia e investigación) y
SURIHVRUHVFRQH[SHULHQFLD\DFWLYLGDGHQODSUiFWLFDSURIHVLRQDO´
x ³'DU XQ PD\RU pQIDVLV D OD FDOLGDG GH OD HGXFDFLyQ TXH D FXDOTXLHU RWUR
criterio en la toma de decisiones internas´
El Modelo Educativo del IPN (MEI, 2003) es congruente con esta visión. La
formulación de planes y programas de estudio se enfrenta entonces a dos hechos
ineludibles y contrapuestos: a) que el acervo de conocimientos relevantes crece
continuamente, y b) que en un lapso dado no puede aprenderse bien más que
cierta porción de él. Por tanto, la decisión de introducir un nuevo tema en los
planes de estudio debe llevar al parejo la decisión de eliminar otro tema de
extensión similar o comparable, o de lo contrario se corre el riesgo de
congestionar al estudiante y volver el aprendizaje superficial (Resendiz, 2008).
En este contexto, el rediseño de la carrera de ISA define dos ámbitos de
desempeño que tendrán los futuros ingenieros(as) en sistemas ambientales.
Dichos ámbitos son el de la Evaluación Ambiental y el de la Ingeniería Ambiental.
En general, el ámbito de desempeño de la Ingeniería Ambiental, se refiere a
prevenir, minimizar y controlar la contaminación en el ambiente así como contribuir
la rehabilitación de ambientes impactados, a través de medidas tecnológicas.
957
/D FRPSHWHQFLD SURIHVLRQDO D GHVDUUROODU HV ³3URSRQHU ORV VLVWHPDV GH
prevención, minimización y control de contaminantes en el ambiente, para
disminuir el grado de deterioro ambiental derivado de las actividades antrópicas,
que permita a los sistemas productivos cumplir no sólo con la normatividad
ambiental vigente y con estándares internacionales, sino proponer sistemas bajo
HOHQIRTXHGHODHFRHILFLHQFLD´
Para lograr esta competencia el ámbito de la Ingeniería Ambiental está
constituido por tres líneas curriculares que responden a la atención de
QHFHVLGDGHV HQ ORV UXEURV ³$JXD \ UHVLGXRV OtTXLGRV´ ³$LUH HQHUJtD \ FDPELR
FOLPiWLFR´\³6XHORV\5HVLGXRV´$VXYH]FDGDXQDGHHVWDVOtQHDVFXUULFXODUHV
tiene su propia competencia general y sus respectivos saberes conceptuales,
procedimentales y saberes sociales.
A continuación se presenta la metodología general que se siguió para la
FUHDFLyQGHOQXHYRSODQGHHVWXGLRGH,6$HVSHFtILFDPHQWHGHODOtQHDGH³6XHORV
\5HVLGXRV´\ODVQHFHVLGDGHVDODs que responde. La inclusión de la temática de
suelos como campo de acción novedoso dentro de la ingeniería en sistemas
ambientales y de la ingeniería ambiental responde a que cada día hay más
demanda de especialistas en remediación y restauración de sitios contaminados.
Sin embargo, este rubro aún no ha sido incluido en la mayoría de los planes de
estudios de universidades que imparten estas carreras.
Procedimiento metodológico
Para el rediseño de la carrera de Ingeniero en Sistemas Ambientales se utilizó
FRPR JXtD HO ³0DQXDO SDUD HO UHGLVHxR GH SODQHV \ SURJUDPDV HQ HO PDUFR GHO
QXHYR PRGHOR HGXFDWLYR \ DFDGpPLFR´ SXEOLFDGR SRU HO ,QVWLWXWR 3ROLWpFQLFR
Nacional (MPR-IPN, 2004). El procedimiento consistió entonces en las siguientes
etapas:
Etapa 1. Análisis del marco de referencia del currículo. Esta etapa consiste en un
análisis de referentes institucionales o referentes internos (como el Modelo
Educativo del IPN, el Programa Interinstitucional) así como un análisis de los
referentes externos (PND 2013 ± 2018, PSE 2013 ± 2018, y diversos programas
específicos, entre ellos el Programa Nacional para la Prevención y Gestión Integral
de los Residuos (PNPGIR, 2008-2012), así como el Programa Nacional de
Remediación de Suelos Contaminados (PNRSC, 2010) entre otros, que dan una
guía sobre las necesidades prioritarias tanto para el país y por ende, para el
Instituto.
Etapa 2. Lineamientos generales para el plan de estudios. En esta etapa se
determina el objetivo y campo ocupacional de la carrera, los perfiles de ingreso y
egreso y los objetivos curriculares por área del plan de estudios.
Etapa 3. Diseño del Plan de Estudios. Consiste en la elaboración de contenidos y
estructura curricular así como en la determinación de los programas de
asignaturas y experiencias de aprendizaje y la puesta en marcha del plan de
estudios (hasta el momento este último paso está pendiente).
958
A partir del análisis de los referentes internos y externos (etapa 1) se obtuvieron
los siguientes hechos relevantes (para la línea de Suelos y Residuos):
En México, hasta el año 2009, la Dirección General de Gestión Integral de
Materiales y Actividades Riesgosas tenía registrados 297 sitios considerados
como pasivos ambientales y tan sólo había gestionado la remediación de
aproximadamente 50 pasivos ambientales que tienen un responsable. Las
principales causas de contaminación han sido derrames de hidrocarburos, la
disposición inadecuada de residuos peligrosos, la degradación de predios
industriales y la disposición histórica de sitios mineros. Existen sitios contaminados
con residuos peligrosos cuya atención es crítica: Metales y Derivados en Tijuana
(Baja California), La Pedrera en Guadalcázar (San Luis Potosí), CYTRAR en
Hermosillo (Sonora), etc. Es por lo tanto una prioridad ambiental para el país
prevenir la generación de nuevos sitios contaminados y reducir el número de sitios
ya existentes con el fin de prevenir y reducir los riesgos a la salud y al ambiente
(PNRSC, 2010; DGGIMAR, 2014).
El Programa Nacional de Remediación de Suelos Contaminados (PNRSC,
2010) establece cinco objetivos principales entre los que destaca por su
SHUWLQHQFLDSDUDHVWHHVWXGLRHOREMHWLYR³&RQWDUFRQHOVRSRUWHDFDGpPLFR\GH
investigación que permita continuar con el desarrollo de la gestión de sitios
FRQWDPLQDGRV´'HHVWHREMHWLYRVHGHVSUHQGHQWUHVPHWDV\ODV siguientes líneas
de acción relevantes:
³,PSXOVDUODLQYHVWLJDFLyQHQLQVWLWXFLRQHVGHHGXFDFLyQVXSHULRU\GHLQYHVWLJDFLyQ
>«@SDUDODUHPHGLDFLyQ\UHYLWDOL]DFLyQGHVLWLRVFRQWDPLQDGRV´
³,PSXOVDU OD LQYHVWLJDFLyQ HQ WHFQRORJtDV D WUDYpV GH OD UHDOLzación de
cooperaciones en proyectos de remediación y proyectos de investigación
HVSHFtILFRV´
³,PSXOVDUODLQYHVWLJDFLyQFRQUHVSHFWRDODHYDOXDFLyQGHULHVJRVDPELHQWDOHV\D
la salud humana a través de la realización de cooperaciones en proyectos de
rePHGLDFLyQ\SUR\HFWRVGHLQYHVWLJDFLyQHVSHFtILFRV´
³,PSXOVDUODLQYHVWLJDFLyQFRQUHVSHFWRDODUHXWLOL]DFLyQ\UHYDORUL]DFLyQGHVLWLRV
contaminados a través de la realización de cooperaciones en proyectos de
UHPHGLDFLyQ>«@´
Por otro lado, el Programa Nacional para la Prevención y Gestión Integral de
los Residuos (PNPGIRC, 2009-2012) así como el documento Diagnóstico Básico
para la Gestión Integral de los Residuos 2012 (DBGIR, 2012) ambos publicados
por la SEMARNAT, tienen como finalidad, respectivamente, establecer la política
nacional en materia de manejo de los residuos y actualizar la información referente
al manejo de residuos sólidos urbanos, residuos de manejo especial, residuos del
petróleo, residuos de la minera y residuos peligrosos. Los documentos anteriores
señalan que en México se generan aproximadamente 9.1 millones de toneladas
de residuos peligrosos anualmente, 37.5 millones de toneladas de residuos sólidos
urbanos y 84 millones de toneladas anuales de catorce corrientes de residuos de
manejo especial.
959
La creación de una línea que incluya el tema de suelos y residuos se justifica
entonces con la existencia del Programa Nacional de Remediación de Suelos
Contaminados así como con el Programa Nacional para la Prevención y Gestión
Integral de los Residuos. En este último se establecen también estrategias y líneas
de acción a seguir. Una vez analizados estos referentes (y habiendo definido los
dos ámbitos de desempeño de la carrera y sus respetivas competencias) se
obtuvo la competencia específica de la línea curricular de suelos y residuos que a
continuación se enuncia: contribuir a la prevención, minimización y control de
residuos (sólidos, especiales y peligrosos), a través de acciones de minimización,
separación en la fuente, reúso y reciclaje de material, de energía y a acciones de
disposición final como última opción. Caracterizar residuos y suelos y proponer
estrategias de remediación de suelos contaminados mediante sistemas de gestión
integral técnicamente factibles, económicamente viables y socialmente aceptables;
SDUD ORJUDU OD SURWHFFLyQ D OD VDOXG GH OD SREODFLyQ \ DO DPELHQWH´ /RV VDEHUHV
asociados a esta línea curricular se muestran en el TABLA 1.
En la Figura 1 se muestra la estructura del mapa curricular general de la
carrera de ISA destacando las unidades de aprendizaje (del área básica,
profesional y optativas) que contribuirán a lograr la competencia y a adquirir los
saberes relacionados con la trayectoria de Suelos y Residuos.
960
TABLA 1. Saberes a desarrollar en el ámbito de desempeño profesional de
Ingeniería Ambiental en la línea curricular de suelos y residuos de la carrera
de ISA
INGENIERÍA AMBIENTAL: SUELOS Y RESIDUOS
SABER CONCEPTUAL
SABER
SABER SOCIAL
PROCEDIMENTAL
Planea la logística para los sistemas Propone
estrategias Realiza un trabajo minucioso en
de tratamientos asociados a suelos y para la remediación de colaboración
con
otros
residuos.
suelos contaminados.
profesionistas.
Realiza muestreos de suelos y/o Diseña
planes
de Genera reportes confiables en
residuos.
manejo de residuos.
colaboración
con
otras
Caracteriza suelos y residuos.
Dimensiona
la disciplinas.
Explica las etapas del manejo infraestructura
Comunica sus ideas en forma
integral de los RSU, RME y RP.
necesaria
para
el verbal y escrita, asertiva y
Selecciona los métodos apropiados manejo de residuos respetuosamente a diferentes
para la remediación y restauración sólidos
urbanos, niveles.
de suelos.
residuos de manejo Coordina equipos de trabajo
Selecciona y aplica leyes y normas.
especial
y residuos multidisciplinarios.
Predice
la
dispersión
de peligrosos.
Colabora
eficientemente en
contaminantes en suelos mediante
equipos
de
trabajo
el uso de modelos
especializados.
Resuelve problemas con base
en objetivos y asume las
responsabilidades de la toma de
decisiones.
961
FIGURA 20. Mapa de las unidades de aprendizaje que contribuyen directa o indirectamente a adquirir los saberes
y lograr la competencia de la línea curricular de Suelos y Residuos.
962
Para alcanzar las competencia y los saberes asociados a esta línea, se han
propuesto las siguientes unidades de aprendizaje: Manejo integral de Residuos (como
unidad de aprendizaje obligatoria directamente relacionada con la línea curricular), y las
siguientes optativas: Muestreo y Caracterización de Suelos y Residuos; Tratamientos
físico-químicos de Suelos y Residuos; Tratamientos Biológicos de Suelos y Residuos;
Manejo Integral de Residuos Sólidos Urbanos; Modelación; Manejo de Residuos
Peligrosos, Remediación y Restauración, y Temas selectos de Suelos y Residuos. La
información básica para el análisis fue provista por el comité de Ingeniería en Sistemas
Ambientales (ENCB-IPN, ISA).
Los programas en extenso de cada una de estas unidades de aprendizaje
deberán estar acorde con el modelo educativo institucional (MEI, 2003), es decir, el
diseño detallado de los contenidos deberá basarse no sólo en los elementos teóricos
que debe aprender un ISA que elija esta línea, sino en una planeación basada en
competencias, donde los conocimientos se refuercen con habilidades, actitudes y
valores que proveaQ³HOVDEHUVHU\HOVDEHUKDFHU´GHORVIXWXURVHJUHVDGRVGH,6$
Conclusión
/D FUHDFLyQ GH OD OtQHD ³6XHORV \ 5HVLGXRV´ GHQWUR GHO SURJUDPD GH HVWXGLR GH OD
carrera de Ingeniero en Sistemas Ambientales de la ENCB se justifica con la existencia
del Programa Nacional de Remediación de Suelos Contaminados así como con el
Programa Nacional para la Prevención y Gestión Integral de los Residuos. La creación
de esta línea responde a las necesidades nacionales en materia de manejo de residuos
y de remediación de pasivos ambientales y la necesidad de generar capital humano
capacitado para atender esta problemática.
No está por demás mencionar que a nivel mundial estos dos rubros también son
de gran relevancia en las tendencias actuales de investigación y formación profesional
en la ingeniería ambiental.
Agradecimientos
Los autores agradecen al Instituto Politécnico Nacional su apoyo para la realización de
este proyecto y sobre todo al Comité de Rediseño de la carrera de Ingeniería en
Sistemas Ambientales, por otorgar la información de base para presentar este trabajo.
Referencias
DBGIR (2012). Diagnóstico básico para la gestión integral de residuos. INEEC-SEMARNAT. México.
Consultado
el
9
de
agosto
de
2014.
http://www.inecc.gob.mx/descargas/dgcenica/diagnostico_basico_extenso_2012.pdf
MPR-IPN (2004). Materiales para la reforma. Manual para el rediseño de planes y programas en el
marco del nuevo modelo educativo y académico. México, D.F.: Instituto Politécnico Nacional.
963
MEI (2003). Un nuevo modelo educativo para el IPN. México, D.F.: Instituto Politécnico Nacional. ISBN
970-36-0077-8
PND (2013-2018). Plan Nacional de Desarrollo 2013-2018. Consultado el 9 de agosto de 2014.
http://pnd.gob.mx/
PNPGIR, 2009-2012. Programa Nacional para la Prevención y Gestión Integral de los Residuos.
Consultado el 9 de agosto de 2014. http://www.semarnat.gob.mx/programas/programas-cargo-de-lasemarnat-2006-201-12
PNRSC, 2010. Programa Nacional de Remediación de Sitios Contaminados. Gobierno Federal.
Secretaría de Medio Ambiente y Recursos Naturales-GTZ. México.
PSE (2013-2018). Programa Sectorial de Educación 2013-2018. Secretaría de Educación Pública.
Consultado
el
9
de
agosto
de
2014.
http://www.sep.gob.mx/es/sep1/programa_sectorial_de_educacion_13_18#.U-azU-OHiSo
Reséndiz-Núñez, D. (2008). El rompecabezas de la ingeniería. Por qué y cómo se transforma el mundo.
México, D.F.: Fondo de Cultura Económica. ISBN: 978-968-16-8444-0
DGGIMAR, 2014. Dirección General de Gestión Integral de Materiales y Actividades Riesgosas.
Consultado
el
9
de
agosto
de
2014.
http://www.semarnat.gob.mx/temas/gestionambiental/materialesactividades
964
CHAPTER 8.5. APROVECHAMIENTO DE RESIDUOS DE JITOMATE (Lycopersicon
esculentum) HIDROPÓNICO COMO SUSTRATO PARA LA PRODUCCIÓN DEL
HONGO SETA (Pleurotus ostreatus)
Alma D. Sánchez-García (1); Rosa A. Guillen-Garcés (1);
Karla González-Velázquez (2)
(1, 2) Universidad Politécnica del Estado de Morelos, Ingeniería en Tecnología Ambiental. Jiutepec,
Morelos, México
RESUMEN
El cultivo de jitomate hidropónico (Lycopersocon esculentum) representa una
alternativa para satisfacer su demanda, ya sea a nivel mundial, nacional o estatal. Para
el estado de Morelos la producción de esta hortaliza es uno de los principales
productos agrícolas y el manejo de la técnica hidropónica se realiza en alrededor del
34% de sus municipios. Sin embargo, su producción también es generadora de miles
de toneladas de residuos orgánicos, se calcula que tan solo la producción con esta
técnica genera alrededor de 1 360 ton/ha/año, y por lo general, no reciben un manejo
adecuado, y se desconoce si existe una revalorización de los mismos, así como su
disposición final. Como una alternativa para el manejo de estos residuos y el
aprovechamiento de los nutrientes incorporados a los sistemas hidropónicos, se
propuso como objetivo general, evaluar el rastrojo de jitomate (RJ) hidropónico como
sustrato para la producción del hongo comestible Pleurotus ostreatus. Los resultados
indican que la producción de setas empleando mezclas de RJ hidropónico y paja de
trigo (PT) puede ser una alternativa para proporcionarle un valor agregado a este
residuo, ya que se obtuvieron EB hasta de un 50%, así como ciclos de cultivo (CC)
desde los 84 hasta los 117 días y tasas de producción (TP) del 0.31 al 0.58%.
Palabras clave: residuos orgánicos, jitomate hidropónico, hongos comestibles
Introducción
El cultivo del jitomate o tomate rojo (Lycopersicon esculentum), es uno de los productos
agrícolas con mayor demanda a nivel mundial, nacional y estatal, debido a la amplia
gama de productos que se elaboran y consumen a partir de este. Su demanda se ve
reflejada en la extensión territorial empleada para su obtención y en la gran cantidad de
empleos que genera su cultivo, cosecha y comercialización. De acuerdo con el Servicio
de Información Agroalimentaria y Pesquera (SIAP), tan solo durante el año 2012 se
calcula que generó alrededor de 72 mil empleos a nivel nacional, considerando la
producción a cielo abierto y de temporal y, en este mismo año, el país se ubicó a nivel
mundial, en el décimo lugar con una producción de 845 588 ton de esta hortaliza. A
nivel nacional, los estados que ocupan los principales sitios en su producción son
Sinaloa y Michoacán, el estado de Morelos no aparece entre los lugares más
destacados sin embargo, es importante mencionar que a pesar de ello, el cultivo de
965
jitomate es considerado como uno de las principales actividades socioeconómica en el
estado, de acuerdo con la Infografía Agroalimentaria de Morelos, publicado por la
SAGARPA y SIAP (2013), este cultivo ocupa el cuarto lugar dentro de la lista de
productos agrícolas destacados, representando el 7.4% del total de la producción
agrícola en el estado, así mismo representa importantes cadenas productivas en la
región debido a que el 54.5% de sus municipios se dedican a esta actividad, ocupando
alrededor de 2, 591 hectáreas de cultivo a cielo abierto, esto considerando la superficie
de riego y de temporal, para lo que se emplean alrededor de 1 800 productores
registrados formalmente, además de una amplia fuente de empleos temporales ya que
durante su ciclo de producción (que es alrededor de 100 días) se utilizan
aproximadamente 1 000 jornales por hectárea cultivada (CESVMOR, 2010).
De la producción total anual de jitomate en Morelos, un porcentaje es generado
por cultivos semihidropónicos, llamados hidropónicos de forma general por los
productores. De acuerdo con encuestas realizadas por el grupo de investigadores de la
Universidad Politécnica del Estado de Morelos, se reporta que en el estado existen
alrededor de 54 productores formales de jitomate hidropónico, distribuidos en 49
invernaderos localizados en 13 de sus municipios (Ayala, Coatlán del Río, Jantetelco,
Jojutla, Jonacatepec, Mazatepec, Miacatlán, Puente de Ixtla, Temixco, Xochitepec,
Yautepec, Yecapixtla y Zacatepec), su producción alcanza alrededor de 1, 386.8 ton al
año, en dos ciclos de cultivo.
Debido a la importancia que este cultivo representa, es de considerarse la
cantidad de residuos orgánicos que se generan a partir del mismo, ya que únicamente
se conocen datos generales, como los citados por Sánchez et al., (2008) quienes
mencionan que anualmente se generan alrededor de 14.4x106 ton/año, sin embargo, la
Semarnat en su informe del Programa Nacional para la Prevención y Gestión Integral
de los Residuos 2009-2012 (Semarnat, 2013) mencionan que actualmente se carece
de información acerca de la generación de los residuos de origen agrícola a nivel
nacional, ya que no se carece de metodologías para el muestreo de generación, así
como de una clasificación de éstos, por lo que se carece de información acerca del
volumen generado por entidad federativa, su manejo y disposición final, debido a que
esto no se encuentra regulado por ninguna normatividad excepto por la LGPGIR, quien
menciona que estos residuos deben considerarse de manejo especial debido
únicamente, al volumen generado. De acuerdo con las técnicas culturales llevadas a
cabo en este cultivo, los productores cultivan alrededor de ocho plantas por m 2, y
dependiendo de la variedad cultivada (crecimiento determinado o indeterminado), al
finalizar el ciclo de producción una planta puede alcanzar alrededor de 17k de peso, lo
que genera aproximadamente 1 360ton/ha, tan solo de jitomate hidropónico en el
estado, los cuales, por lo general son desechados a cielo abierto sin ningún
tratamiento previo o aprovechamiento que permita un manejo responsable de los
mismos (Sánchez et al. 2008), esto sin mencionar el alto valor nutrimental con el que
cuenta el tejido vegetal eliminado, ya que si se analiza, la técnica hidropónica consiste
en desarrollar plantas de importancia económica sobre contenedores que contengan un
soporte inerte y aportar de forma equilibrada una formulación de minerales denominada
solución nutritiva (SN), todo esto bajo el estricto control de factores tales como el pH y
la conductividad eléctrica (Sánchez et al. 2007). La solución nutritiva representa uno
966
de los rubros económicos más elevados de este sistema de producción, y a pesar de
esto el 30% aproximadamente es lixiviada, el resto se destina al desarrollo de biomasa
y al metabolismo vegetal y solo una pequeña parte es empleada en la generación de
frutos. De manera que, gran porcentaje de la mayor inversión de la producción
hidropónica no es aprovechada dentro del mismo sistema, sino que es eliminada como
parte de la materia orgánica generada al final de la cosecha, con muy pocas
alternativas para su manejo, entre las que se pueden mencionar: el apilaje a cielo
abierto, la quema y, en muy pocas ocasiones el compostaje aerobio, sin embargo, se
desconocen los parámetros y medidas de control empleadas en dichos sistemas.
Una posible alternativa para el manejo de estos residuos que permita el
aprovechamiento del valor nutrimental proporcionado al cultivo mediante la solución
nutritiva, además de su revalorización (LGPGIR, 2003), mediante la generación de una
técnica que permita emplear este material de desecho como sustrato para el cultivo del
hongo seta (Pleurotus ostreatus), el cual se caracteriza por su capacidad para degradar
lignina e hidrolizar la celulosa (Sánchez y Royse 2001), debido a que las enzimas
digestivas extracelulares que posee son capaces de degradar hasta el 60% del
contenido de estas sustancia presentes en el tejido vegetal (Guzmán et al. 1993).
Además, con esta alternativa, podría aprovecharse al máximo las concentraciones de
nutrientes minerales proporcionados en la solución nutritiva durante las diferentes
etapas del desarrollo del cultivo. Finalmente, se reduciría de forma considerable, el
volumen de los residuos generados por este cultivo, ya que los hongos de pudrición
blanca a los que pertenece el género Pleurotus son una forma eficiente de
bioconversión de estos residuos en proteína comestible (Sánchez et al. 2008),
facilitando su posterior composteo y generación de humus (Vega et al, 2005). Por todo
lo anterior, en la presente investigación de propuso como objetivo general, evaluar el
rastrojo de jitomate (RJ) hidropónico como sustrato para la producción del hongo
comestible Pleurotus ostreatus.
El presente trabajo se llevó a cabo en las instalaciones de la Universidad Politécnica
del estado de Morelos. Las técnicas empleadas para la determinación de los factores
fisicoquímicos fue la NMX-FF-109-SCFI-2007. La medición de pH se realizó con el
método electrométrico con ayuda de un potenciómetro, para determinar la
conductividad eléctrica (expresada en DSM-1) se empleó el conductivímetro. El
porcentaje de humedad de la muestra se realizó mediante el método gravimétrico
empleando la siguiente fórmula para su cálculo:
Ψ۶ ൌ
ሺ‫۾‬۰ା‫۾‬۶۶‫ۺ‬ሻିሺ‫۾‬۰ା‫܁۾‬۶‫ۺ‬ሻ
ሺ‫۾‬۰ା‫۾‬۶۶‫ۺ‬ሻି‫۾‬۰
ൈ ૚૙૙
[1]
En donde:
% H = Contenido de humedad gravimétrico expresado en Porcentaje (%)
967
PB = Peso del crisol (g)
PHHL = Peso sustrato (g)
PB + PHHL = Peso del crisol más peso húmedo del sustrato (g)
PB + PSHL = Peso del crisol más peso seco del sustrato (g)
Las pruebas para la determinación de cenizas, materia orgánica y carbón orgánico se
realizaron por medio del método de calcinación, considerando las siguientes formulas:
Cenizas
Ψ ൌ
ሺ୔େେି୔େሻ
୫୳ୣୱ୲୰ୟሺ୥୰ሻ
ൈ ͳͲͲ
[2]
En donde:
PCC = Peso crisol con cenizas
PC = Peso del crisol
100 = Para referirlo a porcentaje
Materia orgánica
Ψ‫ ۽ۻ‬ൌ ͳͲͲ െ Ψ
[3]
En donde:
% MO = Materia Orgánica expresada en porcentaje
% Cenizas = Calculada en el punto
Carbón orgánico
Ψ۱‫ ۽‬ൌ
Ψ୑ୟ୲ୣ୰୧ୟ୓୰୥ୟ୬୧ୡୟ
ଵǤ଻ଶସ
[4]
En donde:
%CO = Carbono Orgánico en porcentaje
%MO = Materia Orgánica calculada a partir del % cenizas
1.724 = Factor Van Benmelen
La determinación del nitrógeno total se realizó mediante el método de Kjendahl.
Los cálculos se realizaron mediante la siguiente ecuación:
Ψ‫ ܂ۼ‬ൌ
୚ൈ୬ൈଵସ
୔
ൈ ͳͲ
[5]
En donde:
V = Gasto de ácido clorhídrico consumido en mL
n = Normalidad del ácido clorhídrico
968
14= Peso molecular del nitrógeno.
10 = Para referirlo a por ciento
P = Peso de la muestra en gramos
La relación Carbono-Nitrógeno se calculó mediante la siguiente ecuación.
۱
‫ۼ‬
ൌ
େ୓ሺΨሻ
୒୘ሺΨሻ
[6]
En donde:
C:N = Relación Carbono-Nitrógeno.
CO = Porcentaje de Carbono Orgánico
NT = Porcentaje de Nitrógeno total Kjeldahl
Todas las evaluaciones fisicoquímicas se realizaron por triplicado.
El cultivo se realizó en condiciones de invernadero, bajo la técnica tradicional
conocida rudimentaria (Sanchez y Royse 2001; Garzón y Cuervo, 2008), la cual
consiste en los siguientes pasos: a) reducción del tamaño de partícula a un promedio
de 15cm, b) pretratamiento de los materiales mediante un proceso de pasteurización
utilizando agua corriente a 70°C, durante 45min una vez transcurrido este tipo se
empleo aire frío para generar un shock térmico y eliminar microorganismos así como
reducir la temperatura y eliminar el exceso de humedad hasta el 70%; c) la inoculación
se realizó colocando capas de sustrato intercaladas con inoculo (reaislada sobre
granos de trigo a partir de la cepa de Pleurotus ostreatus (Jacq.: Fr.) Kumm HEMIM 50
del laboratorio de Micología del CIB-UAEM) en bolsas de plástico transparente de 40 x
60cm, perforadas previamente con agujas de disección. Cada bloque contenía 3Kg de
sustrato mezclado (residuo de jitomate/paja de trigo) y 250 g de inoculo. Los bloques
se colocaron en anaqueles para su incubación hasta la aparición de los primordios.
Para estimular la fructificación, los bloques se trasladaron a un espacio con
fotoperíodos de 12h luz al 50%, en donde fueron regadas de forma manual para
alcanzar una humedad relativa del 60% y reducir la temperatura de incubación, para
estimular el desarrollo de los frutos. Las formulaciones empleadas como sustrato
fueron las mostradas en la Tabla 1.
969
TABLA 1. Formulaciones (%) evaluadas para la producción del hongo seta
(Pleurotus ostreatus)
Sustratos
Control
negativo
Tratamiento A
100
B
Residuos de
75
jitomate (RJ)
Paja de trigo
25
(PT)
Las formulaciones se realizaron por quintuplicado
C
50
D
25
Control
positivo
E
0
50
75
100
Mezclas (%)
Para la evaluación de la Eficiencia Biológica (EB) se consideraron las siguientes
características:
a) Tiempo de incubación hasta la aparición de primordios
b) Tiempo de aparición de la primer cosecha
c) Evaluación de la eficiencia biológica
ൌ
୮ୣୱ୭୤୰ୣୱୡ୭ୢୣ୪୭ୱୡ୳ୣ୰୮୭ୱ୤୰୳ୡ୲୧୤ୣ୰୭ୱ
୮ୣୱ୭ୱୣୡ୭ୢୣ୪ୱ୳ୱ୲୰ୟ୲୭
ൈ ͳͲͲ [7] d) Ciclo del cultivo
ൌ ሺ×ሻ ൅ × [8]
e) Tasa de producción
୉୆
ൌ େେ [9]
Tomando únicamente los valores de las tres primeras cosechas.
Finalmente, se evalúo la pérdida de peso, considerando la diferencia entre el peso
húmedo inicial y el peso del sustrato agotado al finalizar el monitoreo en la tercer
cosecha.
970
Los resultados obtenidos a partir del análisis fisicoquímico se muestran en la Tabla 2.
TABLA 2. Análisis fisicoquímico obtenido del sustrato de paja de trigo y rastrojo
de jitomate
Sustrato
Conductividad
Humedad
Cenizas
Materia
Carbón
Nitrógeno
Eléctrica (DSM-1)
(%)
(%)
orgánica
orgánico
total
6.66
0.10
65
6.01
93.99
54.24
0.58
6.67
0.049
69
5.88
94.17
54.62
0.28
pH
Rastrojo de
jitomate
Paja de
trigo
Los resultados del análisis fisicoquímico mostraron poca diferencia entre el
potencial hidrógeno (pH) del rastrojo de jitomate (RJ) y la paja de trigo (PT), no así en
el caso de la Conductividad Eléctrica (CE), ya que en el caso del RJ fue de 0.10 DSM1, y para la PT de 0.049 DSM-1, esto seguramente influenciado por la concentración de
sales minerales proporcionadas al cultivo de jitomate hidropónico por medio de la
solución nutritiva (Resh, 2006). De con Sánchez y Royse (2001) y Garzzón y Cuervo
(2008), los niveles de pH adecuados para el cultivo de setas es de 5.5 y 6.5, otros
autores mencionan (Fracchia et al. 2009; Cayetano y Bernabe, 2008) la adición de
CaCo3 o CaSO4 para regular los niveles de pH a neutro, sin embargo en esta
investigación se consideraron adecuados los niveles de pH naturales de los sustratos.
El porcentaje de humedad obtenido para ambos sustratos (RJ y PT) se calculó en
base a lo recomendado por la literatura (Garzón y Cuervo, 2008; Sánchez y Royse,
2001; Guzmán et al. 1993), que es de 50-80%. Sin embargo, en un bioensayo
preliminar, no se obtuvieron buenos resultados, debido a que el RJ tiende a retener por
más tiempo un porcentaje de humedad elevado, a diferencia de la PT que pierde
humedad durante el período de incubación; esto no solo limito el desarrollo micelial,
sino que favoreció la contaminación biológica del sustrato. Más adelante se mencionan
las modificaciones a la técnica rudimentaria, para lograr el desarrollo de cuerpos
fructíferos.
Existen pocas investigaciones relacionadas al cultivo de Pleurotus ostreatus, en
las que se realicen análisis fisicoquímicos para caracterizar los sustratos. Algunas de
estas investigaciones mencionan que, los hongos pertenecientes al género Pleurotus,
requieren mayores concentraciones de carbono que de nitrógeno para lograr la
degradación aerobia de los sustratos (Garzón y Cuervo, 2008). López et al. (2008)
971
indica que la concentración de carbón y nitrógeno influyen directamente sobre la
invasión micelial, en sus investigaciones determinaron que los sustratos con mayor
concentración de carbono que de nitrógeno se obtuvieron mayores eficiencias
biológicas en menor tiempo. Sus datos obtenidos en las evaluaciones de carbono en el
sustrato de aserrín de encino fueron de 51%, son similares a los obtenidos durante esta
investigación en el RJ 54.24% y, PT 54.62%, los cuales difieren de los obtenidos en el
capacho 28.31%, cáscara de arveja 25.5% y tuza de maíz 18%. Por su parte, las
concentraciones de nitrógeno variaron, de o.11% para el aserrín de encino, similar a la
concentración de nitrógeno en la PT que es de 0.28%, y 0.58% en el RJ, este
incremento posiblemente se deba a la nutrición vegetal que proporcionan los
productores a sus cultivos.
Debido a la retención de humedad por parte del RJ, fue necesario realizar varias
adaptaciones a la técnica tradicional conocida como rudimentaria, con la finalidad de
lograr el desarrollo micelial y la formación de cuerpos fructíferos (Figura 1).
La primer modificación fue el aplicar un pretratamiento al RJ, en la técnica
rudimentaria no es indispensable realizar ninguna técnica de pretratamiento, sin
embargo muchos autores mencionan desde la fermentación de los sustratos (Cayetano
y Bernabé, 2008; Guzmán-Dávalos et al. 1987b), hasta la esterilización en autoclave
(Fracchia et al. 2009; López et al. 2008; Garzón y Cuervo, 2008 y Sánchez et al, 2008),
sin embargo lo recomendado por su eficiencia y bajo costo son los tratamientos de
pasteurización (Guzmán et al. 1993; Sánchez y Royse, 2001; Sánchez et al. 2007) en
los que se emplean vapor de agua a 70-80°C/30-45min. En la presente investigación
fue indispensable realizar una desinfección sumergiendo el sustrato en una solución de
Cl al 10% durante 15min. posteriormente se realizó un pretratamiento de
pasteurización, sin embargo fue necesario incrementar la temperatura y el tiempo de
pasteurización a 80-90°C/60min.
972
x No es indispensable
x 70-80°C de temperatura
durante 30-45min.
x 70% de humedad
x Tamaño de partícula: 20 cm
Pretratamien
to
Pasteurizació
n
x Desinfección con Cl 10%
x 85-90 °C de temperatura
durante 60min.
x 40% humedad
x Tamaño de partícula: 20cm
Inoculación
x 20-25 días
x 20-25°C de temperatura
Incubación
x Desde 15 hasta 37 días,
dependiendo del sustrato
x 25-30°c de temperatura
Indicción a la
fructificación
Cosecha
),*85$ (VWDQGDUL]DFLyQ GH OD WpFQLFD ³UXGLPHQWDULD´ SDUD OD SURGXFFLyQ GHO
hongo seta (Pleurotus ostreatus) en mezclas con rastrojo de jitomate y paja de
trigo.
En la Tabla 3, se muestran los parámetros productivos obtenidos en los diferentes
tratamientos, obtenidas a partir de los promedios de las repeticiones.
No fue posible el cultivo en los tratamientos A y B, debido a que el alto contenido
de humedad retenida causo ocasiono la inhibición micelial e incremento la
susceptibilidad del sustrato a la contaminación biológica.
Los períodos de incubación fueron menores a los obtenidos por Sánchez et al.
(2008) quienes evaluaron este cultivo en rastrojo de jitomate, pero a diferencia de la
presente investigación, hidrataron previamente durante 18h hasta alcanzar un
porcentaje de humedad del 70%, emplearon bloques de 500gr con el sustrato de 0.9
cm de tamaño de partícula y finalmente, lo esterilizaron en autoclave a 121°C/2h. El
tiempo de obtención de la primer cosecha se reporta a los 40 días, en esta
investigación se reportan de 32 a 37 días de incubación para la aparición de primorios y
su posterior cosecha. En la Figura 2, se muestra el desarrollo de Pleurotus ostreatus
973
M O D I F I C A C I O N ES A L A T É C N I C A T R A D I C I O N A L
T É C N I C A T R A D I C I O N A L (Guzmán et al . 1993; Sánchez y Royse, 2001)
C ultivo tradicional
del hongo seta
sobre los residuos de jitomate hidropónico, tanto en su etapa de primordio, como el
cuerpo fructífero desarrollado.
Tabla 3. Parámetros productivos obtenidos
Tratamiento
Días de
EB (%)
incubación
CC
(días)
TP
(%)
Peso
final
(%)
C
36
39
84
0.47
51
D
37
52.75
117
0.31
55
E
32
67
115
0.58
32.8
Notas: EB, Eficiencia biológica; CC, Ciclo de cultivo; TP, Tasa de producción
FIGURA 2. Desarrollo de Pleurotus ostreatus sobre residuos de jitomate
hidropónico. a) Aparición de primordios, b) cuerposfructíferos de Pleurotus ostreatus
sobre mezclas de 50-50% rastrojo de jitomate hidropónico y paja de trigo,
respectivamente.
La EB obtenida fue de 36, 52.75 y 67% para la mezcla C, D y el control positivo,
respectivamente. Sánchez et al (2008) reporta valores de EB de 139.8, 112.4, 111.3%
empleando rastrojo de jitomate, rastrojo de jitomate con madera de vid y rastrojo de
jitomate con paja de trigo respectivamente. Sin embargo, López-Rodríguez et al (2008)
mencionan EB de 56.7% cultivando Pleurotus ostreatus sobre sustrato esterilizado de
tuza de mazorca, Cayetano y Bernabé (2008), emplearon residuos fermentados y
esterilizados de jamaica, paja de arroz y plátano obteniendo EB de 64.7%; datos
similares a los obtenidos por Guzmán-Dávalos et al (1987a, 1987b) quienes lo
cultivaron sobre residuos fermentados y esterilizados de bagazo de maguey y bagazo
de caña de azúcar, con EB de 54.6 y 51.05% respectivamente. Fracchia et al (2009)
sembraron esta misma especie sobre mezclas de residuos esterilizados de Paja de
trigo con Jatropha macrocarpa y Simmondsia chinensis, con EB de 60.3 y 76.2%; estos
974
mismos autores mencionan que por lo general, los valores de EB en paja de trigo
varían entre el 50-100% en la producción a nivel comercial, ya que esto depende de
diversos factores bióticos y abióticos.
La TP permiten analizar si los residuos pueden ser empleados como suplemento
en el cultivo de Pleurotus ostreatus (Curvetto et al. 2002). La TP, los CC así como la
EB obtenida en la presente investigación, puede variar en comparación con varias
investigaciones, sin embargo, los resultados obtenidos en el tratamiento D son
similares con otras evaluaciones similares sobre otros tipos de residuos
agroindustriales (Cayetano y Bernabé, 2008; Curvetto et al, 2002; Fracchia et al. 2009;
Garzón y Cuervo, 2008; López et al. 2008; Sánchez et al. 2008).
Este período de incubación, la EB, CC y TP pueden variar dependiendo del
sustrato empleado, de las condiciones de incubación y cultivo e incluso de la cepa
inoculada.
En cuanto a la pérdida de peso se obtuvo una reducción del peso inicial por
encima del 50% en tratamientos C y D y de 32.8% en el control positivo; esto debido a
que Pleurotus ostreatus presenta la capacidad enzimática que le permite degradar
polímeros como la lignina y celulosa, que componen la mayor parte de los residuos
agroindustriales, por lo que el cultivo de esta especie puede ser una alternativa para la
reducción de volúmenes de desechos agrícolas.
Conclusión
Fue posible OD HVWDQGDUL]DFLyQ GH OD WpFQLFD ³UXGLPHQWDULD´SDUD ORJUDUDSURYHFKDUORV
residuos de jitomate (Lycopersicon esculentum) hidropónico como sustrato para la
producción del hongo seta (Pleurotus ostreatus).
Los datos de EB, CC y TP permiten suponer que el RJ puede emplearse como
suplemento en el cultivo de Pleurotus ostreatus, preferentemente en mezclas de
25%RJ y 75% PT.
Las mezclas de 25% RJ-75% PT, así como 50% RH-50%PT, permiten la
reducción hasta del 50% del peso inicial de estos residuos agroindustriales.
975
Referencias
Cayetano, M., T. Bernabé. (2008). Cultivo de Pleurotus ostreatus sobre residues de las cosehcas de
Jamaica (Hibiscus sabdariffa) y plátano (Musa paradisiaca). Revista Mexicana de Mciología.
26:57-60.
Curvetto, N. R., Figlas, R., Devalis, S. Delmastro. (2002). Growth and productivity of different Pleorotus
spp strains on sunflower seed hulls supplemented with N-NH4 + and/or Mn (II). Bioresourse
Technology. 84:171-176.
Comité Estatal De Sanidad Vegetal Del Estado De Morelos (CESVEMOR). (2010). Manejo fitosanitario
del jitomate. Monitor agricola, 24:14-17.
Fracchia,S., A. Aranda, E. Terrizzano. (2009). Cultivo de una cepa comercial de Pleurotus ostreatus en
desechos de Simmondsia chinensis y Jatropha macrocarpa. Revista Mexicana de Micología. 29:
37-42.
Garzón, J. y J. Cuervo. (2008). Producción de Pleurotus ostreatus sobre residuos lignocelulosicos de
diferente procedencia. Nova publicación Científica en ciencias Bioquímicas. 6, 10:101-136.
Guzmán-Dávalos, L., D. Martínez-Carrera, P. Morales, C. Soto. (1987a). El cultivo de hongos
comestibles (Pleurotus) sobre bagazo de maguey de la industria tequilera. Revista Mexicana de
Micología. 3:47-49.
Guzmán-Dávalos, L., C. Soto, D. Martínez-Carrera. (1987b). El bagazo de caña de azúcar como
substrato para la producción de Pleurotus ostreatus en Jalisco. Revista Mexicana de Micología.
3:79-82.
Guzmán, G., Mata G., Salmones, D., Soto-Velasco, C. y Guzmán-Dávalos, L. (1993). El cultivo de los
hongos comestibles. Ed. Instituto Politécnico Nacional, México. 258 pp.
López-Rodríguez, C., R. Hernández-Corredor, C. Suárez-Franco y M. barrero. (2008). Evaluación del
crecimiento y producción de Pleurotus ostreatus sobre diferentes residuos agroindustriales del
departamento de Cundinamarca. Universitas Scientiarum, 13, 2:128-137.
NMX-FF-109-SCFI-2007. Humus de Lombriz (lombricomposta).Especificaciones y métodos de prueba.
Resh, H. M. (2006). Cultivos hidropónicos. Ed. Mundi-Presa, España. 558pp.
SAGARPA,
SIAP.
(2013).
Infografía
Agroalimentaria
del
estado
de
Morelos.
http://www.siap.sagarpa.gob.mx/infografias/Flippbooks/mor2013/index.html. Recuperado el 10 de
enero 2014.
Sánchez, A., Esqueda, M., Gaitán-Hernández, R., Cordova, A. y Coronado, M. (2008). Uso potencial del
rastrojo de tomate como sustrato para el cultivo de Pleurotus spp. Revista Mexicana de
Micología, 18:17-24.
Sánchez, J. E., Martínez, D., Mata, G. y Leal, H. 2007. El cultivo de setas Pleurotus spp en México. El
Colegio de la Frontera Sur. México. 236pp.
Sánchez, J. y Royse, D. 2001. La biología y el cultivo de Pleurotus spp. Ed. Limusa, México. 294pp.
SEMARNAT. (2013). Programa Nacional para la Prevención y Gestión Integral de los Residuos 20092012. 174pp.
Servicio de Información Agroalimentaria y Pesquera, SIAP. (2012). www.siap.org.mx. Recuperado el 7
de julio de 2012.
Vega, A., Caballero, R., García, J. y Mori, N. (2005). Bioconversion of agroindustrial residues by
Pleurotus ostreatus cultivation. Revista Mexicana de Micología. 20:33-38
976
CHAPTER 8.6. HYDROCARBON ELIMINATION FROM RESIDUAL SLUDGE
OBTAINED IN A WASH AUTOMOTIVE CENTER
Verónica Lavadores (1); Sergio Esteban Vigueras Carmona (1);
Graciano Calva Calva (2); Josefina Pérez Vargas* (1)
(1) TESE, División de Ingeniería Bioquímica, Biotechnology Lab. Edo. Méx, México.
(2) CINVESTAV- IPN, Biotehcnology and Bioengineenig Department, Metabolic Engineering Lab, México
DF, México.
ABSTRACT
Removal of hydrocarbon contaminants in wastewater from automotive washing center
are a problem of environmental pollution, along with the consumption of water employed
for washing the car from 50 to 1000 liters per car depending on the requested service.
Because of the irregularity of the centers of automotive wash is now a requirement that
prior to discharge water to drain to be treated water. Some of these centers have
implemented the use of small plants wastewater treatment, however, still have the
problem of disposal of sewage sludge generated in the wastewater treatment plant. It
was determined that these sludges have a high hydrocarbon content therefore it was
proposed treated with free-living nitrogen fixing bacteria (BFNA) that degrading
petroleum hydrocarbons. It has been demonstrated that cultures of BFNA isolated from
contaminated soils can be successfully removed 120 grams of oil per kilogram of
contaminated soil, so that it was expected that the use of these crops BFNA favored
elimination hydrocarbons present in the sludge obtained from treatment water in car
wash. In studies it has been found that over a period of 8.5 days the amount of total
solids is decreased as much as 88% of an initial quantity of 1396 ppm total solids. The
residual sludge had black color before treatment in the bubble column fermenter
inoculated with a culture of BFNA within 48 hours yellowing without turbidity, and total
hydrocarbon amount decreased to 75% was obtained.
Key words: Hydrocarbons Removal, sewage sludge, BFNA, wastewater, automotive
center, total solids.
Introducción
El desarrollo de pequeñas industrias como los servicios de autolavado, genera miles
de metros cúbicos de agua contaminada por algunos derivados del petróleo como
gasolina, aceite automotriz, limpiadores, líquidos refrigerantes, líquidos de frenos,
anticongelantes, disolvente de limpieza de piezas, diesel, gasolina, desengrasantes,
aceites y grasas lubricantes, champú, siliconas o ceras entre otros, como resultado del
lavado de carrocerías y motores. El agua un recurso natural no renovable, la mayoría
de estas pequeñas empresas han tomado conciencia de la necesidad de reutilizar el
agua, además es un requisito el uso del agua tratada para el servicio de autolavado.
977
El promedio de agua que se consume en los centros de autolavado por vehículo
reportado que es clasificada de acuerdo al tamaño y procedimiento de lavado (Fall
2007) el gasto de agua por vehículo es aproximadamente de 1.9 a 5.6 l/min, Faris
(2012) reportó en un gasto promedio de agua 880 l para camiones, mientras que
Pazéus (1996) reporto un promedio de 1200 l. En la Ciudad de México la cantidad de
agua tratada que utilizan estos centros de autolavado son muy elevadas pues se
estima que mensualmente se gastan entre 1,000,000 m 3 , si por cada pipa de 10 m3 de
agua tratada (INEGI, 2009).
En los servicios de autolavado se utilizan pequeñas plantas de tratamiento en
particular los lodos residuales utilizados en este trabajo son obtenidos después de un
proceso de floculación. Esta agua residual contiene una gran cantidad de sólidos
totales y hidrocarburos y componentes típicos del agua como grasas jabón etc. En la
eliminación de sustancias orgánicas presentes en los lodos residuales (grasas, aceites
e hidrocarburos) ha sido estudiada para encontrar microorganismos que utilizan estos
compuestos como fuente de carbono y energía en donde el producto final sea la
producción de CO2, asegurando la eliminación del compuesto contaminante (Nannipieri
et al 2001). Entre los géneros de microorganismos degradadores de hidrocarburos se
encuentran bacterias, algas, levaduras y los hongos filamentosos. Los más importantes
que se han sido aislado en ambientes acuáticos son: Pseudomonas, Arthrobacter,
Micrococcus, Nocardia, Vibrio, Acinetobacter, Brevibacterium, Corynebacterium,
Flavobacterium, Candida Rhodotorula y Sporobolomyces. Entre los hongos se
encuentran; Penicillium, Cunninghamella, Verticilium, Beauveria, Mortieriella, Phoma,
Scolebasidium (Pérez et al, 2004). Se han estudiado hongos que degradan el 0,4% de
lodos en 3 semanas y por cultivo bacteriano en dos semanas (Spencer, 2004).
También se ha estudiado cultivos BFNA degradadores de hidrocarburos en condiciones
de fijación de nitrógeno (Pérez et al, 2000, Pérez et al 2004) estos cultivos también han
mostrado la capacidad de producir biosurfactantes utlizando hidrocarburos como fuente
de carbono (Pérez et al, 2010), en un sistema de tratamiento de bioaumentación con
biopilas los cultivos de BFNA remueven de 120 g de petróleo crudo por kg de suelo con
una eficiencia del 90 % en un lapso de un año. La propuesta del uso de BFNA fue que
en un se ha probado 20000 ppm de hidrocarburos donde los cultivos de BFNA
remueven un 95% en 15 días (Pérez et al, 2006). Así que se propuso utilizar estos
cultivos ya que en medio líquido y en suelos son eficientes para eliminar altas
concentraciones de hidrocarburos.
Materiales y Métodos
Microorganismo. El cultivo de BFNA aislado de suelo contaminado por hidrocarburos
del estado de Tabasco, fue sembrado por estría en placa en medio Rennie modificado
(1987), se incubó a 27°C, durante cinco días. El cultivo obtenido fue inoculado en
medio liquido mineral usando queroseno como única fuente de carbono, en cultivo
batch, incubado a 27°C, 180 rpm en un agitador orbital incubado durante 72 h. El
cultivo fue utilizado como inóculo después de haber centrifugado a 5000 rpm durante
978
15 min, el paquete celular fué resuspendido con 30 ml de solución salina isotónica
estéril obteniendo una DO de 1 medida a 620 nm.
Lodo residual. Se muestrearon 20 l de aguas residuales de un reactor de
sedimentación después del proceso fisicoquímico de un tren de tratamiento de diez
ciclos en la planta de tratamiento del servicio de autolavado, ubicado en Av. Anillo
Periférico Blvd. Adolfo López Cortines 40000-Bis. Álvaro Obregón. Col. Jardines del
Pedregal (Toyota), en donde se da mantenimiento, lavado de autos, motores y
carrocerías. Los recipientes, muestras y almacenamiento de las mismas se realizaron
de acuerdo a la norma (NMX-AA- 003 -1980). La muestra se guardó en refrigeración
por 48 h, el lodo sedimentado obtenido fue utilizado para realizar los estudios del
presente trabajo.
Caracterización de los lodos del agua residual de un servicio de autolavado. Para
la caracterización de los lodos se determinaron los siguientes parámetros de Densidad
de lodo, pH, sólidos totales, conductividad, grasas y aceites, coliformes y huevos de
helminto, grasas, aceites e hidrocarburos de acuerdo a las normas correspondientes
(NOM- 003-SCFI-1980, Chávez, 2012)
Determinación de
CO2 producido. La medición de CO2 producido por los
microorganismos, durante la incubación de los lodos residuales en un sistema cerrado
se realizó por volumetría de acuerdo al método propuesto por Bartha y Pramer (1965).
Se tomaron muestras de 4 ml de la cámara de reacción con la solución de KOH 0.01
N, al tiempo cero y cada 24 horas hasta los 7 días.
Resultados y Discusión
En la Tabla 1 se presentan los resultados obtenidos para la caracterización del lodo
residual obtenido del agua tratada del centro de automotriz de Toyota. Se observa que
el agua tiene una carga alta de microorganismos por lo que el agua se trató con cloro
activo para ser reutilizada en el centro de lavado. Se realizaron los análisis
microbiológicos nuevamente y disminuyó la carga de coliformes y los microorganismos
totales hasta no detectarlos, el agua obtenida fue reutilizada en la planta para el lavado
de autos. Los lodos obtenidos del proceso de sedimentación en frio fueron estos fueron
tratados en un reactor de columna de burbujeo el cual fue mantenido a 30 oC en un
periodo de 48 horas se obtuvo la disminución del color negro hasta un color amarillo
translúcido para que al final del tratamiento de 7 días este quedara incoloro.
El cultivo de las BFNA permanece activo durante 180 horas en donde empieza a
decaer el crecimiento del cultivo de BFNA, lo que también se ve reflejado en la grasas y
aceites que contiene el lodo, lo que nos dice el cultivo de BFNA los utiliza como fuente
de carbono, y aun cuando aún queda sustrato el cultivo no lo utiliza posiblemente
debido a que su naturaleza química y que el cultivo ya no tiene la capacidad metabólica
para utilizarlos como fuente de carbono. 979
TABLA 1. Caracterización física, química y biológica de los lodos residuales
Parámetros
Sólidos totales (ppm)
Lodos
residuales
(Planta
Toyota)
1396
Límites
normativos
de aguas
de reuso
500
pH
Conductividad (µS)
7
1245
6-7
na
Temperatura
Peso volumétrico (
g/cm3)
Grasas y aceites (ppm)
Color
27°C
0,992764
18- 25°C
0.99
5468.8
Negro aceitoso
na
Sin color
Proteína (mg/l)
255
x
x
x
x
na
na
Jabón (%)
0.64
olor
Aceite,
gasolina
na
Crecimiento
crecimiento
crecimiento
crecimiento
SC
No detectados
SC
Microbiológico
Bacterias totales
Hongos
Coliformes totales
Bacterias fijadoras
de nitrógeno
Huevos de Helminto
Notas: nc, no cuantificada; na, no aplica; sc, sin crecimiento
En la Tabla 2 se muestran la eficiencia obtenida de los parámetros físicos,
químicos y biológicos que se determinaron a los lodos residuales, antes y después del
tratamiento biológico con la BFNA en condiciones de fijación de nitrógeno en
biorreactor. El cultivo de acuerdo a los resultados obtenidos tiene una gran eficiencia
con los lodos residuales lo que nos permite continuar con los estudios de remoción
para proponer el cultivo de BFNA adicional al tratamiento químico en la planta de
servicio automotriz.
En la tabla los resultados obtenidos con respecto a los sólidos totales fue
favorable y nos permite sugerir este tratamiento biológico para poder disponer de estos
sólidos sin ningún tratamiento posterior disminuyendo el costo de la disposición de los
residuos por algún proceso de confinación o tratamiento químico posterior. En el caso
980
de la cuantificación de microorganismos totales se notó una disminución considerable
de la flora microbiana debido posiblemente a que se favoreció el crecimiento de BFNA
por el medio de cultivo mineral en donde no se adiciona fuente de nitrógeno. Al
cuantificar los cultivos de BFNA se observa un crecimiento hasta 100 horas y después
se encuentra que la fase estacionaria termina a las 80 horas. En este lapso de tiempo
se cuantificaron los hidrocarburos y se encontró que hay 60 gramos de hidrocarburos
por litro de lodo residual en donde después de 180 horas se obtiene un 80 % de
eliminación de los hidrocarburos totales por lo que la eficiencia del tratamiento permite
proponer la utilización del reactor para la eliminación biológica de los hidrocarburos,
aunque aún se están realizando experimentos para poder implementar el sistema en la
planta de tratamiento.
TABLA 2. Eficiencia de remoción de la BFNA en el ensayo experimental.
Determinación
Proteínas (g/l)
Microorganismos totales
(ln UFC/ml)
ST(ppm)
Remoción de grasas y aceites
de hidrocarburos (ppm).
pH
Temperatura
Inicial
255
19.67
Final
30.66
2.35
Eficiencia
88%
88 %
858
5468.8
142
2728.5
83%
50%
7
28°C
7
28°C
Sin variación
Sin variación
Conclusión
El objetivo del trabajo era disminuir la concentración de sólidos totales para alcanzar
los valores de la normatividad y se obtuvo una disminución del 88% con lo que se
alcanzan los límites exigidos por la norma para aguas tratadas.
Se obtuvo una eficiencia de remoción de grasas y aceites del 50%, asi como una
disminución de la carga microbiana por lo que los lodos residuales obtenidos pueden
dispuestos sin ningún tratamiento posterior.
Los hidrocarburos totales disminuyeron en 80 % con el cultivo de BFNA por lo que
este cultivo puede ser propuesto para utilizarlo en procesos de remoción de
hidrocarburos en aguas residuales de centros automotrices, se continua realizando
pruebas para el sistema de tratamiento.
Agradecimientos
Se agradece al TESE por el apoyo al trabajo así como a la SEP por el apoyo otorgado
para la realización del proyecto.
981
Referencias
Bartha R, Pramer D (1965). Features of flask and method for measurement of the persistence and
biological effects of pesticides in soil. Soil Sci. 100(1):68-70.
Chávez Pazos Yazmín. (2012). Caracterización fisicoquímica de un efluente de auto-lavado para un
tratamiento de coagulación-floculación. Tesis Lic. Ingeniero Químico. Tecnológico de Estudios
Superiores de Ecatepec.
Fall, C., Lopez. V.M., Jiménez . M. K., Díaz. D.C García. P.D Y Chavez L.M. (2007), Carwash
wastewaters: Characteristics, Volúmen And Treatability by Gravity oil Separation, Revista
Mexicana de Ingeniería Química, 6(2), 175-184.
Faris H. A. (2012). Treatment of oily wastewater produced from old processing plant of north oil
company. Journal of Engineering Sciences. 19(1), 23-34.
Matos. P. R. y Bidois D. E. (2009). Evaluation of the biodegradation of different types of lubricant oils in
liquid medium. Brazilian Archives of Biology and Technology. 52(5).
Nannipieri. P., Ascher J., Ceccherini M., Landi L, Pietramellara. M., Renella G. (2001). Microbial diversity
and soil functions. European Journal Soil Science. 54(4): 655-670.
NMX-AA-003-SCFI-1980 Secretaria de comercio y fomento industrial norma mexicana aguas residuales
muestreo. http://www.conagua.gob.mx/CONAGUA07/Noticias/NMX-‐AA-‐003-‐1980.pdf.
Pérez V. J., Calva .C. G. y Suárez S. J. (2004). Degradación de hidrocarburos por microorganismos.
Tecnocultura (07): 33-35.
Pérez. V. J., Poggi. V. A., Calva. C. G., Ríos. L. E., Rodríguez. V.R., Ferrera. C. R. y García. E.F. (2000).
Nitrogen fixing bacteria capable of utilising kerosene hydrocarbons as a sole carbon source.
Water. Science and Technology. 42(5-6): 407 -410.
Pérez Vargas J., Anaya Reza O., Chang Solis C.K., Membrillo Venegas I., Calva Calva G., (2010),
Biosurfactant Production by Nitrogen Free Living Fixing Bacteria Grown in Hydrocarbons, XV
International Scientific Congress CNIC'2010, National Center for Scientific Research, On line at:
http://www.congresocniccuba.com.
Pérez-Vargas J., Palma-Cruz F., Castañeda López C., García Rivero M., Martínez Trujillo A., CalvaCalva G., (2006), Microbial Degradation of Aromatic Hydrocarbons by Atmospheric Nitrogen
fixing bacteria Isolated from Contaminated Soil of Tabasco, The Second International Meeting on
Environmental Biotechnology and Engineering (2IMEBE), 26-29 Septembre, Poggi-Varaldo H.M.,
Ríos-Leal E., García-Mena J., Esparza-García F., Ponce-Noyola M.T., Robles-González I.,
Sastre-Conde I., Macarie H., Sanz J.L., Watson-Craik I., Foresti E., Reible D., Garibay-Orijel C.,
(Eds.), CD, CINVESTAV, Mexico.
Rennie R. J. (1987). To single for medium the isolation of acetylene reducing (dinitrogen-fixing) bacterium
from soils. Ca. J. Microbiol. 27(1) : 8-14.
982
CHAPTER 8.7. APPLICATION OF A BIOLOGICAL TREATMENT, AFTER FENTON
TREATMENT, IN A SOLID WASTE MATURE LEACHATE
Marbella Amador Cruz (1); Patricia Eugenia Cruz Ortega (2); Rosa Salgado Brito (3);
Gabriel Pineda Flores* (2)
(1)
(2)
(3)
Professional Interdisciplinary Biotechnology Unit-IPN, México DF, México
Mexican Cleaner Production Centre- IPN, México DF, México
Simon Bolivar University, México DF, México
ABSTRACT
The Solid Waste Mature Leachate (SWML) is a liquid waste with a high level of organic
matter. It is formed from the solid waste deposited in confinement sites in longer periods
to 10 years. In the ecosystems, the SWML affects the balance of nutrients, produce
genotoxic effects on local biota and contributes to soil erosion. Chemical methods have
been applied to SWML, as Fenton oxidation, which are effective in removing recalcitrant
organic waste. However, it is possible to increase the removal of these compounds
using a method of biological oxidation after the Fenton oxidation. The objective of this
work was to apply a treatment by activated sludge after Fenton oxidation to a sample of
SWML to increase the removal of organic matter.
It was determined the BOD, COD, total nitrogen, total phosphorus and an UV
visible scan of a SWML sample in order to characterize it. The Fenton method was
applied with a ratio of 0.027 Fe+2/H2O2. Further Fenton treatment the activated sludge
treatment was applied, using a glass column bioreactor with bubbling and an inoculum
of 10% of activated sludge at ambient temperature. The reduction of organic matter was
determined by assessing the COD every 24 hours for 5 days.
The Fenton treatment reduces BOD, COD and total nitrogen 49.06, 91.53 and
73.52% respectively, and total phosphorus in an undetectable level. After applying the
biological treatment the COD is reduced to 99.37%; the BOD and total nitrogen in an
undetectable level. The result of the UV visible scan shows a change in the region
sweeping trend from 100 to 200 nm; this is interpreted as a modification of the
compounds with aromatic structures in the SWML after applying the treatment.
The Fenton treatment reduces total phosphorus to a level allowed in the Mexican
official standard levels, and after applying the biological treatment, BOD and total
nitrogen are also reduced to permissible levels, so we conclude that the sequential
application of these two treatment methods reduce effectively the organic pollutant
content in the SWML.
Key words: mature leachate, solid waste, Fenton treatment, biological treatment.
983
Introduction
The Solid Waste Mature Leachate (SWML) is a liquid waste with a high level of organic
matter. It is formed from the solid waste deposited in confinement sites in longer periods
to 10 years (Deng and Englehardt 2006). In the ecosystems, the SWML affects the
balance of nutrients, produce genotoxic effects on local biota and contributes to soil
erosion (Noaksson et al. 2005, Mor et al. 2006, Sang et al. 2006). Chemical methods
have been applied to SWML, as Fenton oxidation, which are effective in removing
recalcitrant organic waste (Zhang et al. 2006). However, it is possible to increase the
removal of these compounds using a method of biological oxidation after the Fenton
oxidation (Zhang et al. 2005). The objective of this work was to apply a treatment by
activated sludge after Fenton oxidation to a sample of SWML to increase the removal of
organic matter.
Preparation and characterization of solid waste mature leachate. It was performed
a visit to the Bordo Poniente landfill for leachate collection. The stage I was visited and
OLWUHVRIOHDFKDWHLGHQWLILHGDV³PDWXUH´ZHUHFROOHFWHGZLWKWKHVXSSRUWRIVWDIIILOOLQJ
The leachate was placed in plastic containers with tight-fitting lid 2 L capacity. It was
transported to the laboratory in a container with dry ice, and it was refrigerated in the
laboratory at 4 °C until use.
In order to characterize the leachate, it was determined the pH, Chemical Oxygen
Demand (COD), Biochemical Oxygen Demand (BOD), total nitrogen and total
phosphorus before the biological and Fenton treatment, and a spectrophotometrically
scan with a range of 150 to 700 nm was applied. HACH 25651-15 kit, which uses the
sample digestion reactor DRB-200 mark HACH, was used to determine COD. To
determine total nitrogen and total phosphorus, they were used HR kits 1-100 and 10072
HR mark HACH respectively. The BOD was determined according to the method
described in the standard NMX-AA-028-SCFI-2001; dissolved oxygen was measured
according to the volumetric method described in the standard NMX-AA-012-SCFI-2001.
Fenton treatment applied on the mature leachate of urban solid waste. The Fenton
treatment was applied with a combination of hydrogen peroxide 30% plus FeCl 3, the
employed ratio was 0.027 Fe+2/H2O2. Beakers of 500 ml were used to add the leachate
plus the volume of FeCl3 to be tested. With the help of a calibrated burette it was added
the H2O2 30%, then residual peroxide was measured. The COD, total nitrogen and total
phosphorus were evaluated using the methods indicated in previous sub-section.
Biological treatment. This treatment was applied to the leachate previously treated by
Fenton method. The activated sludge treatment was applied, using a glass column
bioreactor with bubbling and an inoculum of 10% of activated sludge at ambient
984
temperature. The reduction of organic matter was determined by assessing the COD
and total nitrogen every 24 hours for 5 days.
Table 1 shows the results of the characterization of the leachate before applying the
proposed treatments.
TABLE 1. Initial characterization of the mature leachate from urban solid waste
Parameter
COD
pH
BOD5
Total nitrogen
Total phosphorus
Concentration
(mg/L, except pH)
4156.15
8.14
398.51
1533.46
131
Figure 1 shows the spectrophotometric scan obtained from the mature leachate of
urban solid waste without treatment.
Figures 2 y 3 show the concentration of COD and total nitrogen respectively, after
Fenton treatment in different concentrations of FeCl3 and H2O2. The Fenton treatment
reduces COD and total nitrogen 91.53 and 73.52% respectively. The results of the
reduction of total phosphorus applying seven concentrations of FeCl 3 and H2O2, were in
all events: total phosphorus in undetectable level.
985
Absorbance
7
6
5
4
3
2
1
0
150
200
250
300
350
400
450
500
550
600
650
700
750
nanometers
FIGURE 1. Spectrophotometric scan of the mature leachate without the Fenton
and Biological treatments
COD 800
(mg/l) 700
600
500
400
300
200
100
0
01:02
01:03
01:04
01:05
01:06
01:08
01:10
Fe+2:H2O2 relationship
FIGURE 2. COD reduction after applying Fenton treatment in different
concentrations of FeCl3 and H2O2
986
700
Total N (mg/l) 600
500
400
300
200
100
0
01:02
01:03
01:04
01:05
01:06
01:08
01:10
Fe+2:H2O2 relationship
FIGURE 3. Total nitrogen reduction after applying Fenton treatment in different
concentrations of FeCl3 and H2O2
Figure 4 shows COD reduction applying biological treatment. After applying the
biological treatment the COD is reduced to 99.37%; the BOD and total nitrogen were in
undetectable level.
Figure 5 shows the spectra of absorption of the mature leachate of solid waste,
before and after receiving the Fenton and activated sludge treatments.The result of the
UV visible scan shows a change in the region sweeping trend from 100 to 200 nm; this
is interpreted as a modification of the compounds with aromatic structures in the SWML
after applying the treatment.
987
COD 350
(mg/l)
300
250
200
150
100
50
0
0
1
2
3
4
5
Time (days)
FIGURE 4. COD reduction applying activated sludge treatment to a mature
leachate of urban solid waste previously treated with Fenton method.
4
3.5
Absorbance
3
2.5
SWML1 after treatment Lixiviado oxidado
2
Deionized water Agua
1.5
Lixiviado puro
SWML Leachate without treatment 1
1Solid Waste Mature Leachate 0.5
0
0
200
400
600
800
Wavelength (nm)
FIGURE 5. Spectra of absorption of the mature leachate of solid waste
Conclusion
The Fenton treatment reduces total phosphorus to a level allowed in the Mexican official
standard levels, and after applying the biological treatment, BOD and total nitrogen are
also reduced to permissible levels, so we conclude that the sequential application of
these two treatment methods reduce effectively the organic pollutant content in the
SWML.
988
References
Deng Y. and Englehardt J.D. (2006). Treatment of landfill leachate by the Fenton process, Water
Research, 40, 3683-3694.
Mor S., Ravindra K., Dahiya R.P. and Chandra A. (2006). Leachate characterization and assessment of
groundwater pollution near municipal solid waste landfill site, Environ. Monitoring Assessment, 118,
435-456.
Noaksson E., Linderoth M., Tjärnlund U. and Balk L. (2005). Toxicological effects and reproductive
impairments in female perch (Perca fluviatilis) exposed to leachate from Swedish refuse dumps,
Aquatic Toxicol., 75, 162-177.
Sang N., Li G. and Xin X. (2006). Municipal landfill leachate induces cytogenetic damage in root tips of
Hordeum vulgare, Ecotoxicol. and Environ. Safety, 63, 469-473.
Zhang H., Choi J.H. and Huang C.P. (2005). Optimization of Fenton process for the treatment of landfill
leachate, J. Hazardous Materials B, 125, 166-174.
Zhang H., Zhang D. and Zhou J. (2006). Removal of COD from landfill leachate by electro-Fenton
method, J. Hazardous Materials B, 135, 106-111.
989
CHAPTER 8.8. IMPACT OF THE SPREADING OF URBAN WASTE ON
AGRICULTURAL SOIL BACTERIAL COMMUNITIES IN THE PERIPHERY OF
OUAGADOUGOU, BURKINA FASO
Edmond Hien* (1); Sabine Favre-Bonté (2); Salomon Bouda (1);
Aboubakar Sidiki Ouattara (1); Dominique Masse (3); Sabine Houot (4);
Sylvie Nazaret (2)
1
Université de Ouagadougou, UFR /SVT, Ouagadougou, Burkina Faso.
UMR CNRS 5557 Ecologie microbienne, UCB LYON1, Villeurbanne, France
3
LMI-IESOL, IRD Campus Bel-Air, Dakar ± Sénégal
4
UMR EGC Equipe Sol, INRA Thiverval-Grignon, France
E-mail : [email protected]
2
ABSTRACT
Recycling organic wastes is of interest in agriculture and horticulture for improving soil
TXDOLW\IDYRULQJSODQWGHYHORSPHQW5HFHQWO\XVXDOIDUPHUV¶SUDFWLFHVLQ%XUNLQD)DVR
involved the use

Environmental Biotechnology and Engineering – 2014

Transcripción

Documentos relacionados

3 Stories of You

Un Siglo de la Industria Aeronáutica Española One Century of

Senhores - Bass fishing trips at Amazing Lake El Salto Mexico by

trabajo sobre la evolución del traje en españa work on the evolution