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RECENT TRENDS IN BIOTECHNOLOGY

BIOREMEDIATION ADVANCES IN RESEARCH AND APPLICATIONS

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RECENT TRENDS IN BIOTECHNOLOGY

BIOREMEDIATION ADVANCES IN RESEARCH AND APPLICATIONS

MOHAMMED KUDDUS EDITOR


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Library of Congress Cataloging-in-Publication Data ISBN: H%RRN

Published by Nova Science Publishers, Inc. † New York


This book is dedicated to my family………



CONTENTS Preface

ix

Chapter 1

Microbes in Bioremediation Irfan Ahmad

Chapter 2

Biodegradation and Detoxification of Environmental Recalcitrant Compounds Jyotsna K. Peter and Sushma Ahlawat

Chapter 3

Chapter 4

Chapter 5

Chapter 6

Chapter 7

The Significance of Microbial Cell Surface Energy in Wastewater Bioremediation Meryem Asri, Alae Elabed, Soumya Elabed, Saad Ibnsouda Koraichi and Naïma El Ghachtouli

1

19

49

The Use of Green Microalgal Cultures for Bioremediation of Freshwater Environments Polluted with Chromium, Nickel and Cadmium Laszlo Fodorpataki, Sebastian R. C. Plugaru, Katalin Molnar, Peter Marossy, Bernat Tompa and Szabolcs Barna

71

Advances in Microbial Degradation of Substituted Phenols with Special Reference to Actinomycetes Namita Panigrahi, Kannan Pakshirajan and Naresh K. Sahoo

113

The Potential of Biosorption Techniques in Heavy Metals Mitigation Akhilesh Bind, Sushma Ahlawat, Veeru Prakash and Somya Agarwal Azo Dye Removal Technologies Maulin P. Shah


151

169

viii Chapter 8

Chapter 9

Contents Plant Growth Promoting Bacteria in Heavy Metals Bioremediation Nezha Tahri Joutey, Nabil Tirry, Wifak Bahafid, Hanane Sayel and Naïma El Ghachtouli Ex Situ Stimulated Bioremediation of a Soil Contaminated with Oil Pollutants: The Dynamics and the Efficiency of Biodegradation of Saturated and Aromatic Hydrocarbons Tatjana Šolević Knudsen, Mila Ilić, Jelena Milić, Gordana Gojgić-Cvijović, Srđan Miletić, Vladimir Beškoski and Miroslav M. Vrvić

185

211

Chapter 10

The Bioremediation of Wastes from the Seafood Industry Saima and Mohammed Kuddus

239

Chapter 11

Strategies for Plastic Waste Management Shikha Raghuwanshi and Reeta Goel

259

Chapter 12

The Role of Cyanobacteria in the Bioremediation for Restoring Aquatic Ecosystems Vinod Rishi, Ravindra Singh and A. K. Awasthi

295

The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) Yu. N. Vodyanitskii and A. T. Savichev

327

Chapter 13

Chapter 14

Biosensors in Bioremediation Ghazala Yunus and Mohammed Kuddus

351

About the Editor

365

Index

367


PREFACE Human activity and increasing urbanization along with irresponsible disposal of garbage and industrial waste is disturbing natural environment continuously. The conventional treatment methods of environmental pollution have recognizable drawbacks. However, the emerging technology of bioremediation offers an alternative method to clean the environment. Bioremediation is the engineered process of biodegradation in which biological means (microorganisms and bioactive compounds) are used to degrade harmful materials under controlled conditions. Bioremediation could provide an efficient solution for pollution-free environment. This book covers fourteen chapters including scientific progress and recent technologies in biodegradation to find an eco-friendly solution for environmental pollution. In the first chapter (Microbes in bioremediation by Irfan Ahmad), authors highlighted potential applications of microbes in bioremediation. Bioremediation is a technique to remove the contaminant from a contaminated site by using microbes. Microbes are so small to contact the contaminants easily and they have ideally possessed all the enzymes that allow them to digest the environmental contaminants as food. This chapter explains how and what types of microorganisms involve in bioremediation process. Second chapter titled ‘Biodegradation and detoxification of environmental recalcitrant compounds’ by Peter and Ahlawat comprises microbial degradation and detoxification of various categories of recalcitrant molecules inclusive of plastics, agrochemicals, dyes and drugs. Also, the source and ecological impact of recalcitrant compounds on biotic and abiotic components are discussed. Similarly, in the third chapter (Significance of microbial cell surface energy in wastewater bioremediation) Asri et al. discussed about various methods commonly used for the determination of surface free energy and crucial effect of surface free energy and its components on wastewater bioremediation. This chapter also presents future research directions for the development of efficient wastewater bioremediation processes. Fourth chapter (Use of green microalgal cultures for bioremediation of freshwater environments polluted with chromium, nickel and


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cadmium), contributed by Fodorpataki and coworkers, is about different tools for developing cost-effective protocols for bioremediation of aquatic ecosystems polluted with different concentrations of chromium, nickel and cadmium ions. In the fifth chapter (Advances in microbial degradation of substituted phenols with special references towards actinomycetes by Panigrahi et al.), authors discussed about contamination levels of substituted phenols in the environment and its toxicity on living system. In addition, biodegradation of substituted phenols by different microorganisms especially actinomycetes in various bioreactor systems for the treatment of contaminated wastewaters has been discussed. Another chapter (sixth in this series) by Bind et al. (Potential of biosorption techniques in heavy metals mitigation) contain literatures about various biosorption techniques used for heavy metal removal from the contaminated site. In the seventh chapter titled ‘Azo dye removal technologies’ by Maulin Shah, author discussed about removal of azo dyes from wastewater by biological treatment. The azo dyes are the largest commercial dyes and account for nearly 75% of all products of textile dyes. Chapter eighth (Plant growth promoting bacteria in heavy metals bioremediation), contributed by Joutey et al., includes different mechanisms of plant growth promotion and of metal remediation by metal-detoxifying Plant Growth Promoting Bacteria as well as the recent progress in exploitation of these bacteria alone or in combination with mycorrhizal fungi to support phytoremediation in the view of environmental restoration. In the ninth chapter (Ex-situ stimulated bioremediation of soil contaminated with oil pollutant-dynamics and the efficiency in biodegradation of saturated and aromatic hydrocarbons by Knudsen et al.), authors studied dynamics and efficiency in biodegradation of saturated and aromatic compounds (constituents of the oil pollutant), during ex-situ bioremediation of the polluted soil. They concluded that the extent of biodegradation in the samples exposed to stimulation was much higher than in the samples where only natural biodegradation occurred, pointing that the stimulated biodegradation is much more efficient process than the natural biodegradation. Another chapter (tenth) titled ‘Bioremediation of wastes from seafood industry’ by Saima and Kuddus focused on the role of chitinase in the bioremediation of wastes from seafood industry. Increased amount of seafood wastes, resulting from the industrial processing of seafood, causes serious problem both for the environment and for the processing plants. Chitin decomposition by conventional methods are harmful for environmental, instead microbial method is used for the degradation of chitinous waste and received increasing attention because of their broad applications in various industry. Chapter eleven contributed by Raghuwanshi and Goel on the topic ‘Strategies for plastic waste management’ described about recalcitrant nature of plastics that possess threat to living creature and environment, and management of plastic waste by using different types of bioremediation techniques. In the twelfth chapter (Role of cyanobacteria in the bioremediation for restoration of aquatic ecosystem by Rishi et al.), authors discussed about pollution of aquatic ecosystem and its bioremediation by various microorganisms.


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They concluded that cyanobacteria can significantly remediate aquatic ecosystem contaminated with heavy metals, domestic and industrial effluents, agro-chemical wastes and phenolic compounds. Another chapter (thirteenth in this series) titled ‘Biodegradation of oil in soil-groundwater under the influence of Fe(III)’ written by Vodyanitskii and Savichev described about biodegradation of oil hydrocarbons under the impact of Fe(III). Finally, the last chapter (fourteenth) deals with different biosensors used in bioremediation (Biosensor in bioremediation by Yunus and Kuddus). Biosensors are analytical device to detect molecules of our interest by using biological and an electrical component. It is also used for the detection and monitoring of pollutants along with various toxic compounds present in the environment. Recently, much attention has been paid on on-site detection of pollutants by using novel approach. This chapter contains literatures about various biosensors and their application in bioremediation process. Hopefully this book will be useful for environmentalist, waste site managers, scientist and academics along with regular students in research and practice of bioremediation. It will also provide useful information for experts in allied fields including ecology, biochemistry, enzymology, environmental microbiology and biotechnology. Due to growing industrialization and irresponsible disposal of hazardous wastes, more and more wastes are being generated every year. The book will help to fuel the research and management activity of environmental pollutants through bioremediation processes. In the last, but not least, I express my deepest sense of gratitude and regards to my family. Words are not sufficient to express my feelings to them for their love and moral support which helped me in completing this book. I would also like to thanks all the authors who have eagerly contributed their chapter in this book. I assured that these chapter will be beneficial for academics and scientific community working in the field of bioremediation. My special gratitude is due to Carra Feagaiga, Manager, Nova Science Publishers, USA, for her efforts to get this book published. Finally, I also express my sincere gratitude to the Publication House for providing this opportunity.

Mohammed Kuddus Hail, KSA February 2018



In: Bioremediation Editor: Mohammed Kuddus

ISBN: 978-1-53613-554-1 © 2018 Nova Science Publishers, Inc.

Chapter 1

MICROBES IN BIOREMEDIATION Irfan Ahmad* Department of Clinical Laboratory Sciences, College of Applied Medical Sciences, King Khalid University, Abha, Saudi Arabia

ABSTRACT Bioremediation is a technique used for the waste management by microbes to remove the contaminant from a contaminated site. Microbes living ubiquitously play an essential role in bioremediation. Microbes have characteristics to contact the contaminants easily, and ideally possessed all the enzymes required to digest the environmental contaminants. Since billions of year microorganisms are involved in bioremediation such as the breakdown of animal, plant and human wastes flourishing the life to continue from one generation to the next generation. Without the activity of microorganisms, the earth would have been full of wastes, and the nutrients as a result the continuation of life would have brought in detritus. Moreover, depending on the type of contamination and the geo-chemical conditions at the site of contamination, microorganisms have the efficiency to destroy the man-made contaminants. This chapter explains how and what types of microorganisms are involve in bioremediation process. In addition, it evaluates the types of contaminants which are most susceptible to bioremediation and delineates the contaminated sites where bioremediation is most feasible to succeed.

Keywords: microbes, bioremediation, pollutants, heavy metals, contaminants

* Corresponding Author Email: [email protected].


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INTRODUCTION Bioremediation is the use of microbes to degrade the environmental contaminants into less toxic forms. When microorganisms are exposed to the organic compounds they start to neutralize it by enzymatic action. The contaminants present in the environment either reduces or induces the function of microbial enzymes. However, only some specific community of microbes has an ability of bioremediation and it depends on the functional and structural groups of toxic compounds. The microorganisms along with the genetic capability to alter the compounds of interest should be found in the metabolism of contaminant in the process of bioremediation. In few cases, bioaugmentation might reduce the lag phase duration. Bioremediation is one of the options that provide the feasibility to decay the various harmful contaminants by using natural biotic activity (Gupta and Mahapatra 2003). Microorganisms have capability to use unlimited pattern of electron donors and electron acceptors for metabolism. However, beside of these oxidation/reduction reactions, microbes have also other approaches to purify the atmosphere. The process of bioremediation employs these principles to gather a more appropriate combination for activity of microbial community, electron oxidation/ reduction/contaminant concentrations and some other practical as well as physical considerations to remediate the targeted pollutant. Bioremediation is frequently more advantageous strategies than conventional strategies since it can be executed in situ (directly at the site of contaminant without transport the contaminated material). A novel in situ process allows biotic treatment of contaminated water by reactive molecules synthesized by microbes. The microbes can be either present in the contaminated site or isolated from other site and transported to the contamination site. In the latter case, this is referred to as bioaugmentation, whereas, if the naturally occurring population is encouraged to proliferate by the addition of extraneous electron donors or electron acceptors, it is called biostimulation. The benefit microorganisms already residing in contaminated environments, they are able to utilize available nutrients and electron pairs ultimately lowering costs and frequently well adapted to environmental conditions. Bioremediation may be divided into two groups, biosorption and bioaccumulation (Figure 1). Biosorption is a passive type adsorption mechanism, which is reversible and rapid (Gadd and White 1993). The metals are reserved by means of physicochemical interaction (e.g., adsorption, ion exchange, crystallization, complexation and precipitation) between the metal and functional groups existing on the surface of a cell (Volesky 2004). Some factors have an effect on the biosorption of metals, such as temperature, pH, biomass concentration, particle size, ionic strength and occurrence of other ions in the solution (Volesky 2004). Both dead and living biomass can take place for biosorption as it is not dependent on cell metabolism. While bioaccumulation consists of both intra and extracellular processes where passive intake plays a limited and not


Microbes in Bioremediation

3

fully defined role (Gadd and White 1993). Hence, living biomass may only take place for bioaccumulation.

Figure 1. Mechanism of biosorption and bioaccumulation by microorganisms.

MICROBES IN BIOREMEDIATION Bioremediation is the process to transform or degrade the toxic compounds into less toxic form with the help of microorganisms (Kumar and Maitra 2016). However, pollutants such as polycyclic aromatic hydrocarbons (PAH) and heavy metals cannot be destroyed easily, consequently, persevere and accumulate in the atmosphere. Their accumulation in the environment is very danger to human health and environments. Recently scientists have focused on bioremediation as a more cost-effective and efficient procedure for the removal of contaminants from the environment as compared to other physical and chemical method(s) of bioremediation (Zhu et al. 2016). Several other advantages of microbial bioremediation require less energy, lesser production of slush and more efficient, etc., (Mukherjee et al., 2017). It is well documented that various microorganisms have a natural power to remove toxic heavy metal ions by biosorption (Singh et al. 2014). Some examples of microorganisms involved in process of bioremediation for pollutants as well as heavy metals are given in Table 1. However, individually microorganisms are not able to mineralize maximum lethal compounds. Whole mineralization outcomes in chronological deprivation by a certain group of microorganisms and includes co-metabolism actions and synergism. Biological communities of microbes in diverse habitats comprise an incredible physiological adaptability, they have an ability to metabolize and frequently mineralize massive amount of organic molecules. Some unknown communities of fungus and bacteria metabolize a


Irfan Ahmad

4

lot of molecules by the activity of microbes in one environmental condition or another. Majority of the bioremediation process is carried out in aerobic conditions, however anaerobic conditions (Colberg and Young 1995) can allow microbes to degrade refractory molecules. Table 1. Microorganisms involved in the bioremediation of pollutants and heavy metals Microorganism Bacteria Enterobacter species Rhizobium meliloti Acetobacterium paludosum Clostridium acetobutylicum Rhodococcus erythroplis Rhizobium sp. Enterobacter sp. Escherichia coli Citrobacter sp. Bacillus subtilis Acinetobacter guillouiae Pseudomonas putida

Pollutants/heavy metals

References

Chlorpyrifos Dibenzothiophene Hexahydro-1,3,5-trinitro-1,3,5-triazine Hexahydro-1,3,5-trinitro-1,3,5-triazine Polychlorinated biphenyl Polychlorinated biphenyl Trichloroethylene Cr Au, Pb  Cu, Ni, U, Hg, Zn Cr Cu Cu, Co, Mn, V, Pb, Ti, Ni

Bacillus cereus Bacillus barbaricus Acinetobacter sp. Providencia vermicola Fungi Agaricus bisporus Pycnoporus sanguineus Aspergillus niger Aspergillus niger and Aspergillus foetidus Penicillium spp. Aspergillus fumigatus Algae Microorganism Cladophora spp. and Spirogyra spp. Rhizoclonium spp. Ascophyllum nodosum Selenastrum capricornutum and Scenedesmus acutus

Cd Cd, Pb F F

Singh et al. 2004 Frassinetti et al. 1998 Sherburn et al. 2005 Zhang and Huges 2003 Chung et al. 1994 Damaj and Ahmad 1996 Kang et al. 2012 Robins et al. 2013 Gunasekaran et al. 2003 Fathima et al. 2010 Majumder et al. 2015 Kamika and Momba 2013 Huang et al. 2014 Sem et al. 2014 Mukherjee et al. 2017 Shraboni et al. 2017

Zn, Cd Polycyclic aromatic hydrocarbons U, Cd, Zn, Ag, Th Fe, Zn, Ni, Mo, V, Mn

Nagy et al. 2014 Arun et al. 2008 Gunasekaran et al. 2003 Anahid et al. 2011

Ar Pb

Loukidou et al. 2003 Ramasamy et al. 2011

Pollutants/heavy metals Cu, Pb

References Lee et al. 2011

As, V Zn, Cd, Pb, Cu Benzo (a) pyrene (BaP)

Saunders, 2012 Romera et al. 2007 Garcia de Llasera et al. 2016



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MECHANISMS OF BIOREMEDIATION Microbes are ubiquitous that dominant in the soil contaminated with heavy metal and may simply transform it into non-toxic forms. Microbes have capability of two-way defense such as synthesis of enzymes for degradation of target pollutants and resistance to appropriate heavy metals. A variety of bioremediation mechanisms are identified comprising biosorption, bioaccumulation, biomineralisation, metal-microbe interactions, bioleaching and biotransformation. Microbes eliminate the heavy metals from the soil by chemical action for their development and growth. They have ability to dissolve metals and oxidizing or reducing transition metals. A variety of approaches by which microorganisms reestablish the environment is immobilizing, binding, oxidizing, transformation and volatizing of heavy metals. Bioremediation may be carried out successfully at a specific site by the microbe tactic, and through the mechanism that controls development and activity of microbes in the contaminated places, their response to environmental changes and their metabolic capabilities. Most of the pollutants are carbon-based solvents that distort the membranes, nonetheless, cells might acquire the resistance mechanisms comprising synthesis of defensive material around the cell membrane (Sikkema et al. 1995). For example, plasmid-encoded and energy dependent metal efflux systems comprising chemiosmotic ion/proton pumps and ATPases are reported for Cd, As and Cr resistance in various bacteria (Roane and Pepper 2000).

Figure 2. Mechanism utilized by the microbial cell for bioremediation.


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Figure 3. Bioremediation of pollutants using biodegradation capabilities of microbes consists of the natural attenuation, though it can be improved by engineered systems, either by adding of certain microorganisms (bioaugmentation) or by biostimulation, wherever nutrients are added. Genetic engineering is also employed to increase the biodegradation abilities of microbes by genetic engineered microorganisms (GEM). Nonetheless, there are numerous factors affecting the efficacy of this procedure and risks related to the usage of GEM in the process.

Microbes and occasionally fungi or plants are employed to deceive or break the contaminants into less harmful and simpler substances such as water and carbon (Figure 2). Microbial bioremediation is mainly two types namely biostimulation and bioaugmentation. For biostimulation, matters such as nutrients and oxygen are added in order to increase the capability of microorganisms that already exist to degrade the pollutants. In the cases where such types of microorganisms do not exist, bioaugmentation must be carried out in which microorganisms are introduced in the specific location where they are not present naturally. Occasionally, these microbes are genetically engineered for the certain purpose of digesting the contaminant (Figure 3).

MICROBIAL ENZYMES IN BIOREMEDIATION Microorganisms release the certain catalytic enzymes to degrade the complex pollutants such as PAHs (Ping et al. 2017). Microbial enzymes also involved in bioremediation process by breaking the chemical bonds and reduce the toxicity of pollutants. Examples of microbial enzymes that involved in bioremediation for pollutants are described below and summarized in Table 2.



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Table 2. Application of microbial enzymes Enzyme Oxidoreductases Oxygenases Monooxygenases Dioxygenases Laccases

Applications Protein engineering, bioremediation, synthetic chemistry Bioremediation, synthetic chemistry Protein engineering, bioremediation, synthetic chemistry Synthetic chemistry, pharmaceutical industry, bioremediation Food industry, paper and pulp industry, textile industry, nanotechnology, synthetic chemistry, bioremediation, cosmetics Food industry, paper and pulp industry, textile industry, pharmaceutical industry, bioremediation

Peroxidases

Oxidoreductases Various bacteria, fungus (Gianfreda et al. 1999) and higher plants (Bollag and Dec 1998) detoxify the toxic organic compounds by oxidative coupling are facilitated with oxidoreductases. Energy is extracted by microbes through a biochemical process facilitated by oxidoreductases biocatalyst to break the chemical bonds as well as to support the relocation of electrons of the reduced organic compound to alternative chemical compound. Through these oxidation-reduction reactions, the pollutants are finally converted to less toxic substances (ITRC 2002). Furthermore, the humification of several phenolic materials, which are formed in the soil environment from the degradation of lignin, occur with the help of oxidoreductases. Similar to this, toxic xenobiotics such as anilinic or phenolic compounds can also be detoxified by oxidoreductases through binding to humic substances or copolymerization, polymerization with other substrates (Park et al. 2006). Microbial enzymes have been exploited in the degradation and decolorization of azo dyes (Williams 1977). Radioactive metals are reduced from an oxidized soluble form to insoluble form by many bacteria. Bacterium receipts the electrons from organic compounds during the energy production process and radioactive metal is used as the final electron acceptor. Radioactive metals are also indirectly reduced with an intermediate electron donor by few of the bacterial species. Conclusively precipitant may be observed within the metal reducing bacteria by the action of redox reactions (Leung 2004). Additionally, chlorinated aromatic organic compounds are one of the plentiful refractory litters material that occur in the discharges produced through the pulp and paper manufacture factories. Partial degradation of lignin produces these compounds during the process of pulp bleaching. Many fungal species have ability to remove the chlorinated aromatic organic compounds from the contaminated surroundings. The activity of fungus is primarily because of the presence of oxidoreductase enzymes (extracellular) such as lignin peroxidase, laccase and manganese peroxidase, which are isolated by the fungal body (mycelium) into their adjacent


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background. Furthermore, fungi could be spread the soil impurities more efficiently as compared to bacteria because of the filamentous structure (Rubiler et al. 2008). The enzymes secreted by the roots of the plants may disinfect pollution of water with phenolic compounds. The plant families such as Solanaceae, Fabaceae and Gramineae release the oxidoreductases enzyme, which involves in the oxidative breakdown of certain soil rudiments. Phytoremediation of organic impurities has been usually based on three major groups of chemical compounds such as hydrocarbon (explosives), petroleum hydrocarbons and chlorinated solvents (Duran and Esposito 2000).

Oxygenases The enzyme oxygenases belong to oxidoreductase group. Oxygenases enzymes are categorized into two groups; monooxygenases as well as dioxygenases, based on the number of oxygen atoms. They involve in the metabolic process of carbon-based compounds by enhancing water solubility or reactivity. It’s having the comprehensive range of substrate and is effective against an extensive compounds range. Normally the overview of oxygen atoms into the organic molecule by oxygenase enzyme results in the break of the aromatic rings. Generally, microbial mono or dioxygenases are the furthermost studied enzymes in bioremediation process (Arora et al. 2009). Halogenated organic compounds consist the major group of environmental contaminants resulting in their widely applicable as fungicides, insecticides, herbicides, plasticizers, hydraulic and intermediary compounds for the synthesis of the chemical. Decomposition of these contaminants is accomplished by particular oxygenases enzyme. Similarly, halogenated ethylenes, ethanes and methanes are dehalogenated by oxygenases (Fetzner and Lingens 1994).

Monooxygenases Monooxygenases include one atomic molecule of oxygen into the substrate. On the basis of the presence of cofactor, monooxygenases are categorized into two subclasses: P450 monooxygenases and flavin-dependent monooxygenases. P450 is a type of heme containing monooxygenases, which occur in prokaryotic as well as in eukaryotic organisms. Flavin dependent monooxygenases contain flavin as a prosthetic group and need NADPH or NADP as coenzyme. Monooxygenases are stereo-selective and region selective and perform as biocatalysts in the bioremediation processes. Most of the monooxygenase enzymes reported earlier are consisting of cofactor, however, some monooxygenases do not need cofactor for their function. Such enzymes need simply oxygen for their accomplishments (Arora et al. 2010). Monooxygenases catalyzed the



9

dehalogenation, desulfurization, denitrification, hydroxylation, ammonification, biodegradation and biotransformation of numerous aliphatic and aromatic compounds. These characteristics have been investigated recently for significant use in biotransformation and biodegradation of aromatic compounds (Arora et al. 2010). Methane monooxygenase is the top most characterized amongst monooxygenases. It is also participated in decomposition of hydrocarbon viz alkanes, cycloalkanes, methanes, alkenes, haloalkenes, aromatic and ethers (Fox et al. 1990). In oxygen-rich environments, monooxygenase carries oxidation of dehalogenation reactions, while in oxygen-less environments, reduction of dechlorination reaction carries out. Dehalogenation produces the labile products due to the oxidation of the substrate, which undergoes chemical decomposition subsequently (Fetzner and Lingens 1994).

Dioxygenases Dioxygenases are multiple component enzymes, which add oxygen molecule into substrate. Mostly aromatic compounds are oxidized by dioxygenases and, consequently, are using in natural bioremediation. Entire family members comprise a couple of proteins for transport of electron foregoing oxygenase constituents. Crystalline naphthalene dioxygenases have assured the occurrence of single nuclear iron and Rieske cluster in all alpha subunit (Dua et al. 2002). Catechol dioxygenases perform a natural scheme to destroy organic compounds in the atmosphere. These enzymes are predominantly occur in the bacteria existing in soil and helped in conversion of aromatic precursor molecules into aliphatic compounds. Intradiol chopping enzyme exploits Fe (III), however, an extradiol chopping enzyme employ Mn(II) and Fe(II) in some situations (Que and Ho 1996).

Laccases Laccases (p-diphenol: dioxygen oxidoreductase) establish a class of multicopper oxidases that formed by various insects, plants, fungus and bacteria which, facilitate the oxidation of a large variety of reduced aromatic and phenolic compounds through simultaneous reduction of oxygen to water molecule (Mai et al. 2000). It is well documented that laccases found in numerous isoenzyme, each of isoenzyme is encoded by a distinct gene (Giardina et al. 1995). However, in certain cases, the genes have been expressed differently which mainly depend on the type of the stimulus (Rezende et al. 2005). The intra and extracellular laccases (which is produced by various microorganisms) are able to catalyze the oxidation of lignins, polyphenols, paradiphenols, aminophenols, aryl diamines, polyamines, and several inorganic ions (Mai


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et al. 2000). It is well documented that laccases involve in the oxidation of methoxy phenolic and phenolic acids as well as in decarboxylation and demethylation. These enzymes are participated in the lignin depolymerization, resulting in a diversity of phenols. This compound is consumed as nutrition source for microbes or repolymerized through laccase (Kim et al. 2002). Amongst the natural enzyme, laccases show a fascinating set of abundant, oxidoreductase, which illustrates potential of proposing excessively probable for bioremediation and biotechnological uses (Gianfreda et al. 1999). The substrate affinity and specificity of laccase may alter with changes in pH. Numerous reagents such as azide, cyanide, halides and hydroxide can inhibit the laccase (Xu 1996). Diverse laccases seem to have different lenience to inhibit by halides, representing differential halide convenience. Productions of laccase depend on the concentration of nitrogen in fungus. More concentrations of nitrogen are generally needed to gain more quantities of laccase. By homologous or heterologous recombinant laccase may also be synthesized (Gianfreda et al. 1999).

Peroxidases Peroxidases are abundant enzymes, which carry the oxidation of lignin and other phenolic compounds by using hydrogen peroxide. Furthermore, peroxidases enzyme could be a haem as well as non-haem proteins molecules. In animals, enzymes participated in various natural activities including hormone regulation and defense mechanism in immune system however in plants, peroxidases are participated in suberin as well as lignin formation, cell elongation, cross linking of cell wall components, auxin metabolism or defense mechanism against pathogens. The hemeperoxidases are categorized into a couple of categories as occur in prokaryotes and eukaryotes. The subsequent categories of peroxidases are subcategorized into three classes based on the comparison of sequence. Class I is an intracellular enzyme comprising plants ascorbate peroxidase, yeast cytochrome-c peroxidase and microbial catalase peroxidases. Moreover, Class II contains the fungal peroxidases like manganese peroxidase (Mnp) and lignin peroxidase (LiP) from Coprinus cinereus peroxidase, Arthromyces ramosus peroxidase (ARP) and Phanerochaete chrysosporium. Class III comprises plant peroxidases like barley, soybean and horseradish. Furthermore, peroxidases enzymes participated in lignifications process and formation of plant cell wall. Non-haem peroxidases develop five liberated families such as alkyl hydroperoxidase, thiol peroxidase, NADH peroxidase, nonhaem haloperoxidase and manganese catalase. Among them, thiol peroxidase is one of the biggest and consisting of two subfamilies like as peroxy redoxins and glutathione peroxidases (Hiner et al. 2002).



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BIOREMEDIATION BY PHYSIO-BIO-CHEMICAL MECHANISM Biosorption is the process that includes higher affinity of a biosorbent to sorbate (metal ions), continued till equilibrium is found between the two components (Das et al. 2008). Saccharomyces cerevisiae used like biosorbent for the elimination of Cd (II) and Zn (II) by mechanism of ion exchange (Chen and Wang 2007). Cunninghamella elegans developed as an auspicious sorbent released out by textile waste (Tigini et al. 2010). Heavy metal decomposition includes energy for cell metabolism. Bioremediation of toxic metal by joint active and passive means is known as bioaccumulation (Brierley 1990). The fungus has arisen as biocatalysts to transform heavy toxic metals into lesser toxic (Pinedo-Rivilla et al. 2009). For example, various fungus such as Stachybotrys sp., Phlebia sp., Allescheriella sp., Botryosphaeria rhodina and Klebsiella oxytoca have potential to bind metal (D’Annibale et al. 2007). It is well known that soil contaminated with Pb (II) might be biodegraded by various fungal species such as Cephalosporium aphidicola and Aspergillus parasitica by the process of biosorption (Akar et al. 2007). Furthermore, Hg resistant fungi such as Verticillum terrestre, Hymenoscyphus ericae and Neocosmospora vasinfecta had potential to transform toxic form of Hg (II) state to a nontoxic form (Kelly et al. 2006). Various pollutants are hydrophobic in nature and seem to be degraded by microorganisms via exudation of biosurfactant as well as by direct association of cell contaminant. Moreover, biosurfactants form complexes with metallic compound by making strong ionic bonds (Thavasi 2011). Bioremediation can involve in both aerobic and anaerobic microbes activities. Aerobic degradation frequently includes involvement of oxygen in the processes facilitated by hydroxylases, oxidative dehalogenases, monooxygenases and dioxygenases. However, anaerobic degradations of pollutants include the early activation reactions followed with the oxidative metabolism that is facilitated by anoxic electron acceptors. Immobilization is a procedure that employed to diminish the mobilization of heavy toxic metals from polluted places via altering the physiochemical form. Moreover, solidification procedure includes partying of chemical compounds at the places of contamination (Evanko and Dzombak 1997). It is well documented that microbes catalyze the heavy toxic metals from the places of contamination by the process of chelation, leaching, redox transformation as well as by methylation of toxic heavy metals. Heavy metallic compounds can certainly not be totally damaged, however, the procedure converts its oxidation form thus they converted less toxic, water soluble and precipitated. Microbes utilize trace elements and heavy metals as electron acceptors or reduce these metals via the process of detoxification that helps in the exclusion of metals present the places of contamination. Furthermore, microbes eliminate heavy metals by such processes that concern with toxic metallic compounds via enzymatic as well as nonenzymatic mechanisms. A couple of mechanisms for resistance development in bacteria includes detoxification that is a conversion of toxic metal form, which makes it not


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available and release the toxic metal from cells by active efflux pump (Silver 1996). The basic redox reaction carries out in the soil between microbes and toxic metallic compounds; microbes oxidize the heavy metallic compounds and trigger them to lose the electrons that are received by alternative electron receiver (ferric oxides, nitrate and sulphate). Oxygen receives the electron in aerobic conditions, while microorganisms oxidize organic pollutants by reducing electron acceptors in anaerobic conditions. Microbes receive energy for development by oxidation of carbon-based compound with Fe (III) or Mn (IV). More availability of Fe (III) stimulates the anaerobic degradation of organic contaminants for reduction of microbes. The metals are used, as electron acceptors are known as dissimilatory reduction of metal. Biodegradation of chlorines from pollutants carries out by dechlorination reduction, in which pollutants as the chlorinated form of solvents performs as electron acceptors in breathing. Many microbes diminish the metallic form and alter their solubility, as Geobacter species decrease the soluble uranium (U6+) to the insoluble state of uranium (U4+) (Lovley et al. 1991).

BIOREMEDIATION PROCESS BY MOLECULAR MECHANISMS Many molecular mechanisms are known to remove the heavy metals by microbes. Deinococcus geothermalis, a genetically modified bacterium, reduction of mercury (Hg) has been pointed out at elevated temperatures via the expression of mer operon with Escherichia coli encoded for reduction of Hg2+ (Brim et al. 2003). Cupriavidus metallidurans strain MSR33, a mercury resistant bacterium was genetically modified by inserting a pTP6 plasmid, which offered genes (merG and merB) controlling the biodegradation of Hg along with the production of mercuric reductase (MerA) and organomercurial lyase protein (MerB) (Rojas et al. 2011). Pseudomonas strain becomes resistant to mercury by the modification of plasmid pMR68 with novel (mer) genes (Sone et al. 2013). There are two different types of mechanisms for degradation of Hg by Klebsiella pneumonia M426 is Hg volatilization through the reduction of Hg (II) to Hg (0) state (Essa et al. 2002). Genetic modification of radiation resistant bacterium, Deinococcus radiodurans reduces Cr (IV) to Cr (III) has been made for ample degradation of toluene with cloned genes of xyl and tod operons of Pseudomonas putida (Brim et al. 2006). Bacterial metabolites such as metal bound coenzymes and siderophores primarily found in the degradative pathway (Penny et al. 2010).

CONCLUSION Toxicity due to pollution is hazardous for environment and human health in all over the world. Bioremediation is an alternate approach to degrade the waste from the



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environment. Removal of toxic compounds from the atmosphere with the help of microbes has numerous advantages including eco-friendly, adaptability, reproducibility, recycling of bioproducts and much more. Also, bioremediation by microbes is a costeffective technology as compared to other conventional method (such as incineration) for the removal of contaminants from the environment. Microbes release the various enzymes, which break chemical bond of the pollutants and make it less toxic. The bioremediation techniques work to control the naturally occurring microbial catabolic properties for degradation, transformation as well as accumulation of variety of pollutants such as polychlorinated biphenyls (PCBs), polyaromatic hydrocarbons (PAHs), radioactive compounds and heavy metals. Genetic manipulation of the bacterial strain to over express metal transports and insert of metal resistant genes into bacteria revealed that these modifications could significantly change the resistance of microbes to heavy metals. These metal resistant microbes are able to increase absorption and accumulation of pollutants through ecofriendly bioremediation techniques.

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Kumar, V. and Maitra, S. S. (2016). Biodegradation of endocrine disruptor dibutyl phthalate (DBP) by a newly isolated Methylobacillus sp. V29b and the DBP degradation pathway. 3 Biotech., 6(2): doi: 10.1007/s13205-016-0524-5. Lee, Y. C. and Chang, S. P. (2011). The biosorption of heavy metals from aqueous solution by Spirogyra and Cladophora filamentous macroalgae. Bioresource Technology, 102(9): 5297-5304. Leung, M. (2004). Bioremediation: techniques for cleaning up a mess. J. Biotechnol., 21(2):8-22. Loukidou, M. X., Matis, K. A., Zouboulis, A. I. and Liakopoulou Kyriakidou, M. (2003) Removal of As(V) from wastewaters by chemically modified fungal biomass. Water Research., 37(18): 4544-4552. Lovley, D. R., Philips, E. J., Gorby, Y. A. and Landa, E. R. (1991). Microbial reduction of uranium. Nature, 350: 413-416. Mai, C., Schormann, W., Milstein, O. and Huttermann, A. (2000). Enhanced stability of laccase in the presence of phenolic compounds. Appl. Microbiol. Biotechnol., 54(4): 510-514. Majumder, S., Gangadhar, G., Raghuvanshi, S. and Gupta, S. (2015). A comprehensive study on the behavior of a novel bacterial strain Acinetobacter guillouiae for bioremediation of divalent copper. Bioprocess Biosyst Eng., 38(9):1749-1760. Mukherjee, S., Yadav, V., Mondal, M., Banerjee, S. and Halder, G. (2017). Characterization of fluoride-resistant bacterium Acinetobacter sp. RH5 towards assessment of its water defluoridation capability. Appl. Water Sci., 7(4): 1923-1930. Nagy, B., Manzatu, C., Maicaneanu, A., Indolean, C., Lucian. B.T. and Majdik, C. (2014). Linear and nonlinear regression analysis for heavy metals removal using Agaricus bisporus macrofungus. Arab. J. Chem., 10: doi: 10.1016/j.arabjc. 2014.03.004. Park, J. W., Park, B. K. and Kim, J. E. (2006). Remediation of soil contaminated with 2,4-dichlorophenol by treatment of minced shepherd’s purse roots. Arch. Environ. Contam. Toxicol., 50(2): 191-195. Penny, C., Vuilleumier, S. and Bringel, F. (2010). Microbial degradation of tetrachloromethane: Mechanisms and perspectives for bioremediation. FEMS Microbiol. Ecol., 74(2): 257-275. Pinedo-Rivilla, C., Aleu, J. and Collado, I. G. (2009). Pollutants biodegradation by fungi. Curr. Org. Chem., 13(12): 1194-1214. Ping, L. F., Guo, Q., Chen, X. Y., Yuan, X. L., Zhang, C. R. and Zhao, H. (2017). Biodegradation of pyrene and benzo[a]pyrene in the liquid matrix and soil by a newly identified Raoultella planticola strain. 3 Biotech., 7: doi: 10.1007/s13205-0170704-y. Que, L., Ho and R. Y. N. (1996). Dioxygenactivation by enzymes with mononuclear nonheme iron active sites. Chem. Rev., 96(7): 2607-2624.



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Ramasamy, R. K., Congeevaram, S. and Thamaraiselvi, K. (2011). Evaluation of isolated fungal strain from e-waste recycling facility for effective sorption of toxic heavy metal Pb (II) ions and fungal protein molecular characterization: A Mycoremediation approach. Asian J. Exp. Biol. Sci., 2(2): 342-347. Rezende, M. I., Barbosa, A. M., Vasconcelos, A. F. D., Haddad, R. and Dekker, R. F. H. (2005). Growth and production of laccases by the ligninolytic fungi, Pleurotus ostreatus and Botryosphaeria rhodina, cultured on basal medium containing the herbicide, Scepter (imazaquin). J. Basic Microbiol., 45(6): 460-469. Roane, T. M., Rensing, C., Pepper, I. L. and Maier, R. M. (2000). Microorganisms and metal pollution. Environmental Microbiology: second edition, Academic Press: London, UK, p. 55. Robins, K. J., Hooks, D. O., Rehm, B. H. A. and Ackerley, D. F. (2013). Escherichia coli NemA is an efficient chromate reductase that can be biologically immobilized to provide a cell free system for remediation of hexavalent chromium. PLoS One, 8(3): doi:10.1371/journal.pone. 0059200. Rojas, L. A., Yanez, C., Gonzalez, M., Lobos, S., Smalla, K. and Seeger, M. (2011). Characterization of the metabolically modified heavy metal-resistant Cupriavidus metallidurans strain MSR33 generated  for mercury bioremediation. PLoS One, 6(3): e17555. https://doi.org/10.1371/journal.pone.0017555. Romera, E., Gonzalez, F., Ballester, A., Blazques, M. I. and Munoz, J. A. (2007). Comparative study of biosorption of heavy metals using different types of algae. Bioresour. Technol., 98 (17): 3344-3353. Rubilar, O., Diez, M. C. and Gianfreda, L. (2008). Transformation of chlorinated phenolic compounds by white rot fungi. Crit. Rev. Environ. Sci. Technol., 38(4): 227268. Saunders, R. J., Paul, N. A., Hu, Y. and de Nys, R. (2012) Sustainable sources of biomass for bioremediation of heavy metals in wastewater derived from coal-fired power generation. PLoS One, 7(5): doi: 10.1371/journal.pone.0036470. Sem, S. K., Raut, S., Dora, T. K. and Mohapatra, P. K. (2014). Contribution of hot spring bacterial consortium in cadmium and lead bioremediation through quadratic programming model. J Hazard Mater, 265:47-60. Sherburne, L., Shrout, J. and Alvarez, P. (2005). Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) degradation by Acetobacterium paludosum. Biodegradation, 16:539-547. Shraboni, M., Priyanka, S. and Gopinath, H. (2017). Microbial remediation of fluoridecontaminated water via a novel bacterium Providencia vermicola (KX926492). J. Environ. Manage., 204:413-423. Sikkema, J., de Bont, J. A. and Poolman, B. (1995). Mechanisms of membrane toxicity of hydrocarbons. Microbiol. Rev., 59(2): 201-222. Silver, S. (1996). Bacterial heavy metal resistance: New surprises. Annu. Rev. Microbiol., 50: 753-789.


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Singh, B. K., Walker, A., Morgan, J. A. and Wright, D. J. (2004). Biodegradation of chlorpyrifos by enterobacter strain B-14 and its use in bioremediation of contaminated soils. Appl. Environ. Microbiol., 70:4855-4863. Sone, Y., Mochizuki, Y., Koizawa, K., Nakamura, R., Pan-Hou, H., Itoh, T. and Kiyono, M. (2013). Mercurial-resistance determinants in Pseudomonas strain K-62 plasmid pMR68. AMB Express, 3:41 doi: 10.1186/2191-0855-3-41. Thavasi, R. (2011). Microbial biosurfactants: From an environment application point of view. J. Bioremed. Biodegrad., 2:104e. doi:10.4172/2155-6199. Tigini, V., Prigione, V., Giansanti, P., Mangiavillano, A., Pannocchia, A. and Varese, G.C. (2010). Fungal biosorption, an innovative treatment for the decolourisation and detoxification of textile effluents. Water, 2: 550-565. Valls, M., Atrian, S., de Lorenzo, V. and La, F. (2000). Engineering a mouse metallothionein on the cell surface of Ralstonia eutropha CH34 for immobilization of heavy metals in soil. Nat. Biotechnol., 18(6): 661-665. Volesky, B. Sorption and biosorption. Quebec: BV Sorbex, Inc.; 2004. Williams, P. P. (1977). Metabolism of synthetic organic pesticides by anaerobic microorganisms. Residue Rev., 66: 63-135. Xu, F. (1996). Catalysis of novel enzymatic iodide oxidation by fungal laccase. Appl. Biochem. Biotechnol., 59(3): 221-230. Zhang, C. and Hughes, J. B. (2003). Biodegradation pathways of hexahydro-1,3,5trinitro-1,3,5-triazine (RDX) by Clostridium acetobutylicum cell-free extract. Chemosphere, 50:665-671. Zhu, X. Z., Ni, X., Waigi, M. G., Liu, J., Sun, K. and Gao, Y. Z. (2016). Biodegradation of mixed PAHs by PAH-degrading endophytic bacteria. Inter. J. Env. Res. Pub. Heal., 13(8): doi: 10.3390/ijerph13080805.




Chapter 2

BIODEGRADATION AND DETOXIFICATION OF ENVIRONMENTAL RECALCITRANT COMPOUNDS Jyotsna K. Peter1 and Sushma Ahlawat2,* 1

Department of Industrial Microbiology, Jacob Institute of Biotechnology and Bioengineering, Sam Higginbottom University of Agriculture, Technology and Sciences, Allahabad, India 2 Department of Biochemistry and Biochemical Engineering, Jacob Institute of Biotechnology and Bioengineering, Sam Higginbottom University of Agriculture, Technology and Sciences, Allahabad, India

ABSTRACT This chapter provides a perspective on microbial degradation and detoxification of various categories of recalcitrant molecules inclusive of plastics, agrochemicals, dyes and drugs. An overview of the ecological impact of recalcitrants on biotic and abiotic components and their sources are provided. Further detailed descriptions on strategies for bioremediation are presented focusing on microbial degradation and detoxification mechanisms, including analysis.

Keywords: degradation, detoxification, agrochemicals, dyes, biomedical waste.

*

recalcitrant,

bioremediation,

Corresponding Author Email: [email protected].


plastics,

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Jyotsna K. Peter and Sushma Ahlawat

INTRODUCTION Growing population on the Earth is bringing urbanization which is likely to increase the use of recalcitrant compounds. The society is day-by-day introduced with discharge of these lesser or more toxic compounds released in the environment. Among various hazardous compounds present in the system, many are easily degradable whilst some of them are hard to degrade; they are often termed as xenobiotic compounds and few are recalcitrant molecules for which very scarce option for degradation could be defined. They are usually non degradable compounds and microorganisms do not have proper cell machinery to digest these molecules for certain reasons as these may be very large molecules, toxic, or lack of enzyme system to recognize it as substrates. Agro-waste is one of the suitable sources of energy and fuel generation, its proper management may lead to increase in biofuel generation. But in some parts of India a practice of residue burning is opted promptly, this could have been an opportunity for the availability of cattle feed as forage or raw material for biofuel generation (Lohan et al., 2018). Plastic is one of the most recalcitrant and persistent substance found on the Earth. The global plastic production has taken a leap of 1.5 million tons per year in 1950s to 288 million tons per year by 2012 (Bayer et al., 2014). Bio plastics are carbon based biopolymers and an appropriate option to replace synthetic plastics (Bharti and Shwetha, 2016). Olivera et al., (2001) genetically engineered Pseudomonas sp. to manufacture bio plastics. Biopolysters polyhydroxyalkonates (PHA) are now a day’s suitable material for bioplastic preparation (Chen, 2009). Some important recalcitrant compounds are discussed in the following paragraphs.

PLASTICS Plastics are cross linked polymers that have an empirical formula of CnH2n (Sangale et al., 2012). The chemical bonds of the interlinked monomers provide high tensile strength to the polymer and subsequently they can be remolded in various forms of varied thickness. The name plastics has come into existence from ‘Plastikos’ which means capable of remolding into varied shapes (Joel, 1995; Kale et al., 2015). Monomer of plastics includes carbon, hydrogen, silicon, oxygen, chloride and nitrogen (Sangale et al., 2012). A polythene has 64% of total plastic, which is a linear hydrocarbon polymers consisting of long chains of ethylene monomers. Polythene terephthalate (PET) is used for food packaging materials. These are usually recycled through landfill or pyrolyzed. High density polythene (HDPE) is long linear polymer chain often used for manufacturing of bottles for milk, detergents, oil containers, toys etc. and it has great strength. It could be decomposed through pyrolysis. Low density polythene (LDPE) is


Biodegradation and Detoxification of Environmental Recalcitrant Compounds

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highly branched from their carbon backbone in comparison to HDPE. This kind of plastic is highly water resistant and are used as plastic bags, trash bags etc. Polyvinyl chloride (PVC) has a composition of 57% chlorine and 43% carbon. The incorporated chlorine provides a thermal resistance to it. It is generally used for cable insulation, medical devices, blood bags, boots, credit cards etc. (Sharuddin et al., 2016).

Mechanisms of Plastic Degradation Breakdown of polymers into smaller units of oligomers or monomers is depolymerization of the plastic (Frazer, 1994; Hamilton et al., 1995). Pyrolysis is one of the mechanisms of plastic polymer degradation that includes cascade of thermal and chemical reactions which leads to depolymerization of the plastic under oxygen free condition. This process leads to generation of plastic pyrolyzed oil, which is a mixture of carbon composition C5-C20 and C10-C30 (Wongkhorsub and Chindaprasert, 2013). This oil after distillation operation could be used for diesel engine. Mineralization also occurs to the plastic polymers with bioconversion of oligomers or monomers into carbon dioxide, water or methane (Frazer, 1994; Hamilton et al., 1995).

Nature of Pretreatments Physical and chemical pretreatments are the two options for the modification of plastic substrate. A range of physical forces active on the degradation of synthetic polymers of plastics under in vivo conditions of polycarboxylates (Winursito and Matsumura, 1996), poly (ethylene terephthalate) (Heidary and Gordon, 1994), polylactic acids and its co polymers (Hiltunen et al., 1997; Nakayama et al., 1985), poly (α-glutamic acids) and polydimethylsiloxanes, or silicones (Lehmann et al., 1996; Xu et al., 1998). The following pretreatments are more commonly applied: 1) Cut into pieces or powdered: Polythene or LDPE plastics are usually cut into pieces prior to experimentation for sake of ease of experiment. Powdered forms of plastics are also used under in vitro investigation. Later it is decontaminated using ethanol and then dried overnight in hot air oven (Das and Kumar, 2015). 2) UV irradiation: Exposure of plastic piece to UV radiation up to few months (Broshkevitch et al., 2013) leads to reduced degradation time of the polymer. 3) Thermal treatment: Treating the experimental plastic to high temperature alone or in combination of the UV irradiation enhances the degradability much fairly. Broshkevitch et al., (2013) used chemical, UV and thermal pretreatments for the degradation of HDPE plastic followed by microbial degradation by


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Jyotsna K. Peter and Sushma Ahlawat Phanerochaete chrysosporium, Pseudomonas putida and Sphingomonas macrogoltabidus. 4) Alkali treatment: Small strips of polythene are soaked in tween, bleach, and distilled water followed by rinse in ethanol then drying for overnight (Talkad et al., 2014).

Routes of Plastic Degradation Degradation of plastic is accomplished via two means i.e., aerobic or anaerobic modes. Aerobic decay of plastic leaches the evolution of carbon dioxide and water (Ivanova et al., 2013). Anaerobic degradation of plastics takes place at landfill sites and sediments that ultimately leads to release of carbon dioxide, water and methane.

Modes of Plastic Degradation a) Oxo-biodegradation: This is a two way treatment of plastic by UV rays followed by its oxidation. Both the processes sufficiently reduces the molecular weight of polymer of plastic and leads to biodegradation (www.willowridgeplastics.com). UV light breaks down end products of plastics and oxidation changes the molecular weight of the polymer. b) Photodegradation: The process involves the irradiation of plastic by ultraviolet radiations (UV) at a range of 290 to 400nm. This range of wavelength of UV radiation is engrossed by the plastic which tends to cleave bond and depolymerization resulting in photo decay (Dilara and Briassoulis, 2000). The photodegradation assay also leads to ketone photolysis which occurs through two routes known as Norrish I and Norrish II (Klemchuk, 1990). Yashchuk et al., (2012) reviewed the biotic and abiotic factors on the composition and mechanical performance of the polythene with oxo-degradable additives. c) Thermal degradation: Polymer degradation accomplished via thermal degradation is achieved at elevated temperatures at which rate of various reaction increases leading to indirect degradation (Albertsson et al., 1992; Dilara and Briassoulis, 2000). It is a kind of molecular degradation as a consequence of overheating of the polymer. At elevated temperatures the long chain backbone of polymer tends to separate and start reacting with each other which leads to change in the properties of the polymer compounds. This kind of chemical reaction alters the physical entity of the material. Thermal degradation of plastic polymers often includes changes of the following categories: reduced ductility, chalking, color changes, cracking etc. (Olayan et al., 1996).



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d) Oxidation: Oxidation of the polymer takes place if subsequent exposure of plastic to heat on time scale leads to breakdown of plastic. Once a polymer is oxidized it shows significant reduction in molecular weight (Yashchuk et al., 2012). e) Biodegradation: It involves the microbial cells to degrade the polymer. The process can be operated aerobically or anaerobically depending upon the type of microorganisms associated with the utilization of the polymer as a sole carbon source or this could be a mixed process involving both aerobic and anaerobic biodegradation approach simultaneously or alternatively. Plastics are biodegraded aerobically in natural form and anaerobically in sediments and landfills. In case of compost plastic polymers tend to be degraded through mixed aerobic and anaerobic type of degradation. Plastics are partly faced to aerobic degradation and partly anaerobic during composting process in soil. During biodegradation process these plastics are decomposed to monomers (Swift, 1997). f) Plastic degrading microorganisms: Microorganisms are the biological agents with great capacity to decompose plastics and polythenes. Nature harbors countless assortments of plastic degrading microorganisms of varying potential to utilize this recalcitrant compound within their cellular metabolism as sole carbon source utilization. The cellular machinery of these microbial entities makes them adept to allow degradation of these plastics. A plethora of miscellaneous biochemical reactions proceed in order to degrade the substrate. Range of external and internal factors governs the rate of microbial degradation process. The cases may differ under in vitro and in vivo conditions. Consequences of microbial degradation of plastic polymers leads to variation (Shah et al., 2008) in their molecular weights. The complexity of the polymers determines the degrading capacity of the microorganisms towards the process (monomers > dimers > oligomers > polymers, branched or complexed polymers); the higher is the complexity of polymers greater would be the degradation potential required. Microbial enzymes (extra or intracellular enzymes system) are the cellular machinery that plays quintessential role in the process of utilization or degradation of plastic polymers (Doi, 1990; Gu et al., 2000b).

Extrinsic and Intrinsic Factors Governing Microbial Degradation of Plastics Plastic polymers attacked by the microorganisms suffer changes in mobility, crystallinity, molecular weight, functional group types, substituents present in structure, plasticizers or additives added to polymers (Gu et al., 2000b). Polymers of plastic are decomposed by quite a few microorganisms involving bacteria and fungi (Gu et al., 2000a). Each category of microorganisms renders the changes in plastics different.



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Fungal detriment causes small scale swelling and bursting due to pressure exerted by growing hyphae and mycelium of fungi into the polymer solid (Griffin, 1980). Microbial attack to plastic polymers is enzymatically degraded wherein the enzyme cleaves monomeric bonds and releases off the methane, water or carbon dioxide molecules. Biosurfactants could be an alternative of microbial degradation of plastics. Mathew and Vimala (2016) performed a setup of in vitro biodegradation of plastics using bacterial cells along with its biosurfactant (surfactin) in mineral salt medium.

Enzymatic Action The depolymerization of poly (caprolactone) is assisted by microbial enzymes (Shah et al., 2008; Ivanova et al., 2013) thereafter the depolymerized monomers are absorbed and degraded by microbial live cells (Goldberg, 19995). Lignin degrading enzymes of microbial origin are a class that potentially degrades plastic. Theses enzymes include laccasses, manganese peroxidase, lignin peroxidases and alkane hydroxylases, ureases, esterases, proteases, polyster hydrolases, chitinase, carboxylesterases etc.

Conventional Methods of Plastic Disposal (Land Filling and Incineration) Plastic waste disposal through landfill is a conventional approach that allows burial of the plastics in a confined area in covered soil. It may remain buried there for a long duration of time. It remains an alternative for burning strategies of the solid waste. The purpose of the landfill is to allow natural biodegradation process. As a consequence of the landfill degradation process gases are emitted (usually methane) are released into the environment and increases global warming. Soluble forms or inorganic forms leach out into the deeper soil and even to aquatic water bodies that increases further toxic pollutants (heavy metals, pesticide chemicals etc) in the environment. Plastic land filling eventually leads to deposition of various toxic secondary environmental pollution viz. benzene, toluene, xylene etc. (Ivanova et al., 2013). Incineration is an optional method but does not solve the problem of plastic disposal rather it is useful for remolding the plastic into new shapes and caste and can be reused for insulation, road construction or other way. Alike landfill approach of plastic disposal incineration removes the hurdle of space requirement but likely to release several gases into the atmosphere (Ivanova et al., 2013).

Methods Adopted for Estimation and Evaluation of Plastic Biodegradation a) Visual observations: Examination of plastic for visual changes as a record of degradation effects is one of the easiest way to observe plastic degradation. The



b)

c)

d)

e)

f)

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effects involve roughening of surfaces, development of holes or cracks, defragmentation, colour alteration, biofilm disposition at plastic surface. It is a qualitative examination more than quantification and also indicates primary degradation approach by microorganisms. To determine these visual observations variety of techniques and instrumentation are available such as AFM for surface analysis, FTIR, DSC, NMR, X-ray Diffraction etc. (Shah et al., 2008). Weight loss measurements: The assay involves determination of loss of polymer mass under treatment. Powdered polymer disintegration is better quantified through the technique using sieving analysis to determine the residual polymer through extraction technique to separate out the residual polymer from the microbial biomass, soil, or compost. Tracking changes in mechanical properties and molar mass: Variations in the mechanical characteristics of the polymer can be examined such as tensile strength that even indicates changes in molar mass. Abiotic degradative processes often lead to changes in mechanical properties of the polymer more significantly in comparison to biotic degradative mode. Estimation of gaseous exchange or evolution: Consumption of oxygen and release of carbon dioxide on utilization of sole carbon sources (plastic, polymer) is significant indicator of polymer degradation rate. To estimate the carbon dioxide release an easy conventional method is approachable i.e., trapping of carbon dioxide in Ba(OH)2 solution tailed by titration. To determine the concentration of oxygen or carbon dioxide in the test container infrared or paramagnetic oxygen detectors are used. Infra-red gas analyzer along with titration system is an air sampling technique for non-continuous aerated closed systems. Carbon dioxide analysis is also applied for solid matrices such as compost and soil. Radiolabelling: It involves the estimation of 14CO2 evolution which is nondestructive and precised method. Net CO2 and 14CO2 evolution magnitudes estimation are naive, non-destructive and quantify crucial biodegradation. If appropriately 14C labelled test material is accessible, their magnitudes estimation and interpretations are moderately candid. Biodegradability examinations using this technique for polymeric resources in different microbial environments show an elevated degree of meticulousness and uniformity (Sharabi and Bartha, 1993). However, labeled materials are high-priced and not at all times obtainable. The licensing and the waste disposal complications associated with radioactive work might likewise be a shortcoming. Examination of halo zone formation: It is a simple technique to screen microorganisms for degradation of polymers in artificial medium under in vitro conditions. The assay is performed in solidified agar plates which are opaque prior to inoculation and formation of halo zones around the colonies is an


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Jyotsna K. Peter and Sushma Ahlawat indication of utilization of the polymer. It is semi quantitative analyses. Low molecular weight hydrocarbons can be degraded by microbes (Nakamura et al., 2005). They are taken in by microbial cells, ‘activated’ by attachment to coenzyme-A, and converted to cellular metabolites within the microbial cell. However, these processes do not function well in an extracellular environment and the plastic molecules are too large to enter the cell. This problem does not arise with natural molecules, such as starch and cellulose, because conversions to low molecular weight components by enzyme reactions occur outside the microbial cell (Chandra and Rustogi, 1998, Pometto et al., 1993, Saroja et al., 2000). g) Enzymatic degradation: Application of pure enzymes or microbial cells as source of enzymes causes the enzymatic degradation of polymer (Ivanova et al., 2013). Microorganisms are the source of endoenzymes and exoenzymes that are capable to cleave of the bonds of varying monomers of the polythene or plastic polymer thus varying the molecular weight change of the polymer. First the polymer is bound by the enzyme and later the enzyme causes catalytic cleavage of polymers into their monomers or oligomers, the process known as depolymerization (Ivanova et al., 2013). Microbial action on plastic degradation is influenced by varied intrinsic and extrinsic factors such as availability of water, temperature, oxygen usage, minerals, pH, redox potential, and carbon and energy source influence the growth of microorganisms (Holmes, 1988; Sand, 2003). Chemical and physical attribute (surface area, hydrophilic, and hydrophobicity, molecular weight, chemical structure, melting temperature, crystallinity) (Tokiwa et al., 2009), Side chain variability of polymer, Degree of crystallinity (Sharma et al., 2015). h) Controlled composting assay: Solid wastes treated in anaerobic digesters for recycling or obtaining valuable transformation is of commercial value option. The method is beneficial for biodegradable components. During composting there are changes in the physico-chemical and physiological factors such as temperature, switch over of aerobic to anaerobic or vice versa, dominant micro flora at different duration of composting (aerobic, anaerobic, mesophiles, thermophiles etc).

Alternatives of Plastic Reuse Ceppa and Marino (2012) has presented alternative management of material used in the food packaging industries. The research highlighted the raw materials used as the packing material of food processed from certain industries in Italy. The type of raw materials used in food industries were jute, nylon, low density polythene (LDPE), high



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density polythene (HDPE), card board (PAP), paper based polyaliminate (CA) and polypropylene (PP). Jute can be used as such for plant nurseries. Plastics first segregated into their respective types, shredded and molded into various forms that could be transformed as sacks and stretching films, technical support fabrics, ropes, flocked yarns, brittle brooms, strapping seals, trolley wheels, polyamide seals, plugs and caps for aerosols cans, bottles, containers, helmets (PPE: personal protective equipment), big bags and LDPE or HDPE components for automotive sector. In another recent research conducted by Appiah et al., (2017) who proposed management of plastics by blending waste thermoplastic polymers viz, HDPE and PP. The samples were first shredded and blended with bitumen in situ. Basic rheological characteristics of the polymer modified bitumen (PMB) were evaluated such as viscosity, softening and penetration values. It was observed that the properties of unmodified bitumen were enhanced in their rheological properties. The most compatible and incompatible blends for HDPE were observed at 2 and 3% polymer loading respectively. The highly compatible homogeneous blend was obtained with PP at 3% polymer loading. The study concluded that the use of waste commodity plastics in binder modifications has an advantage of a cheap and effective means of enhancing conventional bitumen binder performance characteristics and is an alternative way to utilize plastic waste.

Biofuel Generation from Plastic Wastes Conversion of plastic waste into fuel is one of the smart actions towards the Earth’s health sustenance. The molten plastic waste is pyrolyzed in a reactor and subjected to condense into fuel gases or forwarded for fractional distillation for separation of heavy oil, diesel oil, gasoline and gas. Raja and Murali (2011) refabricated fuel from waste plastics whose fractional distillation revealed 60% gasoline and 30% diesel oil. There are several other ways of generating fuels. Lohan and Sharma (2012) and Lohan et al., (2015) has depicted the potential strategies to generate electricity using solar energy, implementation of gasifires, or biogas production utilizing biomass y-products from the environment that are considered waste keeping in view of the energy requirement of the Jammu and Kashmir regions of India. The Ladhakh valley of India receives 320 days of solar energy that could reduce the use of LPG in these places. Apart from the aforesaid options of energy and fuel generation biodiesel production is one of the great opportunities of to replace the use of conventional fuels though more studies and research. Lohan et al., (2013) has stated the use of petroleum as fuel source in India which has become the fourth largest primary consumer of petroleum in the world. Production of biodiesel from agro wastes is one of the potent options (Maurya et al., 2012; Maurya et al., 2015). Maurya et al., (2015) gave their insights on utilization of



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lignocellulosic biomass for generation of biofuels. It is a recent approach to generate biofuels from plastic wastes.

Bioplastics Bioplastics are plastics of biological origin (Chen, 2014). Bioplastics can be prepared via different plant sources such as Tapioca, Potato (Reddy et al., 2013), starch based materials (Lorcks, 1998), corn starch (Reddy et al., 2013) etc. Okoshi et al., (2016) described the flame retardancy of bio based plastics and developed inedible woody biomass plastic. Nano aluminum hydroxide was used as a flame retardant for polyolefin. PHB (polysters of hydroxyl alkonates; Bharti and Shwetha, 2016) is again a novel source of bioplastic preparation of microbial origin (Reddy et al., 2013). The most fascinating attributes of bioplastic is its ecofriendly nature. There are several advantages of use of bioplastic over conventional plastics viz. bioplastics possess lower carbon foot prints; it consumes lower energy costs for manufacturing, its biodegradable under natural conditions. Bacteria are suitable PHB producers viz. Alcaligens eutrophus, Bacillus megaterium, Methylobacterium rhodasianum, Pseudomonas putida and Sphaerotilus natans (Bharti and Shwetha, 2016).

AGROCHEMICALS Agrochemicals are the sources of vivid chemicals in the form of fertilizers, pesticides and antibiotics applied to the plant or soil for obtaining higher yields of crops. These chemicals at elevated concentrations in soil and water lead to serious health concerns leading to a strong urge of science and technologies to find the remedial processes for the existing problem. Implementing phytoremediation (plants and algae), microbioremediation (microorganisms), derivative remediation (biomolecule assisted bioremediation), rhizoremediation (enzymatic digestion and rhizospheric bioremediation) are hopes of the solution to the pollution and damage to lives and the Earth ecosystem. Quite a few groups of microorganisms are able enough of biodegrading the agrochemicals such as bacteria (bacterial remediation), fungi (myco-remediation), algae (phycoremediation).

Phytoremediation The strategy of phytoremediation is to involve plants. The mechanisms concerned in the phytoremediation success comprise numerous bioremediation approaches like



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phytoextraction, rhizodegradation, rhizofiltration, phytodegradation, phytostabilization (EPA, 2006, Dietz and Schnoor, 2001). Many are the elements which influence phytoremediation effectiveness. The enzymes implicated in plant biotransformation are chiefly CYP, carboxylesterases, GST and translocases (Dietz and Schnoor, 2001). Maize (Zea mays) and giant foxtail (Setaria faberi) can biotransform particular herbicides (Hatton et al., 1999). Another crop plants such as brinjal, spinach, radish and rice are able to bioaccumulate pesticides including DDT and benzene hexachloride (Mishra et al., 2009; Velazquez-Fernandez et al., 2012). Also, Basil is potent to bioremediate endosulfan from soil (Ramirez- Sandoval et al., 2011). Barley (Hordeum vulgare) is capable of translocating herbicide metolachlor into vacuoles (Martinoia et al., 1993). Horseweed (Conyza canadensis) can sequester glyphosate in vacuoles (Ge et al., 2010). Likewise, it has been recommended that genetic engineering could be employed to improve phytoremediation capacities of poplars (Yadav et al., 2010) and plants in general (Abhilash et al., 2009). In supplement, few plants can deliver co-substrates and oxygen to rhizosphere microorganisms, intensifying them to biodegrade pesticides. Phytostimulation has evidenced to be one of the supreme valuable strategies because it brings together the bioremediation competencies of plant and its biorizhosphere -bacteria and mycorhiza (Chaudhary et al., 2005; Velazquez-Fernandez et al., 2012).

Rhizoremediation The literature suggested to use rhizoremediation process for remediation of insecticide such as parathion and the herbicide 2,4-dichlorophenoxyacetic (Kuiper et al., 2004; Velazquez-Fernandez et al., 2012). Also, pea (Pisum sativum) can stimulate endophytic bacteria that can help in degradation of 2,4-D (Germaine et al., 2006).

Mycoremediation The fungal peroxidases, dioxygenases and oxidases are able to degrade pesticides more competently than cytochrome P450 (Aust, 1995). Some other biotransformation enzymes like lignin peroxidase, laccases, and dichlorohydroquinone dioxygenase are also produced by fungi such as Phanerochaete chrysosporium, Pleurotus ostreatus, Ganoderma australe and Fusarium ventricosum; the three former are ligninolytic, and the latter is a saprobe. Another fungus P. chrysosporium can degrade endosulfan probably due to intracellular peroxidase (Kullman and Matsumaura, 1996; Velazquez-Fernandez et al., 2012). F. ventricosum has also been reported to disintegrate endosulfan (Siddique et al., 2003). It has been shown that fungal peroxidases and dioxygenases are also involved in biodegradation of pentachlorophenol (Ruttiman-Johnson and Lamar, 1996; Sun et al.,



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2011). The ligninolytic fungus Ganoderma australe, isolated from the stone pine (Pinus pinea), is a suitable for biodegradation of lindane (Rigas et al., 2007).

Phycoremediation In the phycoremediation, an unicellular green alga (Chlorella fusca var vacuolata) is reported to transform a herbicide Metfluorazon by a CYP (Thies et al., 1996). Recently, it is also described that an alga namly Chlamydomonas reinhardtii can bioaccumulate and biodecompose herbicide prometryne (Jin et al., 2012; Velazquez-Fernandez et al., 2012).

Bacterial Remediation Involvement of bacteria for pesticides remediation is of immense importance. Anabaena sp. (cyanobacterium), Pseudomonas spinosa, Pseudomonas aeruginosa and Burkholderia are well reported bacteria to degrade endosulfan (Lee et al., 2003; Hussain et al., 2007). Another research has been reported that Pseudomonas, isolated from endosulfan polluted soil, is potentially degrade endosulphan (Velazquez-Fernandez et al., 2012). Microorganisms of the genera Pseudomonas, Bacillus, Trichoderma, Aerobacter, Muchor, Micrococcus and Burkholderia have been shown to degrade dieldrin and endrin (Matsumoto et al., 2009). The literature suggested that some bacteria could disintegrate parathion and fenitothrion using esterases enzyme activity (Choi et al., 2009; Kim et al., 2009). An Enterobacter related strain from soil have ability to mineralize chlorpyrifos, parathion, diazinon, coumaphos and isazofos (Singh et al., 2004). Equally, a bacterial isolate related to Serratia can decay diazinon (Abo-Amer, 2011).

Enzymes For degradation of these agrochemical molecules by means of microbial sources and their enzymes perform the quint essential function for degradation pathway employed as microbial source or in its pure form. These are considered as biotransformation enzymes and include: Phase I biotransformation enzymes (which catalyzes the modification of pollutant), Phase II enzymes (which catalyzes whole group or biomolecule translocation reactions), Phase III enzymes (catalyzes translocation of pollutant and their metabolite making it unavailable). Significant enzymes appropriate to these kinds are oxidoreductases, hydrolases, transferases and translocases. Amidoxido reductases, the most frequent are monooxygenases (like cytochrome P450), dioxygenases, peroxidases and oxidases. Hydrolases like esterases are involved in biodegradation pathways. There



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are numerous categories of transferases, and they are classified corresponding to the group they conjugate to the xenobiotic: methyl-transferases, acetyl transferases, glutathione S-transferases among others (Velazquez-Fernandez et al., 2012).

Factors That Affect Bioremediation All ecological factors related to the bioremediation process directly or indirectly affect the process such as temperature, pH, water and nutrient availability, microbial consortium acting on substrate etc.

Various Physico Chemical Reactions Occurring during Biodegradation Reductive dechlorination is a chief mechanism opted by the microorganisms for conversion of organochlorine pesticides especially the o,p’-DDT and p,p’-DDT under reducing conditions. The reaction prevails with the substitution of one aliphatic chlorine molecule in exchange of one hydrogen atom. This is an enzyme catalyzed biochemical reaction performed by several microorganisms namely, Aerobacter aerogenes, Bacillus sp, Pseudomonas sp., Arthrobacter sp., Sphingobacterium sp. and Micrococcus sp. among fungi, Penicilliummyczynskii, Aspergillus sydowii, Trichoderma sp., Penicillium raistrickii, Bionerctia sp. are reported (Barragan-Huerta et al., 2007). Daughton et al., (1979) examined Pseudomaonas to degrade methyl phosphonate, O-isopropyl methyl phosphate, O-pinacolyl methylphosphonate (the initial breakdown product of Soman), ethylphosphonate or O-ethyl ethylphosphonate as sole phosphorus source. Lindane and DDT are few recalcitrant pesticides examples that can be biodegraded using microorganisms. These pesticides tend to accumulate in living cells and can persist at various concentrations in different trophic levels. Organophosphorous compounds are detoxified through microorganisms via hydrolytic cleavage of p-O alkyl and p-O aryl bonds (Singh and Walker, 2006). The hydrolyzed by product of the pesticide is found less toxic to living cells. A long list of bacteria and fungi are responsible for the detoxification of organophosphorous compounds such as Enterobacter sp. (Singh et al., 2003), Micrococcus sp. (Guha et al., 1997), Pseudomonas diminuta (Serdar et al., 1982), Flavobacterium sp. (Mallick et al,. 1999). Among fungi; Phanerochaete chrysosporium (Bumpus et al., 1993), Hypholamafasicularae (Bending et al., 2002), Coriolus versicolor (Bending et al., 2002) Aspergillus sp. (Obojska et al., 2002), Trichoderma harzianum (Omar, 1998), Penicillium brevicompactum (Omar, 1998). Documented parathion degrading bacteria in past literature reveals potential of bacteria in soil to detoxify these pesticide, e.g., Flavobacterium sp. (Sethunathan and Yoshida, 1973), Pseudomonas diminuta (Serdar et al., 1982), Pseudomonas stutzeri (Daughton and Hsieh, 1977),


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Arthrobacter spp. (Nelson et al., 1982), Agrobacterium radiobacter (Horne et al., 2012b), Bacillus spp. (Nelson et al., 1982), Xanthomonas sp (Rosenberg and Alexander, 1995). Methyl parathion degradation has been reported by bacteria by several researchers as for examples: Pseudomonas sp. (Chaudhry et al., 1988), Bacillus sp. (Sharmila et al., 1989), Plesimonas spM6 (Zhongli et al., 2001), Pseudomonas putida (Rani and Lalitha Kumari, 1994), Flavobacterium balustinum (Somara and Sidavattam, 1995). Another organophosphorous pesticide is glyphosate that is being reported to be degraded by several bacteria like Pseudomonas spp. (Kertesz et al., 1994a), Alcaligens sp (Tolbot et al., 1984), Bacillus megaterium (Quinn et al., 1989), Rhizobium sp. (Liu et al., 1991), Agrobacterium sp (Wacket et al., 1987), Geobacillus caldoxyticus (Obojska et al., 2002). Few fungi are also documented to detoxify the compound like Penicillium citrinum (Pothuluri et al., 1998), Penicillium notatum (Pothuluri et al., 1992), Penicillium chrysogenum (Klimek et al., 2001), Trichoderma viridae (Zboinska et al., 1992b), Aspergillus niger (Zboinska et al., 1992), Alternaria alternate (Lipok et al., 2003). Caumaphos is another category of organophosphorus pesticides often degraded by bacteria such as Agrobacterium radiobacter (Horne et al., 2002a), Nocardiodes simplex (Mulbry, 2000), Pseudomonas montelli (Horne et al., 2002b). Dazinon is being degraded by Flavobacterium sp. (Sethunathan and Yoshida, 1973), Pseudomonas spp. (Rosenberg and Alexander, 1979) and Arthrobacter sp. (Barik et al., 1979). Fenitrothion is decomposed by Flavobacterium sp. (Adhya et al., 1981) and Burkholderia sp. (Hayatsu et al., 2000). Degradation studies of methyl parathion was recorded by Pseudomonas aeruginosa, Bacillus megaterium and Staphylococcus aureus collected from rhizospheric region of cabbage, guava and tomato in Tripticase Soy broth at varying concentrations, growth of the bacteria were analyzed (Peter et al., 2014). Prior to opt for a bioremediation technology certain conditions and criteria must be kept in mind viz. a contaminant heterogeneity and concentration of the dye present, persistence and toxicity of the compound, interaction of the compound with biotic and abiotic factors, favorable conditions for biodegradative approaches, knowledge of natural bioprocess at contaminated sites, authenticated microbial degradation operating protocols, knowledge of soil enzymes (Singh, 2008).

DYES Colorants have been in use since ancient times across the globe. Archaeologists record the history of the process of dyeing (application of dye to substrate) to the Neolithic period. The serendipitous discovery of Mauveline (Mauve) by Sir William Henry Perkins in 1856 pioneered the modern synthetic dye industry. Dyes are soluble, organic compounds that impart color through selective absorption. They may be natural



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(derived from plants, invertebrates or minerals) or synthetic (i.e., chemically derived). Dyes possess color because of the following reasons (Abrahart, 1977). 1) 2) 3) 4) 5)

Absorption spectra of dyes falling in visible range of light (400-700nm) Presence of minimum of at least single chromatophore group in the dye Dye containing conjugate groups (possess alternate double and single bonds) Resonating electrons of the chromophore group Possess auxochromes (carboxylic acid, sulfonic acid, amino or hydroxyl groups)

The largest family of the organic dyes is constituted of one or more azo groups (i.e., azo dyes). Acid dyes are the dyes commonly used for polyamide and proteinaceous substrates for instance nylon, wool and silk. It possesses anionic charge, making it suitable for polyamides which are present in nylon and protein present in wool, silk and leather. Basic dyes bear positive charge imparting affinity to polymers as poly (acrylonitrile) and protein fibers as silk and wool having carboxylate groups. Disperse dyes are chiefly used for substrates that are hydrophobic in nature like polyester and acetate. These are hydrophobic dyes, apt for Poly (ethylene terephthalate) and cellulose acetate type of substrates. Direct and reactive dyes are used for substrates which possess cellulose for example cotton, rayon, linen, and paper. Dyeing efficacy depends on affinity with the substrate. Dye manufacture therefore takes into consideration the target substrate and endues application. The chromophore group is the colour imparting group of the dye which is characterized by electron arrangement within conjugated double bonds or single bonds that may be scattered including heteroatoms eg. N, O, and S, with non-bonding electrons is accountable for the light absorption (Gomes, 2001). Different auxochromes (-NH3, -COOH, HSO3 and -OH) i.e., electron donors are even enhance color. The choromogen (benzene, naphthalene or anthracene ring) composes part of the chromogenchromophore assembly along with the auxochrome. The chemical classes of dyes more often hired on an industrial scale are azo, anthraquinone, indigoid, xanthene, arylmethane and phthalocyanine derivatives. Azo dyes possess -N=N- groups and are the principal and highly versatile dye in industrial application. Their stability towards light, wash processes and resistance towards microbial attack builds azo dyes the most extensively cast off dyes in textile, pharmaceutical and printing trades, with the textile manufacturing being the leading punter for azo dyes. It has been estimated that 10% of the dye material does not fix to the fiber and therefore, are expelled in the ecosystem especially as effluent in water bodies. These dyes are not swiftly degenerative under ordinary circumstances and are not normally eliminated during the waste water treatment process. These compounds thus, comprise a group of xenobiotic complexes that are stubborn in the biodegradation process. The chemical nature of the waste from textile industry ranges from organochloride built trash pesticides to heavy metals concomitant with dyes and dying


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process (Correia et al., 1994; Faraco et al., 2009). Acute toxicity tests showed that most textile dyes are not particularly toxic (Cripps et al., 1990). Nevertheless, their persistence and resulting long exposure time is of particular concern for the discharge of waste dye effluent, since these substances may exhibit chronic effects such as mutagenic, teratogenic damage and carcinogenicity towards biota (Pierce et al., 2003; Bae and Freeman, 2007). Majority of the azo dyes are water solvable and are freely absorbed through skin contact and inhalation directing to the threat of cancer and allergic responses. It behaves as irritant for the eyes and may be exceedingly toxic, if inhaled or ingested. For example, p-phenylene diamine (PPD) which is even termed as 1,4-diamino benezene or 1,4-phenylene diamine, is an aromatic amine, principal constitute of azo stains. The constituting azo colourants are lethal also responsible for skin soreness, develop dermatitis, chemosis, lacrimation, exopthamlmose, and long-lasting blindness. Consumption of PPD goods directs to the quick progress of edema on face, neck, pharynx, tongue and larynx along with respiratory distress (Sudha et al., 2014). Certain azo colorants are accompanied with various forms of cancers, sarcomas, chromosomal aberrations and nuclear anomalies. Dye discharges also involve oil fertility and plant growth by shifting the biological and chemical makeup of soil and water. Exposure of fish (C. carpio) to sub lethal concentrations of dye run-off adversely alters the rate of feeding, absorption and conversion (Amutha et al., 2002).

Removal of Dye from Wastewater Treatment of dye effluent involves not only decolorization but also degradation and mineralization. Various technologies adopted (Pereira and Alves, 2012) for treatment of dye effluents are described below. 1) Physical methods: Physicals methods are not reliable as they result in effective removal of color, but the dye molecules are not degraded, resulting in the dye becoming concentrated and needs appropriate disposal. Membrane filtration process i.e., nanofiltration, reverse osmosis and electrodialysis is one of the options. On the other hand sorption techniques i.e., physical sorption (van der waals, hydrogen and dipole-dipole) and chemical sorption (covalent and ionic interactions). 2) Chemical methods: A large amount of sludge generated due to chemical methods of treatment (e.g., precipitation) is highly toxic and problematic to safe disposal. Various strategies have been adopted to implement chemical degradation or decomposition of dyes are in use such as coagulation, flocculation combined with flotation, and filtration, precipitation- flocculation with Fe(II)/CaOH2, electroflotation electrokinetic coagulation conventional oxidation methods.



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3) Biological methods: Aerobic and anaerobic microbial degradation using microbial enzymes is the most preferred method for the degradation.

Bioremediation Strategies Biological treatment offers a cheap and environment friendly alternative for removal of dyes. Dyes are unwavering agents contrary to breakdown through numerous microorganisms and major dyes do not biodegrade over aerobic biological remedial operations. Most, including azo-dyes decompose through anaerobic circumstances, and the aromatic amines thus formed further degrade aerobically. The electron retracting character of azo linkages prohibits the susceptibility of azo dye particles for oxidative reaction. Bacteria with azoreductases (azo dye reducing enzymes) are able for this kind of degradation. Aromatic amines resulting from anaerobic reduction are potentially harmful and not further degraded anaerobically. Anaerobic degradation, therefore, is considered as the first stage of degradation. Second stage for complete degradation involves conversion of aromatic amines under aerobic conditions (Saranjraj, 2013). Proteus vulgaris exhibited decolorization of azo food dyes (Walker, 1970) as per theory; a redox mediator could facilitate the relocation of reducing equivalents from NADPH to the substrate dye. Pseudomonas strains been reported with the ability to degrade carboxyorange possessed aerobic azoreductases having monomeric flavine free enzymes that use NADPH and NADH as co-factors and reductively cleaved several sulphonated azo dyes (Zimmerman et al., 1982). A distinctive model for the non-specific reduction of azo dyes through bacteria that does not require passage of the azo dyes or reduced flavins through the cell membranes for Sphingomonas xenophaga was suggested by Kudlich et al., (1997). Quinone acts as redox mediators in the system reduced by quinone reducatse enzyme situated in the cell membrane of S. xenophagathis leads to reduction of azo dyes in the culture supernatant in a purely chemical redox reaction. The decolourization of dyes by facultative bacteria was documented by Hu in 1998. The process of decolourization occurs during the logarithmic growth phase and COD/BOD reduction during the maximum stationary growth phase of bacteria. Wong and Yuen (1998) described reduction of methyl red by Klebsiella pneomoniae and brilliant green by Kurthia sp.; Walker and Weatherley (2000) reported Bacillus gordonae Tectilon blue 4R-01; Chang and Lin (2001) reported the reduction of reactive red-22 by Escherechia coli; Diniz et al., (2002) showed Congo red dye decomposition through Sphingomonas xenophega; Kim et al., (2002) showed degradation of bromophenol blue using Clostridium perfringens; Parshetti et al., (2006) spoke about the malachite green degradation by Kocuria rosea MTCC 1532; Kalme et al., (2007) utilized Pseudomonas desmolyticum for degradation of Direct Blue-6; Ozdemir et al., (2008) evaluated the decomposition of Acid Black-210 through Vibrio harvei; Deng et


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al., (2008) had shown the Anthraquinone dye degradation via Bacillus cereus strain DC II; Kalyani et al., (2009) showed the degradation of reactive red-22 by Pseudomonas sp SUK-1; Dave and Dave (2009) analyzed the degradation of Acid Red-119 through Bacillus thuringiensis; Ponraj et al., (2011) conducted the Orange 3R degradation via Klebsiella sp.; Peter and Vandana (2014) revealed the Congo red dye decolourization by Pseudomonas aeruginosa; Fungi like Aspergillus terreus has been known for its role in industrial processes in the production of citric acid, kojic acid, cellulases glucanases etc. These industries serve as a constant source of fungal biomass which has in turn shown biosorption efficiency that can be harnessed in the decolorization of dyes from wastewater biodegradation of recalcitrant complexes as azo dyes can be assisted by the presence of surfaces that have appropriate pores to stimulate formation of microaerophilic niches and permit microbial growth. Moreover, it shows traits of adsorptive capacity of the pollutant and nutrients to assure fungal feeding. In this logic, a few agro-industrial remains could be used as opportune supports. Fungal decolorization was found as assuring substitute to replace or increment to present treatment process (Ramamurthy and Ummamaheshwari, 2013). White rot fungi usually possess dye decolorizing activity which has been correlated with their capability to generate lignolytic enzymes e.g., laccases, lignin peroxidase, manganese peroxidase and manganese independent peroxidase which exhibit broad substrate specificity (Chivakula and Renganathan, 1995; Heinfling et al., 1998). Phanaeochaete chrysosporium in particular has demonstrated degradation of a large spectrum of azo, anthraquinone and triphenylethane dyes with decolorization efficiencies more than 90%. Studies have shown algae possessing ability to degrade dyes. Mechanism of dye degradation in algae is also attributed to azoreductases (Jinqui and Houtian, 1992). Vyayakumar and Manoharau (2012) studied the degradation by indigenous cyanobacterium species of a textile effluent containing the dyes remazol and venyl sulfone and observed not only colour removal, but also reduction in the levels of the inorganic compounds such as nitrites, phosphates, ammonia, calcium and magnesium. The main cyanobacterium strains reported to be responsible for the removal of nutrients and chemicals from the industrial effluents were Westiellopsis sp., Lynghya sp., Oscillatoria sp. and Chlorella sp. (Jinqui and Houtian, 1992; Shah et al., 2001; Parikh and Madamwar, 2005; Omar, 2008).

BIOMEDICAL WASTE MANAGEMENT Microbial Assisted Degradation A persistent residual content of antibiotics results in its association with certain kinds of microorganisms that may lead to its resistance towards those microorganisms. Gupta et



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al., (2014) quantified the Tetracycline residues in poultry meat using HPLC technique. This shows the lateral entry of antibiotics to poultry and gain of resistance towards antibiotics by the associated microorganisms. Gupta et al. (2014) also analyzed ronidazole in samples at their CRL levels using LC-MS/MS technique. They performed analytical studies to validate the confirmation of dimetridazole residues in aqueous in compliance with the Commission Decision 2002/657/EC (Gupta et al., 2014).

Factors Which Affect Drug Stability Quite a good number of factors affect the stability of pharmaceutical products or drugs, few of them are discussed in the chapter. A drug in medicinal formulation is stable in its temperature optima. Most of the drugs are stable in cold temperature. Elevated temperature could induce oxidation, reduction or hydrolysis that may result in drug degradation or deformation. A drug is chemical compound even if of natural origin they are greatly affected by variation in pH. An acidic or alkaline pH influences the drug decomposition by altering their oxidative state. The pH from 4 to 8 is a range where most of the drugs are said to be stable. Change in pH of the drug can alter the solubility of the compound. Water induces chemical reactions to proceed like hydrolysis or redox reactions. Presence of moisture even promotes microbial growth and can route microbial contamination of the therapeutic compound. Few drugs are sensitive to light and may be lead to photooxidation of the drug. Presence of microbial cells may lead to degradation and it may become a source of new infection.

CONCLUSION Increased use of recalcitrant compounds in daily life are causing environmental problems and health issues. It is much needed to degrade these compounds in an ecofriendly way. In this chapter, various physical and biological techniques are discussed with special emphasis on microbial biodegradation. In comparison to other methods, bioremediation is relatively efficient and cost effective technology for reducing environmental pollutions. Recently, a variety of microbes and microbial enzymes has been described that are able to degrade various recalcitrant compounds. Moreover, genetic engineering can be employed to increase the catalytic efficiency of microorganisms used in bioremediation. Along with single specific microbes, a group of microbes called microbial consortia can be used for bioremediation of various recalcitrant compounds.


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Chapter 3

THE SIGNIFICANCE OF MICROBIAL CELL SURFACE ENERGY IN WASTEWATER BIOREMEDIATION Meryem Asri1, Alae Elabed1, Soumya Elabed1,2, Saad Ibnsouda Koraichi1,2 and Naïma El Ghachtouli1,* 1

Université Sidi Mohamed Ben Abdellah, Faculte des Sciences et Techniques, Laboratoire de Biotechnologie Microbienne, Morocco 2 Université Sidi Mohamed Ben Abdellah, Centre Universitaire Regional d’Interface, Route Immouzer, Morocco

ABSTRACT Biological processes, using microbial metabolism to eliminate or transform different hazardous contaminants, are among the most pertinent biotechnological applications for wastewater treatment. Expanding the base of knowledge on these applications allows a better control of the transition from laboratory scale to a fully commercialized technology. There are many findings in the literature proving that the microbial cell surface properties are important factors in controlling the interactions between pollutants and microbial surface. Among these properties, the surface free energy (SFE) between interacting surfaces has a significant role in the dictation of the strength of solid-solid and solid-liquid interactions. It presents a major effect in different important phenomena involved in bioremediation applications such as microbial adhesion to differently used supports in biofilm based processes, aggregation, flotation, flocculation and transport. This chapter includes: (i) An essential insight into methods commonly used for the determination of SFE and a critical description of different approaches, (ii) offers a discussion for understanding the crucial effect of SFE and its components of microbial cells and supports on wastewater bioremediation, and (iii) presents future research directions for the development of efficient wastewater bioremediation processes. * Corresponding Author Email: [email protected].


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Meryem Asri, Alae Elabed, Soumya Elabed et al.

Keywords: bioremediation, wastewater, surface free energy, microbial cells

INTRODUCTION Water scarcity is one of the most serious worldwide issues of our time. Currently, the global population lives in a water-stressed situation. The challenge of providing sufficient and safe drinking water gets further complicated by climate change and the contamination of available freshwater due to industrialization and population growth. Thus, water will become one of the scarcest resources in the coming decades. Therefore, many wastewater infrastructures and desalination systems were implemented in order to increase available water resources (Elimelech and Phillip 2011). The worsening unsanitary conditions due to heavy industrialization and urbanization induce a pressing need to the planning of suitable wastewater treatment projects. This would contribute to the remediation of water shortage in an efficient and economical manner. After more than 100 years of successful use, biological systems of wastewater treatment continue to predominate among the best environmental infrastructures for wastewater treatment. Extensive reviews have dealt with biological wastewater treatment systems regarding influencing parameters, including the environmental conditions, the bioreactor designs, etc. Recently, the increasing role of surface properties in many important scientific and technological areas led to a great debate regarding the understanding and the determination of the SFE of materials. Many works have reported that a thorough knowledge of the SFE parameter is of prime importance in various scientific areas such as the biomedical field (Van Der Valk et al. 1983), biofilm adhesion (Elabed et al. 2013), implant and tissue engineering technology (Hallab et al. 2001), metallic biomaterials (Ponsonnet et al. 2003), polymers characterization (Owens and Wendt 1969) and organic electronics (Janssen et al. 2006). However, works on the influence of surface properties for the efficiency of wastewater treatment systems are still restricted. Microbial cells or inert materials in a wastewater treatment system are exposed to different conditions in the aquatic environment. They usually exhibit different physical properties and energy characteristics. The determination of the SFE is of great relevance in the understanding of surface properties in wastewater treatment processes. This chapter reviews different approaches of the determination of SFE. It gains important insights into the effect of surface properties, in particular surface free energies of the involved phases of a wastewater treatment in different important phenomena within the system. Future research directions for the development of efficient wastewater treatment systems are also presented.


The Significance of Microbial Cell Surface Energy …

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MICROBIAL CELL SURFACE FREE ENERGY: COMPONENTS, ESTIMATION METHODS AND APPROACHES Free surface energy provides a more profound understanding of solids and liquids interactions which is of extreme importance in a wide diversity of technological applications, including surface properties such as adhesion, coating, printing, flotation and lubrication (Karagüzel et al. 2005), biomedical field (Van Der Valk et al. 1983), biofilm control (Elabed et al. 2013), tissue engineering technology (Hallab et al. 2001), biomaterials engineering (Ponsonnet et al. 2003), polymers characterization (Owens and Wendt 1969) organic electronics (Janssen et al. 2006). Microbial surface characteristics had been identified as important parameters in various applications. Indeed, numerous investigations have showed that depth research concerning SFE parameter and its components contribute directly to the modelling studies in microbial cell-cell interactions and cell-surface for the diversified technology applications (Zhang et al. 2010). The natural free surface energy of microbial cells is mainly due to specific properties of the cell wall (i.e., the biosorption sites), which strongly depends upon the presence of polysaccharides, proteins and lipids that form a biopolymer layer (Zouboulis et al. 2004). Numerous investigations attempted to find reliable methods and to evaluate different approaches for a better SFE computation (Sharma and Hanumantha 2002). Since the first application of SFE methods, there was a great deal of excitement and a debate for a long time. The different used methods were criticized in several reviews (Volpe and Siboni 1997; Zdziennicka 1999). However, one of the most successfully used methods for SFE determination is contact angle measurement (Jalqczuk et al. 1983). In this method, contact angle of filtrated cell lawns or solid materials using the sessile drop technique are measured, it is followed by a theoretical interpretation allowing the SFE determination (Nguyen et al. 2011; Visa et al. 2012; Maataoui et al. 2014; Zhang et al. 2014). By employing a number of different polar and apolar liquids with known SFE components, it yields a determination of the solids SFE (γS) and its components including the lifshitzvan der waals (γSLW) acid-base component (lewis) (γSAB) the electron acceptor (γS-) and electron donor (γS+) parameters. These measured contact angles are the basis of the surface energy calculation, while different thermodynamic approaches may be used and the results are different depending on the followed approach. The most used approaches are Fowkes, Equation of state, Geometric mean and Lifshitz-van der waals acid-base (LW-AB) approaches (Sharma and Hanumantha 2002). Among these approaches, LW-AB approach was found to be able to yield consistent values of SFE and its components and was hence the most reported in the literature (Bayramoǧlu et al. 2005). The total SFE, γTOT, can be expressed by two components: γTOT = γLW + γAB


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γLW and γAB refer to the dispersive and acid–base components of the free surface energy, respectively. Substituting the appropriate expressions, the following equation is obtained:

 L (cos   1)  2  sLW  LLW    s L   ( s L )1/2  1/2



1/2



Where γ+ and γ− are proton and electron donating character of acid–base component, respectively.

APPLICATION OF SURFACE FREE ENERGY IN WASTEWATER TREATMENT Among the most important technological applications, wastewater treatment has received a great part of scientific interest. In this chapter, we are reporting the importance of SFE on different phenomena in wastewater treatment including flocculation, granulation and adhesion in biofilm-based systems and pollutants removal.

Flocculation and Sedimentation A crucial part of the wastewater treatment is the solid-liquid separation, where the treated water is separated from the activated sludge (Christensen et al. 2015). Thus, a good functioning of many types of wastewater treatment process depends mainly on the flocs formation which are aggregates of microorganisms constituted by microbes, inorganic particles and exocellular polymers (Urbain et al. 1993). Flocculation was for a long time realized using flocculating chemicals to recover the microbial biomass from the aqueous suspensions. This method was reported to be energy requiring process and the chemicals addition is potentially contaminating active biomass (Ozkan and Berberoglu 2013a). Consequently, bioflocculation received increased attention over conventional methods for suspended biomass harvesting, which is involving other flocculating microbes able to improve cell-to-cell association (Salim et al. 2011). Thus, accurate knowledge of the decisive factors of a good flocculation and a profound understanding of the cells interaction mechanisms are very important so as to develop proficient flocculation process (Higgins and Novak 1997; Salehizadeh et al. 2000). Above all, surface properties were demonstrated to have a great role in this microbial strong association (Liao et al. 2001; Ozkan and Berberoglu 2013a). The SFE and its components were proved to be correlated to flocculation behavior (Olofsson et al. 1998). A relation



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between this surface parameter and the flocculation potential within the flocs has been found. Thus, the hydrophobic cells were able to adhere within flocs while the hydrophilic cells did not (Olofsson et al. 1998). It was also showed that bacteria exhibiting low SFE (hydrophobic) have better capacity of attachment to activated sludge flocs than positive SFE (hydrophilic) bacteria (Zita and Hermansson 1997). In the case of algal biomass, SFE and acid-base component are the main mechanisms promoting the initial algal cell to cell attraction (Ozkan and Berberoglu 2013a). Zhang et al. (2014) showed that for microalgal cells, SFE has a direct impact on their bioflocculation behavior. Other data have reported that by decreasing the SFE (i.e., enhancing their hydrophobicity) of Cyanobacteria microsystis, an increase in the cell-cell attraction was obtained (Ozkan and Berberoglu 2013a). Obviously, microalgae strains with high SFE or exhibiting hydrophilic character are most likely in a planktonic state without flocculation as a result of the difficulty of excluding water among the algae cells (Ozkan and Berberoglu 2013a). By contrast to these studies revealing the importance of SFE parameter in microbial bioflocculation, some other data showed that the flocculation capacity is not only related to surface properties but also to other factors such as exopolysaccharides (EPS) production (Becker 1996). Sedimentation of the formed flocs is also a step of extreme importance for the good quality of the resulted effluent. During this step dispersed materials in the aquatic environment including the microbial cells and the small flocs will be attached to the flocs (Olofsson et al. 1998). Therefore, for the management of the wastewater system, it is important to study both flocculation and subsequent sedimentation mechanisms. Intensive works aimed a sound understanding of influencing factors on this rate-limiting step (Urbain et al. 1993; Wilén et al. 2003; Badireddy et al. 2010). Numerous investigations tested the role of crucial parameters on microbial sedimentation behavior. It was reported that a better knowledge of the SFE of microalgal cells might help to predict and control the sedimentation behavior of microalgal cells. Thus, it allows reducing the cost and the required energy for microalgae cultivation and harvesting (Zhang et al. 2014).

Adhesion in Biofilm-Mediated Systems Biofilm systems of wastewater treatment represent a proficient strategy increasingly used for assessing environmental effects. It is mainly due to the great capacity to respond rapidly to physical, chemical and biological stress. Several biofilm based processes were found sufficiently attractive to pass from laboratory scale to a larger scale (Mitra and Mukhopadhyay 2016). A biofilm is defined as a complex coherent structure of cells and cellular products forming dense granules or attached on a static solid surface or adhered to suspended carriers (Lettinga et al. 1980; Nicolella 2000). Existing in the biofilm mode is advantageous for bioremediation because of the greater tolerance to pollutants,


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environmental stress and the ability to degrade varied harsh pollutants via diverse catabolic pathways. In aquatic environments, the interactions between microbial cells and solid surfaces imply a variety of crucial factors. Among these parameters, bacterial surface properties and substratum characteristics have been identified to be of extreme importance in adhesion process. Microbial attachment to the employed supports in a bioreactor is a key step for its treatment efficiency. There has been a great progress in the understanding of bacterial attachment mechanisms (Elabed et al. 2012). Many works dealt with the biofilm formation in the environmental field, by developing the understanding of microbial adhesion phenomena in wastewater systems, it allows an optimization of the biofilm treatment performance during its functional lifetime. Despite the fact that the different types of interactions that might determine adhesion are now well known, their experimental evaluation has been of extreme difficulty. However, many theoretical approaches were developed attempting to understand the involved interactions in microbial attachment (Hermansson 1999; Nguyen et al. 2011). The most reported theoretical approaches are the Derjaguin–Landau–Verwey and Overbeek (DLVO) theory, thermodyamic models and extended Derjaguin–Landau–Verwey and Overbeek XDLVO theory. The DLVO theory takes into account the attractive van der waals interactions and repulsive electrostatic interactions. It predicts two different distances of separation at which there is net attraction between the surface and an approaching microbial cell. At the “secondry minimum” the separation distance is larger and the microorganims are not strongly adhered to solid surface and adhesion is reversible, in other words the attached cell can easily be removed by shear forces. At this distance the approachment is inhibited by repulsive electrostatic force. Second distance is “primary minimum” where attractive van der waals forces are more strong and adhesion is hence irreversible at this distance (Fletcher 1996). According to this approach the total energy of interaction is written as a function of the separation distance (d) between approaching microbial cell and solid surface: GDLVO (d) = GLW (d) + GEL (d) A non-negligible number of studies have used the DLVO theory as a predictive tool for microbial adhesion study, but in most cases this theory was unable to predict experimental results, alternatively other theoretical models have been suggested. In theromdynamic theory, the adhesion is considered as an equilibrium process, where the biofilm formation is a spontaneous phenomena accompanied by a decrease in the system free energy. The employment of this theory requires the determination of interfacial free energies of both microbial cells and substrata. This can be obtained indirectly by contact angles measurment technique. Likewise, this thoery was incapable to give consistent results with experimental observations in all cases, in many cases they were not explained by this model (Bellon-Fontaine et al. 1990). The uncapacity of this theory to describe



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experimental results may be due to different reasons. The great heterogenous nature and the chemical complexity of microbial surface which makes the identification of SFE components of great difficulty. On the other hand, thermodynamic theory considers an equilibrium situation, whereas microbes are known for their adhesive polymers production which may be directing metabolic energy into adhesion initiation or attachment stabilization. Moroever, the determination of SFE values used for prediction is particulary difficult in the absence of reliable standardized methods for the determination of this surface property (Fletcher 1996). XDLVO model was successfully reported for its capacity of prediction and the study of biofilm formation (Hamadi et al. 2009; Elabed et al. 2013; Asri et al. 2017). It takes into account the surface free energies of the different constituent phases of the biofilm including microorganisms, substratum and separating liquid. The SFE of biofilm constituents is commonly obtained using the liquid contact angles on their surface, where the formed angle by the drop of liquids with known parameters placed on the surface allows the determination of chemical and physical interactions between the surface and the liquid (Hamadi et al. 2004). In addition to electrostatic and van der Waals forces in the DLVO approach, the XDLVO approach takes into account the contribution of the acid-base (AB) interactions in the biofilm formation. AB forces refer to electron transfer interactions between polar components of the cell and the surface. Based on the hydrophobicity of the interacting surfaces, these interactions may be attractive (hydrophobic attraction) or repulsive (hydrophilic repulsion) (Van Oss 1986). The total interaction energy GTOT is given as (Ozkan and Berberoglu 2013b): GTOT = GAB + GLW + GEL Although the successful use of theoretical models to adhesion prediction may give an indication on the most significant physico-chemical parameters and attachment conditions, it is clear that the process is determined by a combination of numerous chemical, biological and temporal factors. Moroever, microbial cells may overcome the free energy barrier by protruding pili and flagella or EPS including chemical groups to interact with organic matrix forming different types of bonds i.e short range electrostatic, covalent or hydrogen ones which bridge the separating distance between cells (Fletcher 1996). Although the contact angle measurement is the most reported as a reliable, rapid and easy for surface free energies estimation, its realization on dried microbial lawns has been criticized. It is the case because the changes in the adsorbed macromolecules may have inaccurate results that can imply a discordance between theoretical predictions and experimental adhesions. Thus, alternative methods using microbial cells within liquids, which are in a more natural state, were proposed such as bubble contact angle method (Fletcher 1996). Moreover, the prediction models of surface interaction consider that surfaces are energetically homogenous, which is invalid for microbial cells (Costerton et


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al. 1978; Busscher et al. 1990). Thereby, no physicochemical approach was able to fully explain microbial adhesion to solid surfaces. In wastewater treatment applications, biofilm adhesion was discussed for its great importance in these processes. It has been shown that microalgal strains with a lower SFE (i.e., more hydrophobic) had a higher adhesion density and strength on solid surfaces used for algae cultivation for wastewater treatment or other biotechnological applications of algal cells (Ozkan and Berberoglu 2013b; Zhang et al. 2014). The efficiency of in-situ bioremediation of contaminated aquifers is often dependent on the ability of the effective transport of microorganisms to the contaminant site, where the propensity of cells adhesion to porous media surfaces is one of the affecting phenomena. Physicochemical interactions between these two components have been reported to govern the adhesion phenomenon (Smets et al. 1999).

Granulation or Aggregation Sludge granulation in biological wastewater treatment reactors is referring to the selfimmobilization of microbes, which results in a compact structure of aerobic and anaerobic granules (Sheng et al. 2010). Granulation process involves cell-to-cell interactions including chemical, biological and physical phenomena. These granules are mainly formed by microbial consortia packed with different bacterial species containing millions of organisms per gram of biomass. The activated sludge systems is by far the most common process, however, alternative processes such as biofilm-based technologies or granules systems also exist (Christensen et al. 2015). For the latter, granulation process is a key step in these systems functioning. In order to optimize the granules systems treatment efficiency, important phenomena such as agglomeration should be thoroughly understood. Many factors may affect microbial granulation ability. Thus, several investigations have provided insight into the role of different characteristics on granulation phenomenon (Hou et al. 2015). The mature granules in granular systems are depending mainly on the bacterial attachment. Several models have been suggested in order to better control the structured aggregate formation which is a crucial step in the granulation process. The starting point of the microbial cells granulation process is their attachment to a solid surface or by self-immobilization of microbial cells (Liu et al. 2003). The granulation process can be expressed in terms of total energy of interaction between microbe-to-microbe or microbe-to-solid surface. Thus, on the basis of thermodynamic point of view, some physico-chemical models have been suggested to model the granulation process. Two of the suggested models are related to DLVO theory, where interfacial surface energy is a decisive parameter. In the “multi-valence positive ion-bonding” model, microbial attachment may prevent the approach of one cell to another in an aquatic environment since the attachment according to multi-valence positive ion-bonding model is based on simple electrostatic interaction between



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negatively charged bacteria and positive ion. This model is in concordance with the DLVO theory showing that when the two surfaces carry a charge of the same sign, a free energy barrier occurs between them and acts as a repulsive force. Thus, according to this model adding a positive ion (such as calcium, ferric, aluminum or magnesium ions) to sludge may decrease electrical repulsion among negatively charged bacteria, would promote cell-to-cell association as a decisive step in granulation process (Liu et al. 2003; Schmidt and Ahring 1993; Yu et al. 2001; Zita and Hermansson 1994). Cell-to-cell interaction could also be modeled by “secondary minimum adhesion model” based on DLVO theory for colloidal particles. This model considers only the thermodynamic aspects of microbial interaction, taking into account both the surface charge of microbial cells and SFE or hydrophobicity as important parameter for long and short range forces. According to this DLVO related model, the cell-cell association starts with a reversible attachment in the “primary minimum” which becomes irreversible in the “secondary minimum” (Liu et al. 2003). Hydrophobic interaction model for microbial granulation also had been proposed by Wilschut and Hoekstra (1984) who showed that irreversible interaction among hydrophobic bacterial cells (low SFE) are strong. According to this model, increasing cell surface hydrophobicity or in other words decreasing cell surface energy would initiate and strengthen the cell-to-cell interaction and may also facilitate their solid-liquid separation. This theoretical model was supported by experimental observations showing that microbial surface hydrophobicity is crucial for granulation initiation (Karr et al. 1978; Liu et al. 2003; Loosdrecht et al. 1987; Wu et al. 1991). A further model “surface tension model,” suggests that the microbial granulation is a process requiring an interruption of preexisting individual bacteria-liquid interface and the creation of a new interface granule-liquid. The free energy of adhesion between the two adhering bacterial can be expressed as (Rouxhet et al. 1990): ΔGadh = 2(γb1/2 - γl1/2) (γl1/2 - γs1/2) Where γb, is the SFE of bacteria, γl is the SFE of liquid and γs is the SFE of inert particle. On the basis of this equation, the aggregation ability is related to the relation between the bacterial SFE and the liquid SFE. Thus, it is favored when γb is lower than γl which means the decrease of the total free energy of interaction. On the opposite, the association in unfavorable when the SFE of the bacteria is higher than that of the liquid. A “general” model for microbial granulation based on four steps was proposed as the most complete to depict the granulation process. This model defined the different forces involved in granulation process. Among the most important attractive forces to keep stable multicellular contacts according to this model are physical forces including Van der Waals forces, opposite charge attraction and thermodynamic forces including the SFE and the surface tension (Liu et al. 2003). Apart of these models, others were reported in


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literature, and attest the contribution of the SFE parameter on the initial granulation process in wastewater reactor (Liu et al. 2003).

Pollutants Removal Depollution behavior in the aqueous environments has been known to be influenced by the physicochemical interactions between the biosorbent surface and the contaminant. There may be various different interactions categories associated with different types of contaminants: particulate, inorganic and organic matters. In many bioremediation applications, relevant works have shown that the interaction between microorganisms and the matrix surfaces is of paramount importance. In microbially-mediated in-situ pollutant degradation, these interactions are the key event that determines the microorganisms’ migration in the geological formation and consequently the efficacy of the pollutant removal (Grasso et al. 1996; 1998). The chemical composition of microbial cell surface was found to be in correlation with their surface energy components. Thus, different biomass types exhibit different microbial cell surfaces properties. The surface of bacterial cells is essentially composed of polysaccharides, proteins and lipids offering a dominant electron donor character due to the presence of many negatively charged functional groups such as carboxylate, hydroxyl, thiol, sulphonate, phosphate, amino and imidazole groups (Gong et al. 2005; Liu et al. 2011). Their presence is involved in their bioremediation of polluted aqueous media and their capacity of binding metal ions (Gong et al. 2005). Fungal cell walls contain a great amount of polysaccharides and proteins offering many functional groups such as carboxyl, hydroxyl, sulphate, phosphate and amino groups involved in their bioremediation capacity (Veglio’ and Beolchini 1997). Fungal biomass cell walls also contain chitin, chitosan, polyuronides and melanin which have been shown to be able to sequester metal ions (Bayramoğlu et al. 2005). In the case of the removal of charged coumpounds such as heavy metals, the surface modification is mostly influencing the polar component of the SFE. This is due to the importance of the charge-charge interaction withing the heavy metals removal. Fungal treated biomass using heat, acid or base treatment resulted in the increase of polar component of SFE and concenquently the interaction with hexavalent chromium ions (Cr6+) (Bayramoǧlu et al. 2005). Treatment like heat and acid/base treatments are known to change the components of the SFE of biomass, essentially the increase in the polar component of its SFE. This increasing in surface polarity influences the charge-charge interaction between the biomass and metallic ions and hence their adsorption capacity (Bayramoğlu et al. 2005). Microbial cells presenting low SFE can be used to remove aromatic and xenobiotic organic compounds from aqueous environment. In fact, hydrophobic microbial cells were shown to have high ability to accumulate and decompose these highly hydrophobic and toxic pollutants (Kobayashi et al. 1999).



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Although microorganisms exhibiting low SFE were reported to have better bioremediation potentials, hydrophilic microorganisms were also able to play a considerable role in organic wastes treatment from aquatic environments due to the high resistance to hydrophobic pollutants (Krasowska and Sigler 2014). In some cases, the hydrophilic microbial cells were able to prevail over the hydrophobic microorganisms due to the great amount of lipopolysaccharides in their membranes offering a protection against organic molecules attachment (Kobayashi et al. 1999). Moreover, these microorganisms are advantageous in bioremediation processes when the fast dispersion of inoculated microorganisms is desirable due to their low adhesion tendency (Obuekwe et al. 2009). The limitation in hydrophilic microbial cells use is that they are not able to bind hydrophobic pollutants like hydrocarbons exhibiting extremely low water solubility and low bioavailability. In addition, in granulation process, microorganisms exhibiting high SFE are not able to form stable flocs in granular systems. To overcome this limitation, it was proposed to develop strategies to increase the accessibility by modification of the microbial cell surfaces properties (Krasowska and Sigler 2014). Not only biological systems of wastewater treatment are influenced by the free energy components, but also other depolluting strategies. Numerous works have been realized to understand the relation between membrane properties and their applications in water and wastewater treatment. Membranes post-treatment steps, such as the application of a surface coating layer, are proposed to change the membrane surface physicochemical properties in order to protect membrane surface and to improve membrane depollution performances. The rigorous study of the optimal membrane surface physiochemical properties gives to theusers an aid to choose the most suitable properties for specific applications. Different membranes may have different abilities of adsorption and capture of the particle due to their intrinsic characteristics such as hydrophilic character (Madaeni 1999). Membranes fouling are a complex phenomenon including organic and inorganic components and biofouling (caused by microorganisms). It is regarded as one of the most important limitation for the membranes filtration process and their practical applications in wastewater treatment. To date, fouling is the most important factor that has limited the use of membrane technology (Madaeni 1999). Thus, a lot of research efforts have been made to control this problem and develop anti-fouling strategies. Physico-chemical properties were proved to be of extreme importance in this regard. Indeed, membrane hydrophobicity parameter had been believed to play an essential role in the membrane separation processes. Hence, improving the hydrophilicity of a membrane was shown to be related to the surface adsorption properties and can reduce its fouling ability (Yan et al. 2005). A developed membrane showing hydrophilic and oleophobic surface properties presented high resistance against organic and biological fouling (Zhu et al. 2013). Another recent work reported results supporting the long established view indicating that hydrophilic membranes are less susceptible to fouling (Onate et al. 2015). As the most available


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membranes are hydrophobic, several surface modifications were carried out to increase membrane hydrophilicity such as the addition of Al2O3 particles (Yan et al. 2005), the hydrophilic substances adsorption and the surface coating (Akthakul et al. 2004; Combe et al. 1999; Nunes et al. 1995); hydrophilic groups production by strong acids application or plasma (González Muñoz et al. 2006; Yu et al. 2005), incorporation of hydrophilic polymers on membrane surface (Asatekin et al. 2007; Zhao et al. 2007), functional monomer or polymer grafting on the membrane (Pieracci et al. 1999; Howarter and Youngblood 2009; Zhao et al. 2010). Regarding biofouling caused by microbial adhesion on membrane surface, the possible forces that promote this process are electrostatic, Van der waals, ionic, hydrogen bonding and hydrophobic interactions. It has been claimed that the hydrogen bonding and electrostatic forces appears to be of minor importance in the adhesion process. Hydrophobic interactions have been considered to be the main driving force for bacterial adhesion on membrane surface. Madaeni (1999) explained hydrophobic interaction by considering the free energy for bringing two solute molecules together from infinite separation. The possible antifouling mechanism of a hydrophilic membrane surface has been attributed to the formation of a compact hydration layer on the surface that may minimize the direct contact between the membrane surfaces and fouling substances such as proteins, natural organic matters and bacteria (Chen et al. 2011). In the case of oil-water separation, the improvement of membranes hydrophilicity is reducing hydrophobic contaminations adsorption potential included in oil–wastewater (Yan et al. 2005). In order to minimize the strength of organic foulants adhesion and to facilitate membrane cleaning, membranes of low SFE have been believed to be advantageous. They showed a good resistance to the fouling of oils under dynamic stirring which may be attributed to the weaker adhesion. But these membranes have also showed poor affinity with water and hence had a very low water permeability or water flux (Hamza et al. 1997). However, highly hydrophobic membranes were indeed commonly observed to be easily fouled by biological and organic foulants such as microorganisms and humic substances, etc. (Zhu et al. 2013). The ability of a membrane to have the fouling problem is then depending on the relative interactions among water and the membrane surface and foulants in water. The surface tension of water and the interfacial free energy of the membrane surface and foulants in water are regarged as the most conditioning parameter. Organic and biological foulants in water are often completely or partly hydrophobic exhibiting much lower interfacial free energy than the surface tension of water. A preferential accumulation on a hydrophobic membrane would be resulted because of the natural repulsion by aqueous medium which may explain the reason why a hydrophobic membrane is more susceptible to organic and biological fouling. While in the case of hydrophilic membrane, the hydration layer on the membrane surface may reduce the adhesion of the foulants (Zhu et al. 2013).



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Microbial Fuel Cells Microbial fuel cells (MFCs) are a promising technology for electricity production from a variety of organic components and can be advantageously combined with applications in wastewater treatment. Since 2002, the number of scientific publications on MFCs has been exponentially increased. This growth rate is indicative of the significant potential of this technology. A conventional MFC comprises an anode compartment, a cathode compartment and a proton exchange membrane separating them. Electrons produced by oxidizing the organic matters in anode compartment are transferred to anode electrode from which they flow to cathode electrode via external resistance. MFC studies are reported with different accent or emphasis including: (i) microbial metabolism, (ii) biocathodes used on the cathode side, (iii) biocatalyst used in both anode and cathode, (iv) different MFC configurations and designs, (v) different construction and materials, and methods of data analysis and reporting in order to give information to the MFC developers. MFC studies also concerned performances and optimization of operating parameters of the MFC systems, anodic electron transfer processes, different substrates used in MFCs, applications of MFCs, microbial communities and different electrode materials and designs both for the anode and cathode side. Electrodes used in MFCs are the most important factors affecting performance and cost of MFCs. Electrode materials in MFCs and their configurations have been focused on in recent years and various materials have been studied. Carbon is the most commonly used electrode material in MFCs, generally as graphite rod (Zhang et al. 2013), graphite fiber brush (Liu et al. 2013), carbon cloth (Lu et al. 2014), carbon paper (Alatraktchi et al. 2014) or carbon felt (Flexer et al. 2013). The electrode material and morphology should facilitate bacterial attachment and subsequent biofilm formation. At the same time, anode surface chemistry along with the biofilm formation should enhance electron transfer from bacteria to the electrode. Several thermal or chemical treatments have been described to reduce MFC start-up time by facilitating rapid cell attachment and biofilm development for enhanced power output in MFC (Rinaldi et al. 2008; Wei et al. 2011). Thermal treatment of the electrodes leads to modification of the surface roughness and porosity and thus enhances cell concentration and biofilm development (Manickam et al. 2013; Wang et al. 2009). Depending on the gas atmosphere (e.g., nitrogen, oxygen, ammonia) used in thermal treatment, hydrophilic functional groups can be added on the electrode surface (Cheng and Logan 2007). Several compounds, such as nitric acid (Zhou et al. 2011; Zhu et al. 2011), ammonium nitrate (Zhou et al. 2012), ammonium persulfate (Zhou et al. 2012), polyaniline (Lai et al. 2011) have been used for surface chemical modification of carbonaceous electrodes. However, as previously shown (Li et al. 2014), the chemical treatment of carbonaceous surfaces (e.g., carbon cloth) affects both the chemistry and the morphology (e.g., roughness and porosity) of surfaces. Recently, Guo et al. (2013) studied the influence of


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the surface charge and hydrophobicity on the biomass accumulation, taxonomic distribution and electrochemical activity of multiple surface-modified anodes operated in half-cell bioelectrochemical systems. Anodes consisted of modified glassy carbon electrodes through electrochemical grafting with aryl dyazonium salts. As a result, the surface of each anode has been altered distinctively to be hydrophilic (−OH, − SO3,−N(CH3)3+) or hydrophobic (−CH3) with positive, negative or neutral charge (Guo et al. 2013). The researchers found that the most positively charged and hydrophilic surfaces were associated with improved biofilm formation and selection of electroactive microbes such as Geobacter spp. (Guo et al. 2013). A similar conclusion has been reported by Picot et al. (2011), who observed significant increases in anode current output when the surface was amended with positively charged phenylphosphanium cations.

CONCLUSION Surely, SFE is a parameter of extreme importance in environmental research. Its estimation is an important parameter that has proven to play a paramount role in different phenomena in wastewater treatment. However, it can be concluded that the understanding regarding the effects of SFE on wastewaters treatment processes is far from complete. Hence, much research is still needed to fully control these parameters. Some of these future research niches are therefore outlined. Since the surface adsorption of pollutant on microbial cells is the first step in the removal process, the favorable SFE components for the removal of each pollutant has to be determined. In addition, the influence of SFE on the EPS biosynthesis by microbial biofilms, which is an integral role for the depollution capacity is a task of great importance that should be understood. The most important open research questions are regarding the ability of the surface properties modification of microbial cells or other materials to modify their removal potential.

ACKNOWLEDGMENTS The authors are grateful for the financial and scientific support provided by Microbial Biotechnology Laboratory of Faculty of Sciences and Innovation City, Sidi Mohammed Ben Abdellah University, Fez, Morocco.



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Chapter 4

THE USE OF GREEN MICROALGAL CULTURES FOR BIOREMEDIATION OF FRESHWATER ENVIRONMENTS POLLUTED WITH CHROMIUM, NICKEL AND CADMIUM Laszlo Fodorpataki1,*, Sebastian R. C. Plugaru2, Katalin Molnar3, Peter Marossy1, Bernat Tompa1 and Szabolcs Barna1 Hungarian Department of Biology and Ecology, “Babes-Bolyai” University, Cluj-Napoca, Romania 2 Faculty of Materials and Environmental Engineering, Technical University, Cluj-Napoca, Romania 3 Department of Horticulture, Sapientia Hungarian University of Transylvania, Targu-Mures, Romania 1

ABSTRACT Planktonic microalgae, as the main primary producers in aquatic environments, exhibit not only a crucial capacity of promoting essential biogeochemical cycles, but also an inducible metabolic plasticity which enables them to bioaccumulate various polluting agents, thus contributing to natural bioremediation processes. Axenic cultures of selected algal strains ensure controlled conditions for studying the optimal exposure time, heavy metal concentration ranges, pH values and other adjustable parameters for a high efficiency of bioremediation. Soluble cationic forms of cadmium, nickel and hexavalent chromium are frequent and dangerous water pollutants of industrial origin, and they induce specific changes in vital functions of algae. In this chapter we exemplify how selected physiological markers of heavy metal tolerance, such as parameters of induced *



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Laszlo Fodorpataki, Sebastian R. C. Plugaru, Katalin Molnar et al. chlorophyll fluorescence, photosynthetic pigment ratio, dry biomass production and dynamics of cell density may be valuable tools in developing cost-effective protocols for bioremediation of aquatic ecosystems anthropically polluted with different concentrations of chromium, nickel and cadmium ions, with an exposure time ranging from a few days to at most two weeks in static systems, at moderately acidic pH values. Knowledge of algal reactions to water pollution with heavy metals is indispensable for a correct planning of bioremediation technologies for various aquatic environments.

Keywords: heavy metal tolerance, microalgae, phytoextraction, water pollution

INTRODUCTION Microalgae are widely distributed in every aquatic environment of the Earth, and as the main primary producers of new organic compounds in freshwater and marine ecosystems, they play a determinant role in the aquatic trophic networks, as well as in the biogeochemical cycle of vital chemical elements. Various microalgal species and intraspecific varieties are differently sensitive to metal toxicity and therefore are suitable biological indicators for an early detection of potential toxic effects of heavy metals dissolved in the aqueous solution. Due to their metabolic plasticity, the ability of selected microalgae to sequester various metal ions makes them very efficient in the biological remediation of large water bodies which contain lower concentrations of heavy metal ions [155]. Removal of metals from wastewaters containing higher amounts of such contaminants may be cost-effective using nonviable algal biomass as biosorbent, while bioconcentration of heavy metals from polluted water, with the possibility of extraction of valuable metals may be achieved by using microalgae immobilized in diferrent embedding agents. Algal metabolic and developmental parameters may be very useful in identification and selection of algae with different physiological characteristics that enable an optimized remediation of polluted water. Resistant microalgae have the ability to avoid the uptake of heavy metals, but they adsorb toxic metal ions to cell wall constituents or they secrete metal-binding organic products to the surrounding environment. Heavy metals are very stable environmental contaminants because they cannot be degraded or destroyed, but any process that enhances transformation of free metal ions into bound forms results in a reduction of their toxicity to living organisms. In contrast with resistant ones, tolerant algae enable the absorption and accumulation of heavy metals inside their cells in metabolic-dependent processes that are largely influenced by energy sources and by environmental conditions. Bioconcentration of heavy metals usually involves active uptake through membrane transporters of cationic micronutrients, followed by detoxification achieved with binding metal ions to chelating agents such as phytochelatins, metallothioneins, polyphosphate inclusions, several organic acids and


The Use of Green Microalgal Cultures for Bioremediation of Freshwater …

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amino acids, produced in metabolic processes induced or up-regulated by the presence of heavy metals [5, 32, 45, 51, 96, 114, 141, 150, 181]. Hexavalent chromium, nickel and cadmium are frequent contaminants of continental freshwater ecosystems affected by human activities, especially by mining, combustion of conventional fuels, metallurgy. Even micromolar concentrations of these heavy metals induce direct alterations in enzyme activities, in photochemical processes and in biomembrane functions, affecting algal growth, metabolism and reproduction. On the other hand, by competition with cationic essential trace elements, they have an indirect negative influence on algae by causing mineral nutrient deficiency syndromes, as well as by enhancing oxidative stress, thus impairing ion homeostasis and redox state in different algal cell compartments. In this context, identification of sensitive biochemical and physiological markers which indicate the impact of heavy metal pollution on algal vital processes may be very useful in the wastewater treatment technologies which rely on the remediative capacity of microalgae. And because usually industrial and domestic wastewaters do not contain only one type of heavy metal, antagonistic or synergistic (competitive or cooperative) interactions between different heavy metals are also of great interest in determining the bioremediative efficiency of microalgae in case of cocontamination. From among the various physiological markers of integrated algal reactions to heavy metal toxicity, certain parameters of the induced chlorophyll fluorescence, of the photosynthetic pigment content and of algal growth and reproduction, evaluated simultaneously, may be useful tools in our attempt to optimize biological treatment of heavy metal-contaminated water in an environmental-friendly approach, without using xenobiotic chemicals or genetically modified organisms [18, 52, 63, 120, 130, 145, 173, 180]. The aim of this chapter is to summarize the possible benefits of using microalgae in bioremediation of water polluted with certain heavy metals, as well as to reveal physiological changes induced by different concentrations of chromium, nickel and cadmium in a green microalga which is ubiquitous in freshwater environments, in order to optimize bioindication and remediation of water pollution by using the capacity of this alga to survive and reproduce in heavy metal-contaminated environments, and to bioaccumulate and sequester heavy metals during a given exposure time.

ADVANTAGES OF MICROALGAE IN BIOINDICATION AND REMEDIATION OF WATER POLLUTION Microalgae have useful characteristics of both plants and micro-organisms such as: 1. Because of their basically autotrophic way of life, microalgae incorporate external light energy in their organic compounds, they do not depend on the


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2.

3.

4.

5.

6.

existence of organic nutrients, and photosynthetic carbon assimilation enables a wide plasticity to their metabolism, i.e., algae, like higher plants, are able to produce a wide range of organic substances, primary (ubiquitous) and secondary (specific) metabolites. Autotrophy confers algae a crucial role in the entire energy flux and primary production of aquatic ecosystems. Changes in composition of algal community directly influence all the consumers and decomposers in the trophic network of water bodies [110]. In contrast with higher plants and seaweeds, microalgae do not spend energy for sustained growth of their individual body, for skeletal structures, for nutrient storage sites and for internal long-distance transport systems, a higher part of their energy being invested in reproduction, in formation of new generations of individuals within short periods of time. Being eukaryotic micro-organisms, microalgae exhibit a high adaptive capacity to changes of environmental conditions, and because of their high reproductive rate, they easily occupy new habitats and spread through the aeroplankton from one water body to another. Many microalgae reproduce asexually, which enables formation of genetically homogenous populations. This feature is convenient when bioindication of changes in water quality is made by registering modulation of physiological processes, because genetic differences between individuals would lead to large deviations from the mean reaction of the whole population consisting of billions of algal organisms. Due to their high number and ubiquitous presence in different aquatic environments, they enable biomonitoring and remediation protocols under standardized conditions, which is important in the context of their general use in the management of water quality [177]. Also because of the high number of individuals in a given volume of aquatic environment, it is easy to select new varieties or certain mutants formed randomly under the selective pressure of extreme life conditions related to accumulation of polluting agents with adverse effects on metabolic and developmental processes. Microalgae react very quickly to adverse environmental impacts, being able to make directed changes in their physiological processes disturbed by external stressors. After perception of stress agent and a transitory (acute) decline of sensitive processes, a longer phase of hardening begins, new steady-state levels of the affected functions are set, and enhanced tolerance leads to long-term acclimation to the modified life condition caused by eutrophication, deversion of pesticides or other organic xenobiotics, or increased concentrations of watersoluble heavy metals. This functional adaptation of algal individuals to adverse conditions enables their persistance and reproduction in various polluted waters, as a prerequisite for an efficient remediation. Their quick reaction to environ-



7.

8.

9.

10.

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mental changes makes possible an early detection of polluting agents, which is important for prompt interventions aiming diminuation of pollution. Being widespread in practically every aquatic habitat, microalgae exhibit a great diversity, which facilitates selection of suitable forms for remediation of water pollution with very different chemical agents and their combinations, under diverse ecological conditions. There are more than 21000 algal species known at present, with more than 100000 intraspecific varieties and cultivated cell lines (strains), while many other algae are probably not known yet. This means that algae represent a vast potential biological resource for different biotechnological applications, including wastewater treatment and remediation of degraded aquatic ecosystems. It is estimated that about 70% of the biomass on Earth is given by algae, most of them belonging to microalgae (microscopic algae, individuals consisting of single cells or associations of unicellular, ontogenetically related algae). This means that algae are fundamental components of aquatic living communities even from a quantitative approach, being the most abundant submersed organisms which may extract and accumulate pollutants from the aqueous solution [86]. Due to their plastic metabolism on which the acclimative capacity is based, microalgae are able not only to accumulate and sequester water pollutants, but also to biotransform them into less harmful derivatives, thus performing a detoxification of the aquatic habitat. This capacity is a key feature for biological epuration of wastewater, when autotrophic microalgae are used in combination with heterotrophic bacteria. Depending on the chemical nature of water pollutants, several microalgal species can modify they nutrition and intermediary metabolism, in order to consume the polluting compounds and to convert them in parts of their natural biomass. Many microalgae are able to change their nutrition from photoautotrophy to photoheterotrophy (e.g., in presence of simple organic compounds), to mixotrophy (use of both inorganic and organic carbon sources), to amphitrophy (alternation of autotrophic and heterotrophic periods) or to auxotrophy (autotrophy combined with use of certain organic substances) [38, 89, 176]. Microalgae need basicly sunlight, water and small amounts of mineral nutrients for growth and reproduction. Planktonic species are equally present in the whole volume of the water body to the depth of light penetration. They easily develop high tolerance to the presence of various polluting agents, they accumulate, sequester and even detoxify higher amounts of chemicals. All these features ensure that bioremediation of water pollution by use of microalgae can be costeffective and environmental friendly [165]. Wastewater treatment with microalgae can be combined with the reduction of air pollution caused by large scale emissions of carbon dioxide, because algal


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Laszlo Fodorpataki, Sebastian R. C. Plugaru, Katalin Molnar et al. systems are very efficient converters of carbon dioxide into useful organic compounds resulting from the photosynthetic carbon assimilation. In parallel with remediation of polluted water, the resulting algal biomass may be a good source of proteins (the so-called single cell proteins), vitamins and essential micronutrients in aquaculture and in animal farming (if the algal biomass does not contain toxic compounds), or it may be used for extraction and recycling of valuable substances adsorbed or absorbed by the algal cells (e.g., rare metals). Algal biomass from wastewater oxidation or stabilization ponds may be also used for production of methane, bioethanol or biodiesel as alternative, regenerable sources of fuel [6].

The best algal types for bioindication and remediation of water pollution are those with large tolerance to variations of temperature, light regime, pH, osmotic potential and dissolved oxygen concentrations, with low sensitivity to photoinhibition and to photooxidative stress, with low compensation point of the photosynthesis for the watersoluble inorganic carbon source, and with a pronounced capacity for biosynthesis of protective metabolites induced by specific water-polluting agents [156].

LIMITATIONS OF USING MICROALGAE IN MONITORING AND REDUCING CONTAMINATION OF AQUATIC ENVIRONMENTS Beside their several advantages in early indication and cost-effective remediation of water pollution, microalgae also possess following characteristics which limit their use in wastewater treatment [14, 47, 64, 95, 123, 127]. 1. The ability of microalgae to indicate by internal changes the nature and degree of water pollution, as well as their capacity to accumulate and immobilize harmful chemicals, largely depend on the availability of essential nutrients and of light energy source, because starving algal cells are much more sensitive to environmental stress factors, and energetic limitation impairs hardening and biosynthesis of protective metabolites. For example, several microalgal species accumulate much less cadmium ions and are more sensitive to cadmium toxicity under conditions of phosphate deficiency than in eutrophicated aquatic habitats in which phosphate concentration is high. 2. Repeated or long-lasting exposure to polluting agents enhances the development of tolerance to the chemical stress factor, so algae become less and less sensitive, and their indicative capacity decreases. This is why exposure time greatly influences the remediative potential of microalgae.



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3. Resistance or tolerance of vital physiological processes to polluting agents has its limits, thus algae cannot be used in remediation of aquatic environments heavily polluted with high concentrations of chemicals. 4. Because they are at the base of trophic pyramids, algae have a lower bioconcentration capacity than their primary and secondary consumers. 5. Physiological and biochemical markers used to select more efficient algae in bioindication and phytoremediation may vary in similar extent and direction not only under the influence of water pollutants, but also upon variations of natural environmental factors, and this fact may result in false positive responses which may be misleading when one evaluates the degree of pollution and the capacity of algae to tolerate chemical stress conditions. For example, the quantity of different photosynthetic pigments (chlorophyll-a, chlorophyll-b, carotenes, different xanthophylls) and the ratio between these pigments is usually a very useful marker of algal tolerance or sensitivity to adverse growth conditions, but may considerably vary even in non-polluted water if radical changes occur in photon flux density. This is the main reason why the simultaneous evaluation of several stress markers is recommended for a better accuracy of the evaluation of impacts of polluting agents on algal metabolism and development. Elimination of false positive reactions can be also achieved by using algal cell cultures grown under controlled conditions, where the most important environmental factors are kept in the optimal range. 6. When algal biomass developed during wastewater treatment is intended to be used for recycling of useful compounds or for biofuel production, harvesting processes and species control will increase the overall costs of remediation.

BIOLOGICAL PROCESSES INVOLVED IN TREATMENT OF WATER POLLUTED WITH HEAVY METALS Phytoremediation of heavy metal-contaminated aquatic habitats with microalgae exhibiting plant-type metabolism and stress reactions, represents the use of algae to extract, to sequester and sometimes to detoxify metal pollutants dissolved in the aqueous medium in which algae grow and reproduce, in order to clean-up the water environment. The use of planktonic microalgae for removal of heavy metals from contaminated waters has a lower cost than other remediation technologies, it may generate algal residue rich in recyclable metals, causes only minimal environmental disturbance, and because of these advantages has a large public acceptance. It may be achieved by different processes, based on distinct physiological properties of selected algal species. Thus, a better knowledge of biochemical mechanisms on which remediation relies is highly beneficial


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for the optimization of water treatment technologies based on algal abilities implied in natural, biological reduction of water pollution. Most of the algal species used in remediation of heavy metal-contaminated water are green microalgae belonging to Chlorococcales (with thick cell wall and without flagellum), and a few flagellate species included in Volvocales and Euglenales. Diatoms are more sensitive to higher concentrations of heavy metals, but they may be used for an early indication of water pollution with small amounts of metal ions [74, 83]. Biosorption of heavy metals by algae involves chemical and physical processes resulting in binding of heavy metals to cell wall structures or to secreted polysaccharides and mucilage sheaths. The heavy metals fixed on algal cell surfaces are removed from the aqueous solution, thus biosorption is associated with stabilization and sequestration of heavy metals, without entering in the living cells. Biosorption is achieved mainly by resistant microalgal species and also by dead algal cells. The more extended relative surface of algae (cell shapes distant from spherical, with extensions and emerging patterns) enables a better biosorption, and the process is also largely influences by thickness and chemical composition of algal cell wall [56, 100, 200]. Active uptake and accumulation of heavy metal ions through energy-driven membrane transport leads to extraction and bioconcentration of these metals. Living algal cells actively accumulate heavy metals against their electrochemical gradient, mainly through membrane transporters of cationic micronutrients which are less specific for one certain chemical element, but transport several, chemically similar metal ions. For example, iron transporters may allow uptake of cadmium, chromium, nickel and other water-polluting heavy metals into the algal cells. Intracellular bioconcentration of heavy metals is linked to chemical energy (ATP) consumption, so it has to be supported by metabolic processes which store energy reserves. This is why enhanced aerobic respiration is usually associated with accumulation of high amounts of heavy metals, while growth rate and synthesis of organic storage metabolites are down-regulated, in order to fuel more energy to the active membrane transport of metal ions. Not only the plasma membrane which surrounds the algal cells, but also the internal membranes which delimit different organelles from the cytosol exhibit selective permeability, thus the heavy metals are distributed among the different cell compartments. In algal cells, which usually do not develop a large central vacuole, most of heavy metals are accumulated in the cytosol, in the stroma of chloroplasts and in the matrix of mitochondria. Hyperaccumulation of harmful heavy metals has probably an adaptive significance, because it may have been evolved in order to confer chemical protection against consumers and parasites, including pathogenic micro-organisms. Active accumulation of heavy metals by the algal cells contributes to phytoextraction, bioconcentration and sequestration processes [30, 59, 70, 98]. The main algal process in phytosequestration of heavy metals is complexation with organic chelating agents produced by algal metabolism upon sensing the increased



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concentration of certain heavy metals inside the cells. Phytochelatins and metallothioneins (inducible peptides with several cystein residues that serve to bind heavy metal ions to their sulfur atom) are the most specific chelating agents used by algae and higher plants to bind heavy metals. The presence of high amounts of heavy metals in the cytosol activates the phytochelatin synthase enzyme, which produced phytochelatin molecules from glutathione residues. Heavy metals also activate the genes which encode for metallothioneins, thus these peptides are newly synthetized in ribosomes and accumulate in the cytosol in an extent which is related to the quantity of heavy metal ions reaching the cytoplasm. Chelated heavy metals may be transported and sequestered in different cell organelles, leading to a compartmented distribution of metal ions. In many algae, polyphosphate bodies, which primarily serve as reserves of phosphorus, magnesium, calcium, potassiun and other essential metal nutrients, may bind heavy metal ions, thus reducing their concentration in the aqueous phase of the intracellular space. Algae may also bind heavy metal ions to organic acids such as malic acid and citric acid, while some free amino acids may also serve to stabilize certain metal ions (e.g., nickel is frequently bound in algae to histidine). Because heavy metals complexed with chelating agents cannot move freely in the algal cells, this way of sequestration can be also considered as a mean of detoxification [31, 57, 140, 183]. Detoxification of heavy metals may be achieved not only through complexation with chelating agents, but also by biochemical reactions of the algal metabolism. Enzymatic change of the oxidation state may result in bioconversion to a less toxic metal form. For example, the hexavalent chromium is quickly reduced in algal cells to trivalent chromium ion, which being insoluble in water has a considerably reduced toxicity. Similar detoxification processes are represented by conversion of mercury (II) to Hg 0, of gold (III) to Au0 or of molibdenum (VI) to Mo(III). Change in redox state may be also considered a way of biotransformation of metal form. Enzymatic methylation of heavy metals in the algal metabolism is another way of detoxification, because methylation prevents heavy metals from reaction with sulfhydryl groups of vital protein molecules, thus avoiding structural damage, denaturation and inactivation of thes proteins (enzymes, membrane transporters, transcription factors, etc.). Enzymatic methylation is also a form of biotransformation [59, 164, 195]. In case of a limited number of heavy metals and metalloids, methylation results in generation of volatile products, which leave the cells and reach the atmosphere, being removed from the aquatic medium. This is the case of methylated mercury, arsenic, bismuth and antimony, and the process is known as phytovolatilization [202]. Phytostabilization implies reduction of the mobility of heavy metals in the aquatic environment, as well as bioconversion of metals to insoluble forms which precipitate from the aqueous solution. Precipitation of insoluble metal complexes with reduced mobility and bioavailability is usually performed by conversion into sulfides (e.g., in case of chromium and aluminium). Binding on cell surfaces and to extracellular products, as it


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was mentioned earlier, also results in stabilization of heavy metals and in limitation of their toxic effects. Concerning the physico-chemical characteristics of the aqueous environment, the general trend is that increase in pH (water becoming more more alcaline) and decrease in redox potential (water becoming more reducing) result in decreased availability of heavy metals for algae [71, 135, 138, 144]. Different microalgal processes related to resistance or tolerance to heavy metal contamination of the water make possible several means of phytoremediation, these processes being interrelated and forming an integrated network in which the different components complement each other (Figure 1).

Figure 1. Microalgal processes involved in different ways of phytoremediation of heavy metal-polluted aquatic environments (original).

Concerning their behavior related to heavy metal contamination of the aquatic habitat, algal species may be metal excluders, pseudometallophytes, metal indicators, moderate accumulators or hyperaccumulators. 1. Metal excluders are resistant algal species which prevent heavy metals from entering their metabolism, thus avoiding harmful effects of metal toxicity. They may contribute to biosorption and surface stabilization of metal ions, but do not biotransform and sequester intracellularly these contaminants. They usually start to experience toxic effects only at high heavy metal concentrations, and physiological changes are not proportional with the metal content of polluted water [16, 188].



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2. Pseudometallophytes include algae which survive and reproduced in severely metal-contaminated water without bioconcentrating the metal ions. They may contribute to stabilization of metals by reducing their mobility, e.g., through precipitation of insoluble metal complexes or through metal fixation to extracellular biopolimers, heavy metals being converted into biologically inert forms [27]. 3. Heavy metal indicator algae are those in which metal content approximately reflects the metal level in the surrounding water. They do not exclude and do not bioconcentrate these heavy metals, but they establish an equilibrium between external and internal metal content. Their phytochelatin content may be directly corelated with the degree of heavy metal pollution, thus being suitable for bioindication and even for long-term monitoring of contamination level of the aqueous medium. They are neither resistant nor very tolerant to heavy metal toxicity, and many of their physiological processes exhibit proportional changes with the degree of water pollution. They still tolerate contaminants to a greater extent than non-specialist, sensitive species or intraspecific varieties [29, 54, 87]. 4. Moderate accumulators are able to bioconcentrate heavy metals inside their cells in ten to a few hundred times higher concentrations than in the surrounding water, but only if the effective metal quantity does not exceed the micromolar range. They are suitable for extraction and sequestration of heavy metals from moderately polluted aquatic environments, and they perform an active transport of metal ions through less specific membrane transporters of cations, which primarily serve for uptake of essential inorganic micronutrients. Many microalgal species with heavy metal tolerance are included in this category [3, 26, 73, 85]. 5. Hyperaccumulator algae can concentrate heavy metals to levels far exceeding those in their aquatic habitat, and they experience little or no toxic effects because they efficiently sequester these metals through complexation with chelating agents produced in high quantity in their metabolism, or they detoxify metals by enzymatic reactions. In their case, bioconcentration factor for heavy metals may reach the level of one thousand. Generally, algae which may contain more than 0.1% of chromium, nickel, cadmium, copper, mercury, silver or more than 1% zinc or manganese on a dry weight basis are considered hyperaccumulators, irrespective of the concentration of heavy metal in the aqueous solution. A given hyperaccumulator algal species generally may bioaccumulate to high extent only one type of heavy metal, so a hyperaccumulator alga will never extract in the same extent different heavy metals from the contaminated water. The extent of bioconcentration largely depends on the prevailing nutritional, illumination and other growth conditions, as well as on possible interactions among co-contaminating chemicals. Under adequate developmental


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Laszlo Fodorpataki, Sebastian R. C. Plugaru, Katalin Molnar et al. conditions, hyperaccumulators give the most efficient remediation of heavy metal-polluted water [75, 101, 143, 146, 154, 168, 190].

USE OF IMMOBILIZED MICROALGAE IN REMEDIATION OF HEAVY METAL-CONTAMINATED WASTEWATER Industrial and urban wastewaters may contain high amounts (over 1 mg L-1) of different heavy metals, which at these concentrations exhibit ecotoxicity, and their removal is a priority in wastewater treatment. Several planktonic microalgae are capable to accumulate heavy metal ions, by biosorption to cell wall structures and by active membrane transport into the intracellular compartments, with a bioconcentration factor of several hundreds or even a thousand (i.e., the concentration of heavy metal becomes hundreds-thousand times higher in the algal biomass than in the surrounding water solution). If algal cells are harvested, even the isolation and recovery of heavy metals is possible (e.g., for gold and uranium). Algae start very quickly to accumulate and immobilize heavy metals first by a passive process resulting in adsorption to surface structures (adhesion to colloidal particles of mucilage sheath or secreted biopolimers, coordinative or electrostatical binding to cell wall structures), then a slower, but metabolically active uptake process follows (absorption into the living cells), which reaches its maximum in a few hours. The highest degree of heavy metal accumulation occurs within the first day of exposure, this rate is maintained over the next days with no more enhancement, and finally the bioaccumulative capacity of living algal cells decreases in time, first moderately, then more and more abruptly, during the next one or two weeks. If algae die because of heavy metal toxicity, they may exhibit for a short period of time a greater binding capacity for the metal ions, probably because biomembranes loose their selectivity and more intracellular binding sites may be accessible for the metals. In statical wastewater treatment systems (where contaminated water is supplied only at the beginning of the cleaning process) it takes usually two weeks for microalgae to decrease the heavy metal content of water with 50%, while in dynamic systems (with periodic supply of new wastewater quantities) this decrement is achieved within three weeks. In some cases, inorganic nitrogen and phosphorus source has to be supplied and the pH value of the medium has to be adjusted in order to maintain the high bioaccumulation capacity of algae over several weeks. Heavy metalresistant algal species and strains, as well as dead algal cells, extract and immobilize heavy metals from the aqueous environment through adsorption to surface structures, but heavy metals are not concentrated intracellularly. In contrast, heavy metal tolerant algae actively transport heavy metals through the less specific micronutrient transporters of their



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membranes and bind these metal ions to intracellular organic metabolites, in order to immobilize and detoxify them [69, 112, 159]. If heavy metal concentration is high, algal cells become intoxicated and cease further metal uptake. Instead, their decomposition release a part of the formerly sequestered metals back in the water environment. It was observed that microalgae which are able to produce extracellular polysaccharides and form a mucilage sheath around their cells, are better protected against toxic effects and tolerate higher concentrations of heavy metals. In analogy with this natural phenomenon, different gel particles were created in which algal cells may be embedded, thus being protected from the sudden contact with polluting agents, and being able to perform water decontamination for a prolonged period of time. For metal ion removal and recovery, algal cells may be entrapped in a matrix of insoluble calcium alginate, polyacrylamide gel (based on the free-radical polymerization of acrylamide in algal cell suspension), or silica gel. The matrix has pores that enable the metal ion diffusion into the beads, where interactions with the algal cells may occur progressively. For example, immobilized Chlorella cells can be used efficiently to extract and recover gold, uranium, copper, lead, nickel, chromium from contaminated water containing up to 0.1 mM of dissolved heavy metal. Heavy metal binding to immobilized algal cells is usually enhanced by increased temperature (up to 30°C with living cells and up to 50°C when embedded algae are dead). Conversion of chromium(VI) to the less toxic chromium(III) form, as a useful detoxification method, achievable with immobilized algae in a more cost-effective and environmental-friendly way than the method which uses chemical reagents, may also be enhanced with increasing temperature in the aqueous medium of the immobilized algal cells. pH has also a significant influence on the metal-binding capacity of embedded microalgae, and one of the most important limiting factors of heavy metal removal with immobilized algae is that the optimal pH for heavy metal uptake by algal cells is different (more acidic) from the optimal pH of algal growth and reproduction. Cationic heavy metals (e.g., cadmium, nickel, copper, uranium, lead, zinc) may be successfully bioaccumulated by embedded living algae at pH values between 5-6, but anionic forms (e.g., gold cyanide, platinum chloride, chromate, molibdenate) can be immobilized only at very low pH (around 2) with destroyed algal cells embedded in gel particles. In case of silver and mercury, binding to immobilized algae was found to be rather independent of pH [69, 99]. Because effluents from a column with adsorbent particles usually have lower residual heavy metal concentrations than the supernatant of a batch reactor, it may be concluded that incorporation of algal cells into a gel matrix which can be introduced in a packed column is a practical approach for alga-mediated heavy metal extraction and recovery from polluted water. Because the metal binding properties of different algae vary greatly, and the immobilization methods applicable for wastewater treatment are also numerous, one can predict that embedded microalgae have a high potential in heavy metal extraction and recovery from contaminated aquatic environments [84, 201].


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INFLUENCE OF CHROMIUM ON ALGAL DEVELOPMENTAL AND METABOLIC PROCESSES Heavy metals exert some common influences on algal physiological processes. These may be a) direct effects related to their high affinity to sulfhydryl groups in proteins, thus causing disturbances in the function of enzymes, membrane transporters, receptors, transcription factors and other regulatory proteins, and b) indirect effects resulting from uptake competition with essential metal micro-nutrients, as well as from integrated results of their accumulation on growth, cell division rate and balance between bioregulators. Beside these general impacts on vital functions, the different heavy metals exert specific influences on distinct metabolic and developmental steps. All these effects determine the capacity of different algae to react to water pollution with these metals, to sequester and to detoxify them, and finally the entire success of bioremediation procedures [68]. Chromium impact on aquatic ecosystems is poorely documented, in comparison with the other heavy metals which may accumulate in water ponds exposed to anthropogenic influences. It is a non-essential element for algal nutrition, in small concentrations it may stimulate the enzymatic activity of phosphoglucomutase, thus influencing glucose metabolism. Chromium ions may accumulate in aquatic habitats from production of refractory materials, from metallurgy (production of stainless steel and chrome-plated metals) or from chemical manufacturing (e.g., production of pigments for textile industry) [15]. Hexavalent chromium is toxic because of its high oxidation potential and its ability to cross biological membranes, probably through anion channels, because chromate and bichromate show chemical similarity with phosphate and sulfate anions. Inside the algal cells, hexavalent chromium is quickly reduced to trivalent chromium (in a process which requires reduced glutathione, ascorbic acid and NADPH or NADH), which is not soluble in water, ant thus its toxicity is much lower. This is why biological reduction of hexavalent chromium to the trivalent form may be considered a process of detoxification and stabilization. Inside the algal cells most of the chromium is associated with heat-stable proteins and peptides of the cytosol and of the internal compartment of organelles (it may be incorporated in metal-rich granules), a smaller part being adsorbed to the cell wall [1, 50, 103]. According to our present knowledge, chromium does not have any specific transporter and uptake mechanism in algae, and hexavalent chromium is taken up actively through the plasma membrane, the energy being supplied by the proton-ATPase pump [39]. The harmful effect of high amounts of chromium is based on several actions. It inhibits cell division and reduces dry matter production. It may decrease chlorophyll content because of inhibition of aminolevulinic acid dehydratase and protochlorophyllide reductase involved in chlorophyll synthesis. It reduces the size of peripheral lightharvesting antennae around photosystems in the thylakoid membranes of chloroplasts,



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thus leading to a decreased chlorophyll a/b ratio in green algae. In high concentrations it causes a decline in carbon dioxide assimilation because it decreases the cabroxylase activity of the photosynthetic key enzyme, the ribulose-1,5-bisphosphate carboxylase oxygenase (Rubisco). It may also bind to cytochrome a3, thus interfering with mitochondrial respiratory electron transport towards the oxygen molecule. There are also reports on decreased malate dehydrogenase and amylase activities and enhanced protease activity upon exposure to water pollution with hexavalent chromium. Because it favors the generation of harmful hydroxyl radicals, it triggers higher ascorbate peroxidase, superoxide dismutase and glutathione reductase activities during the antioxidative defense processes, and it stimulates biosynthesis of vitamin E as a main anihilator of hydroxyl radicals [78, 134, 148]. If oxidative damage induced by high chromium concentrations persists, peroxidation of unsaturated fatty acids in membrane lipids occurs, and the first protective enzyme which exhibits a strong decline it its activity is usually catalase. The marked decline of reduced glutathione pool of algal cells under chromium toxicity may be due to its oxidation in processes of antioxidative defense, and also to consumption of glutathione in the synthesis of phytochelatins which will sequester and inactivate chromium ions by complexation. Exposure to high concentrations of chromium may also lead to genotoxicity, because this heavy metal interacts with DNA, causing local mutations, hypermethylation of pyrimidines, mitotic irregularities, undivided nuclei, chromosome fragmentation, and overall increase in DNA polymorphism [24, 65, 151, 162, 169, 174, 186, 193]. The most obvious indirect effect of chromium toxicity is related to its uptake antagonism with several essential mineral nutrients. It competes for surface sites and transport binding mainly with iron, manganese, phosphorus and sulfur, it adversely affects nitrogen assimilation, and it inhibits the activity of plasma membrane proton ATPase, thus impairing uptake of every mineral nutrient which enters algal cells through active membrane transport [23, 61, 125, 128, 152, 191]. Scenedesmus and Chlorella species (freshwater green microalgae) were able to tolerate relatively high amounts of chromium in polluted water, and during several weeks of exposure they accumulated chromium (both by adsorption to cell wall and by absorbtion inside the cells) with a bioconcentration factor ranging between 230 and 320, when the initial hexavalent chromium content of water was 500 µg L-1 [106, 151].

CHANGES INDUCED BY NICKEL IN ALGAL CELLS Nickel is considered by many scientists an essential micronutrient for algae and higher plants [4, 22, 25, 48, 49, 104, 199], and only some of them consider that nickel is not essential, but it may be in certain cases beneficial for plant metabolism [124]. Argumentation of its essentiality is based on its role in the enzymatic activity of urease,


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hydrogenase, glyoxalases, peptide deformylases and methyl-CoA reductase. But these enzymes do not have essential physiological functions or even do not exist in most of the algae and generally in plant cells. For example, urease is not needed for the normal nitrogen metabolism of plants, because they use inorganic nitrogen source in form of nitrate or ammonium ions. If no inorganic nitrogen source is available and urea is present in the surrounding aqueous medium, algae and higher plants switch to nitrogen heterotrophy and use urease (which needs nickel) to decompose this organic nitrogen source to ammonium and carbon dioxide. Hydrogen metabolism, methane production and acetogenesis, in which some of the above mentioned, nickel-containing enzymes are involved, are not characteristic for plant metabolism. This is why nickel cannot be considered an essential nimeral nutrient for algae and higher plants, even if it is less toxic than most of the other heavy metals which may pollute water due to human activities. Its critical toxicity levels are above 10 mg kg-1 dry weight in sensitive species, above 50 mg kg-1 dry biomass in tolerant species and strains, and above 500 mg kg-1 dry weight in hyperaccumulators. Nickel toxicity is usually reflected by inhibition of mitotic cell divisions, by iron, zinc, manganese and copper deficiency symptoms, by decreased chlorophyll and carotenoid pigment content, by impairment of light-driven electron transport in the core complexes of photosystem I and photosystem II [19, 36, 43], by depletion of oxygen production in the water-splitting complex (if nickel accumulates in chloroplasts in micromolar concentrations). Since nickel ions have a stable oxidation state, nickel is not a redox-active metal, so it cannot directly induce oxidative stress by enhancing formation of reactive oxygen species. Still, upon extended exposure to high nickel levels, increases in the concentrations of superoxide anion, hydrogen peroxide, hydroxyl radicals and nitric oxide radicals may be experienced by algal cells, most probably because nickel reduces the activity of enzymes involved in the antioxidative protection. This is why nickel toxicity can induce depletion of cellular glutathione pool and can cause membrane damage through lipid peroxidation [12, 13, 20, 40, 42, 58, 76, 105, 111, 126, 153]. Nickel ions entering the aquatic ecosystems originate mainly from combustion of coal and gasoline, from mining activities, smelting, vehicle emissions, sewage sludge, alloy manufacture, electroplating, disposal of electrical batteries. Because nickel shares the same transport system with copper and zinc to enter the algal cells, its increasing amount in the aqueous solution inhibits absorption of these essential trace elements. In may also exhibit uptake competition with iron, manganese and calcium, while high amounts of magnesium may reduce its intracellular accumulation in toxic concentrations. On the other hand, high nickel levels may prevent uptake of increased amounts of more toxic heavy metals from co-contaminated water, such as cadmium and mercury [2, 10, 60, 88, 94, 149, 166, 175, 182, 203]. In case of microalgae belonging to Scenedesmus, Chlamydomonas and Euglena genera, it was found that the nickel concentration which causes death of 50% of algal individuals is about an order of magnitude or more higher



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than in case of cadmium. Toxicity is related to the free divalent nickel ion, while the main sequestered form is bound to histidine, nicotianamine, citric and malic acid. The optimal pH for extraction and bioaccumulation of nickel from polluted water is slightly alcaline, being situated aroun the value of 8, and the bioconcentration factor varies between 400 and 3000, depending on species, growth conditions and external nickel concentration [7, 9, 11, 17, 21, 77, 107, 131, 160, 167, 196]. Elevated concentrations of free histidine and of metallothioneins, as well as increased serine decarboxylase activity are useful biochemical markers for selection of nickel-tolerant and hyperaccumulator algal species, in order to introduce them in a “green” technology for treatment ow water polluted with nickel.

EFFECTS OF CADMIUM ON ALGAL METABOLISM AND GROWTH Cadmium has no physiological function in plants, so its toxicity manifests even at very low concentrations. It may get into aquatic environments mainly from plastic stabilizers, fertilizers, paints, batteries and electroplating [35, 142, 157]. Toxic effects of cadmium rely on its influence on many metabolic processes. It inhibits the activity of several enzymes. For example, carboxylase activity of Rubisco in the stroma of chloroplasts is impaired, but its oxygenase activity becomes enhanced, thus carbon dioxide assimilation rate in the Calvin cycle decreases, but photorespiration associated with carbon assimilation becomes more intense, this resulting in a lower net biomass production. Photosynthetic pigment content is also altered in the presence of cadmium, this fact leading to decreased efficiency of light energy harvesting. Because synthesis and degradation of different pigment types is influenced in different degrees, modifications of chlorophyll-a to chlorophyll-b ratio and of total chlorophylls (a + b) to total carotenoids (carotenes+xanthophylls) ratios are usually much more sensitive markers of negative effects of cadmium than the absolute quantities of the different photosynthetic pigments. For example, under constant photon flux density and controlled growth conditions, in statical cultures, the ratios between the main types of photosynthetic pigments, determined spectrophotometrically after extraction with dimethylformamide, exhibited in the green microalga Scenedesmus acuminatus distinctive variations in relation with the concentration of cadmium in the polluted water medium (Figure 2). Chlorophyll a/b ratio moderately, but significantly increased upon exposure for 5 days to 50 µM and 500 µM cadmium, indicating that chlorophyll-b content decreased in a higher extent than chlorophyll-a content (overall chlorophyll content was lower in the presence of cadmium as compared to control populations – data not shown).


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Figure 2. Photosynthetic pigment ratios in the green microalga Scenedesmus acuminatus, after 5 days of exposure to different concentrations of cadmium, under constant illumination and controlled growth conditions. chl. – chlorophylls; car. – carotenoids (n = 5, vertical bars represent means ± SE, different letters indicate statistically significant differences at P < 0.05 between components of the same series, according to post-ANOVA Tukey HSD test).

This may be explained by the restriction of peripheric light-harvesting antennae around the photosystems, where there is a higher number of chlorophyll-b molecules than in the inner antennae. In contrast with chlorophyll a/b ratio, the chlorophylls to carotenoids ratio decreases progressively with the increased severity of cadmium pollution, because chlorophyll content is impaired in a higher extent than carotenoid pigment content. In the photosynthetic apparatus, cadmium toxicity may also result in inactivation of the water-splitting complex and in inhibition of electron transport on the acceptor side of photosystem I, as well as in damage to the thylakoid membranes because of stimulation of lipoxygenase activity. The reduced assimilation rate of carbon dioxide due to Rubisco inhibition leads to feed-back down-regulation of the entire light phase of photosynthesis, because chemical energy resulting from photochemical reactions cannot be used in synthesis of new organic compounds in the Calvin cycle [37, 147, 189]. An indirect harmful effect of cadmium is induction of iron, zinc, manganese and copper deficiency, because it competes with these essential trace elements for the same membrane transporters (mainly via the high-affinity, nonspecific zinc-iron protein transporter family). It also disturbes calcium homeostasis of plant cells and inhibits nitrate reductase in the nitrogen assimilation pathway [197]. High amounts of cadmium may also inactivate DNA dismatch repair processes, thus exerting a mutagenic effect. Another indirect effect of cadmium in plant cells is oxidative damage of proteins, membrane lipids and nucleic acids. In sensitive species a decrease of antioxidants may be



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observed, while development of tolerance in related with a significant increment of protective antioxidants [33, 67, 91, 136]. Cadmium is one of the most effective activators of phytochelatine synthase, which is a constitutively expressed enzyme in plant cells, but requires post-translational activation by heavy metals that accumulate in the cytosol. Because the favored ligands of cadmium are thiols, cadmium ions are mainly complexed and sequestered by phytochelatins, which confer tolerance to cadmium toxicity in tight connection with up-regulation of sulfur metabolism [8, 28, 44, 121, 158]. Chlorella, Chlamydomonas, Euglena and Scenedesmus species were found to accumulate cadmium with a bioconcentration factor between 100 and 200, when initial cadmium content of the polluted water ranged between 5 and 100 µg L-1. It was also established that a more efficient cadmium ion removal may be achieved by adsorption to dead algal cells embedded in gel particles [81, 118, 187].

PHYSIOLOGICAL MARKERS OF THE IMPACT OF CHROMIUM, NICKEL AND CADMIUM ON MICROALGAE GROWN IN POLLUTED WATER Presence of heavy metals in polluted water is perceived as a disturbing condition which initiates stress reactions that lead to an increased chance of survival and further development. During these stress reactions coordinated physiological modifications occur, which indicate the capacity and degree of tolerance achieved through individual, reversible acclimation to the worse living conditions. These functional changes may be valuable markers of the effects of heavy metals on algae if we succeed to reveal the cause-effect relations that underly acclimative responses. The most common structural indicator of stress condition is algal biomass, which may be evaluated based on cell density of populations, ash-free dry weight, chlorophyll-a content or biovolume. Productivity is a useful functional marker, and may be determined through biomass change in a given period of time, through carbon assimilation rate and through oxygen production in light. Because variation in oxygen content in a given volume of algal culture may be measured both in darkness and during constant illumination, it enables determination of net production, brut photosynthetic production and respiration rate. Dynamics of photosynthetic pigment content, inhibition or enhancement of the catalytic activity of several protective enzymes, degree of membrane lipid peroxidation, synthesis of specific metabolites (e.g., phytochelatins), and chlorophyll fluorescence parameters related to energetic efficiency of light use in photochemical processes, may all be valuable biochemical and physiological markers for a correct evaluation of algal reactions to heavy metal stress, as well as for selection of algae and adjustment of growth conditions in order to optimize remediation of heavy metal-polluted aquatic environments using selected microalgae [72, 80, 108, 119, 124, 133, 137, 161, 170, 178, 192, 204].


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In order to exclude interference of other external factors with heavy metals, it is preferable to study the above-mentioned functional parameters under controlled experimental conditions. Axenic monoalgal cultures have the advantage that bacterial processes do not interact with algal metabolism, and the response is not a mixture of the reactions of several different species. Culture media with known chemical composition allow an equilibrated supply of essential nutrients. It is also recommended to keep a constant optimal temperature around the algal cultures, and to set a photon flux density of photosynthetically active radiation which provides suficient, but not excessive light energy for an algal culture with a given cell density. For example, a photon flux density of 350 µM m-2 s-1, a temperature of 22°C, and a rotary agitation at 100 rpm (to homogenize the algal cell culture) ensure adequate developmental conditions to study the impact of heavy metal pollution on the selected algae and to evaluate the remediative potential of algae for heavy metal-contaminated water ponds. If several physiological parameters can be followed simultaneously, the results will give a better reflection of what really happens under natural conditions [55, 171]. Because algae are photoautotrophic organisms, their primary production of new organic compounds depends largely on the efficiency of using light energy under the given environmental conditions. This is why impact of heavy metals on their energetic metabolism determins their vitality, and leads to differential tolerance and sensitivity degrees of algal species and varieties against toxicity of various heavy metals occurring in polluted water. Parameters of induced chlorophyll fluorescence are very sensitive markers of photosynthetic efficiency, vitality and stress tolerance of algae inhabiting contaminated water. Chlorophyll fluorescence may be induced in dark-adapted algae (with stopped photosynthesis because of lack of light), and in algal cultures exposed to a constant background illumination. In dark-adapted algae the so-called conventional fluorescence parameters can be determined, while in constantly illuminated algal cultures pulse amplitude-modulated parameters of chlorophyll fluorescence may be registered and computed, upon application of modulating light flashes with determined period, wavelength and intensity. By combining parameters of conventional and modulated chlorophyll fluorescence, it is possible to appreciate potential (maximal) and effective quantum yield of photosynthesis, to determine the vitality index of algae under the given conditions (through the relative fluorescence decay), and to determine the protective capacity through dissipation of excess energy by heat emission, reflected by the nonphotochemical quenching (NPQ) of fluorescence signal. This parameter is based on the difference between the non-modulated maximal fluorescence (Fm) and the modulated maximal fluorescence (Fm’), and is obtained through the relation NPQ = (Fm – Fm’)/Fm’. Under normal growth conditions the numeric value of non-photochemical quenching is very small, but as stress conditions induce defensive reactions in the photosynthetic apparatus, its value increases significantly, reflecting that a certain degree of tolerance is developed against the disturbing environmental factor. This is why



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increase in NPQ is a valuable marker also of heavy metal stress and of algal capacity to cope with unfavorable conditions in the changed aquatic habitat [53, 82, 139, 204]. For example, when axenic monoalgal cultures of Scenedesmus acuminatus were exposed for 5 days under constant experimental conditions to different concentrations of cadmium, chromium and nickel, non-photochemical quenching of chlorophyll fluorescence was proportionally increased with the concentration of cadmium, but in case of chromium and nickel only concentrations as high as 50 µM and 500 µM, respectively, resulted in a significant raise of heat dissipation from light-absorbing chlorophylls, indicating a protective reaction of the photosynthetic apparatus against heavy metal toxicity (Figure 3). It is also observable that similar increases in NPQ values are induced by 50 µM chromium and 500 µM nickel, suggesting that nickel is less toxic to photochemical processes than chromium. Changes in NPQ depend both on type and concentration of heavy metal, so using this parameter a selective characterization of various heavy metal pollutions may be performed in situ and in vivo, without sacrificing the algae for determination.

Figure 3. Non-photochemical quenching (NPQ) of induced chlorophyll fluorescence in populations of the green microalga Scenedesmus acuminatus grown under controlled conditions and exposed for 5 days to different concentrations of cadmium, chromium and nickel (n = 5, vertical bars represent means ± SE, different letters indicate statistically significant differences at P < 0.05, according to postANOVA Tukey HSD test).

Dry algal weight is a basic structural indicator for evaluation of net biomass production in the presence of undesirable heavy metals. It enables us to distinguish between the toxicity levels of different heavy metals, to establish concentration tresholds for algal survival, and to select proper species and strains for remediation of water pollution [90]. For example, under the experimental conditions mentioned above for nonphotochemical fluorescence quenching evaluation, dry weight of Scenedesmus


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acuminatus populations exposed for ten days to different heavy metals registered distinctive changs (Figure 4). In case of cadmium, a concentration of 5 µM already caused a reduction in algal biomass, while the same concentration of nickel induced a moderate, but statistically significant stimulation of algal dry weight upon development of hardening reactions. Chromium started to decrease algal biomass only at 50 Mµ, and this inhibitory effect became more intense at 500 µM. Nickel started to moderately decrease algal dry weight only at the concentration of 500 µM. These results demonstrate that biomass production of algae is an easily determinable, but sensitive growth parameter to differentiate between impacts of different heavy metals and their different concentrations, and also to identify more tolerant or resistant algal varieties, suitable for phytoremediation of variously polluted aquatic habitats.

Figure 4. Dry biomass production of the green microalga Scenedesmus acuminatus exposed for ten days to different concentrations of cadmium, chromium and nickel in the aquatic medium (n = 5, vertical bars represent means ± SE, different letters indicate statistically significant differences at P < 0.05).

INTERACTIONS BETWEEN HEAVY METALS IN MICROALGAE EXPOSED TO CO-CONTAMINATED WATER Contaminated aquatic ecosystems rarely contain only one type of pollutant. Biosorption, transport, sequestration and biotransformation of heavy metals may be significantly influence by the presence of other heavy metal species, of different mineral nutrients, and even of organic xenobiotics (hydrocarbons, pesticides, surfactants). This is why antagonistic or synergistic interactions between essential and non-essential metals, and between heavy metals and non-metallic contaminants may occur, and these interactions have to be considered when the bioremediative efficiencies of algae are determined [27, 41, 102, 184, 185]. The most obvious competitive interactions exist



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among polluting heavy metals and metallic micronutrients during their membrane transport into cell compartments, while cooperative interactions among co-contaminating heavy metal species can be observed in concern with their toxic effects on various metabolic and developmental processes. Usually, non-essential heavy metals utilize the mechanisms evolved for the transport of essential mineral nutrients, so chemically similar essential and non-essential metals compete with each other over a shared transporter and for the chemical energy necessary for an active membrane transport which enables a bioconcentration of heavy metals against the electrochemical gradient. Cation diffusion facilitator and cation exchanger membrane proteins are such sites of uptake antagonism between different metal species. For example, cadmium enters plant cells through the transporter which normally performs the uptake of zinc, iron and manganese. This is a main reason why water pollution with micromolar amounts of cadmium causes iron, zinc and manganese deficiency symptoms. Similar competition for common transporters exists between nickel and copper, between chromium and iron. On the other hand, the lack of an essential metallic micronutrient may trigger the up-regulation of membrane transporters that will non-specifically uptake other metals, including toxic, non-essential heavy metals. Attenuation of each other’s accumulation in algal cells is also based on competition of co-contaminating heavy metals for a limitid number of binding sites due to the shared use of the same trnsporter. This is why the quantitative ratio between the competing metal species is more importand than their absolute amount, in the context of competitive or cooperative interactions. In some cases, heavy metals existing in the aqueous solution may also compete with anionic mineral nutrients for entering algal cells. For example, chromate utilize the transport system which primarily serves for sulfate uptake, while other anionic compounds containing heavy metals compete with phosphate for the same membrane transporter, which is less specific for one kind of inorganic anion [46, 62, 66, 113, 122, 172]. Solubility and uptake of heavy metals are considerably influenced by the presence in the aqueous solution or in cell compartments of organic chelators, which may be synthetic (e.g., ethylene-diamine-tetraacetate, ethylene-diamine-disuccinate) or natural compounds (e.g., citrate, malate, proline, histidine). Their existance facilitates bioextraction and sequestration of heavy metals, which is good for remediation purposes, but in the same time they enhance metal toxicity because they may increase intracellular concentration of harmful metal species. Through biotransformation of different organic pollutants, micro-organisms present in the polluted water also have a serious impact on the fate of different non-essential and essential metals concerning their availability for algae and their synergistic or competitive interactions [34, 79, 93, 97, 115, 116, 117, 132, 163]. Additivity or synergism may be oftenly observed when toxicity of heavy metals is investigated in co-contaminated aquatic habitats. In some cases, harmful effects of different heavy metals occur in different actions sites in an organism, and there is no direct interference between these pollutants. But in many cases, simultaneously present


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heavy metals act on different steps of the same physiological process, or development of tolerance and defense towards one heavy metal type confers similar protection or hardening to other existing heavy metals. This is a phenomenon known as cross-tolerance to different stress factors, and it is a common feature of plant reactions to disturbing external agents. For example, if one heavy metal induces in algal cells an increment in generation of reactive oxygen species, an enhanced activity of the antioxidative enzymes will confer protection to the same effect of other, co-existing heavy metals, so the simultaneous effect will not be the sum of effects registered upon separate exposure to the different heavy metal species [92, 109, 122, 179, 194, 198]. For studying antagonistic or synergistic relations between different heavy metals existing in the same polluted water, developmental and metabolic markers (structural and functional indicator) may be equally suitable, and it is desirable to simultaneoulsy investigate more parameters. For example, in experiments performed with axenic batch cultures of the green microalga Scenedesmus acuminatus under controlled, stable growth conditions, physiological markers which are relevant for evaluation of the remediative capacity, exhibited different patterns of variation when the alga was exposed for ten days, in Bold’s basic nutrient medium, to separate and combined pollution with the same concentration (50 µM) of cadmium, chromium and nickel. Concerning dynamics of the cell density of algal populations (which reflects the relation between the rate of reproduction through cell divisions and the rate of cell death), it was noticeable that nickel reduced the inhibitory effect of cadmium over time, but enhanced the toxic effect of chromium. Chromium could not compensate for cadmium effect on net reproductive rate (Figure 5). From the several parameters of induced chlorophyll fluorescence related to energetic efficiency of light use in photosynthesis, the non-photochemical quenching of fluorescence, which corelates with the protective capacity through dissipation of excess energy, is mostly increased by the simultaneous presence of chromium and cadmium, while nickel does not modify NPQ increment caused by cadmium, but it enhanced dissipative processes initiated by chromium (Figure 6). Changes in dry algal biomass upon exposure to co-contamination with associated heavy metals are also useful in establishing possible interactions between different heavy metals accumulated in algal cells. For example, nickel may reduce the inhibitory influence of cadmium, but in association with chromium it exerts a moderate, but statistically significant disturbing effect. Chromium also enhances toxicity of cadmium (Figure 7).



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Figure 5. Dynamics of cell density (determined cytometrically) in populations of the microalga Scenedesmus acuminatus exposed to separate and combined pollution of the aquatic medium with similar concentrations of nickel, chromium and cadmium (n = 5, vertical bars represent ±SE from means).

Figure 6. Non-photochemical quenching (NPQ) of induced chlorophyll fluorescence in populations of the green microalga Scenedesmus acuminatus grown under controlled conditions and exposed for 5 days to separate and combined pollution of the aquatic medium with similar concentrations of nickel, chromium and cadmium (n = 5, vertical bars represent means ± SE, different letters indicate statistically significant differences at P < 0.05).


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Figure 7. Dry biomass production of the green microalga Scenedesmus acuminatus exposed for ten days to separate and combined pollution of the aquatic medium with similar concentrations of nickel, chromium and cadmium (n = 5, vertical bars represent means ± SE, different letters indicate statistically significant differences at P < 0.05, according to post-ANOVA Tukey HSD test).

CONCLUSION Several microalgae are effectively able to contribute to remediation of heavy metalpolluted water, by performing bioextraction, concentration, stabilization, sequestration and even detoxification of certain heavy metals. Their use in phytoremediation of aquatic habitats relies on their ability to bind, to accumulate, to immobilize and to enzymatically convert heavy metals to less toxic forms. These abilities largely depend on prevailing growth conditions, on exposure time and metal concentrations, on resistance, tolerance or sensitivity of algal processes to heavy metal stress, on availability of inorganic nutrients, and on the presence of other pollutants. In many cases, immobilized algal cells embedded in gel particles perform a more sustained extraction, with the possibility of recovery of useful metals in a concentrated form. Selected biological markers for heavy metal tolerance are useful tools for identification of suitable algal species and strains, in an attempt to optimize phytoremediation of polluted water. Interactions between simultaneously present heavy metal species are inevitable in most of the real-world applications of bioremediation, this is why efficiency of microalgae in decontamination of water bodies has to be studied also in the presence of several heavy metals, if possible not only under laboratory conditions but also in largescale natural ponds, in order to reveal antagonistic relations or altered toxicity due to cross-tolerance. The NPQ parameter proves to be a reliable physiological marker not only for screening the protective capacity of algal photosynthetic apparatus against different heavy metals, but also for revealing relevant interactions among co-contaminating heavy



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metals. The easily determinable algal dry weight and reproductive rate, which are direct reflectants of productivity under stress conditions, as well as the quantitative ratios between the main types of photosynthetic pigments, are also good tools for evaluation of bioremediative applicability of different algae in variously heavy metal-polluted aquatic environments. Extension of tolerability experiments to large-scale water bodies polluted with heavy metals is a promising perspective for using selected microalgae in efficient and environmental-friendly technologies for bioremediation of anthropically affected aquatic habitats.

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Chapter 5

ADVANCES IN MICROBIAL DEGRADATION OF SUBSTITUTED PHENOLS WITH SPECIAL REFERENCE TO ACTINOMYCETES Namita Panigrahi1, Kannan Pakshirajan2 and Naresh K. Sahoo1,* 1

Department of Chemistry, Environmental Science and Technology Program, Institute of Technical Education and Research, Siksha‘O’Anusandhan (Deemed to be University), Bhubaneswar, India 2 Department of Biosciences and Bioengineering, Indian Institute of Technology, Guwahati, India

ABSTRACT The presence of substituted phenols in contaminated water and soil causes serious public health concern due to their detrimental effects on the human and many other living organisms. Therefore, removals of these toxic compounds have gained much attention in the last few decades. Biodegradation of phenolic compounds employing bacterial systems are well reported in the literatures, but biodegradation of substituted phenol is not fully addressed so far. This book chapter includes many problems associated with wastewater, contaminated with substituted phenols, such as i) contamination levels of substituted phenols in the environment and their toxicity on living systems, ii) biodegradation of substituted phenols by different microorganisms especially actinomycetes, with special emphasis on their metabolic pathway, biomass growth and degradation kinetics, iii) performance of various single or integrated bioreactor systems for substituted phenols biodegradation by different microorganisms.

* Corresponding Author Email: [email protected]; [email protected].


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Namita Panigrahi, Kannan Pakshirajan and Naresh K. Sahoo

Keywords: substituted phenols, biodegradation, actinomycetes, bioreactor systems, wastewater treatment

INTRODUCTION The rapid growth of industrial sectors discharges a huge amount of phenolics wastewater to the receiving environment. The major industries that discharge phenolic wastewater include petroleum refineries, coke oven plants, textiles, phenolics resin, petrochemicals, leather, insecticides, pesticides, pharmaceutical and wood preservatives, etc. The worldwide annual industrial production of chlorophenols (CP), bromophenols (BP) and nitrophenols (NP) is estimated to be 0.29 × 106 metric tonnes, 9500 tones and 20 million kilogram, respectively (Tiirola et al. 2002; IUCLID 2003). The concentration levels of these phenolic compounds in industrial wastewater are reported to be very high, for instance, 10,000 mg chlorophenol kg-1 soil in saw mill waste (Trapido et al. 2000); whereas, 26-3690 μg kg-1 of bromophenol found to be in Norwegian Sea water. Similarly, concentration of 4-NP in industrial wastewater is reported to be in the range of 10-17 × 103 mgl–1. Further, these substituted phenols are reported to be carcinogenic and teratogenic in nature. Due to the observed toxic effects of these substituted phenols on human, animal and microbial cells, these compounds are listed as priority pollutants by the U.S. Environmental Protection Agency (US EPA), which recommends the concentration of 4-CP and 4-NP to be lower than 1µgl-1 and 10 ngl-1, respectively in natural water (Erba et al. 2007). The US EPA has also set a limitation on the concentration of total toxic organic pollutants in the effluents of electrical and electronic industries at 1.37 mgl-1. The maximum permissible concentration of total organics (including nitrophenols) in the effluent of electroplating plant is restricted to 2.13 mgl-1 for plant discharge less than 10,000 gallons of wastewater per day, whereas, the maximum monthly average industrial effluent concentration of 4-NP fixed at 162 gl-1 (EPA 1988). Therefore, it is highly essential to eliminate these substituted phenolic compounds from contaminated wastewater before being discharged into the environment. Recently, many important technologies have been employed for the treatment of phenolic wastewater. Among them, adsorption, flocculation, precipitation and advanced oxidation processes such as photocatalysis, fenton reaction and sonication are widely studied by many researchers. However, generation of toxic by-products, high cost and low efficiency are some of the major limiting factors of these technologies. Therefore, the most economical and eco-friendly biodegradation technology seems to be an ultimate solution to the problem. But, the presence of chloro, bromo and nitro groups in phenolic compounds increases the resistance of the aromatic ring against biodegradation. Hence, only selective microorganisms such as Flavobacterium, Acaligenes, Trametesversicolor,


Advances in Microbial Degradation of Substituted Phenols …

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Pseudomonas, Rhodococcus and Arthrobacter species are widely employed for degradation of 4-CP, 4-BP and 4-NP, which uses these compounds as sole source of carbon and/or nitrogen for their growth. Among these microbial candidates, bacteria belong to the species actinomycetes are very efficient in degrading these substituted phenols by secreting both extracellular as well as intracellular enzymes. Therefore, actinomycetes are very efficient in degrading these substituted phenols. Among the different bioremediation techniques, combined or integrated anaerobic– aerobic wastewater treatment process has received a great attention owing to their numerous advantages; for instance i) consumption of low energy, ii) minimization of sludge volume production, iii) energy and resource recovery. Further, development of intrinsic tolerance of the anaerobic methanogens to dissolved oxygen (DO) facilitated coexistence of aerobic or micro-aerophilic microorganisms in the mixed culture system (Erguder and Demirer 2008). Therefore, mixed anaerobic–aerobic cultures systems have been widely investigated in the last two decades.

SUBSTITUTED PHENOLS Toxicity The toxicity of substituted phenols in living cells is due to its uncoupling effect on mitochondrial oxidative phosphorylation, which consequently increases the membrane permeability to proton and shifts the transmembrane pH gradients along with electrical potential (Escher et al. 1996). The uncoupling activity of these substituted phenols is fatal to microorganisms. In addition, formation of phenolic dimers further exaggerates the uncoupling activity. The extent of toxicity depends upon several factors such as pH, degree of substitutions on to the aromatic ring and their positions. The most toxic undissociated (molecular) forms of substituted phenols are found to be predominant in acidic environment, whereas, at higher pH, the less toxic dissociated form is predominant. The dissociation constant i.e., pKa of these phenolic compounds is strongly governed by the type of substituents. For instance, the pKa value of 4-CP is 9.3 and it is 7.18 for 4-NP; whereas, phenol has a pKa value of 10. Furthermore, lipophilicity of the halogenated compounds increases with an increase in the number of substituents and as a result of which the microbial uptake of the pollutants found to be enhanced, which further aggravates its toxic effect on microbial cells. In general, meta and para-substituted phenolic compounds are more toxic than substitution at the ortho position, which is ascribed to the shielding effect of the ortho-substituted chlorine to the OH group and consequently, it enhances the toxicity by interacting with active sites of living cells (Grimwood and Mascarenhas 1997). There are several literature reports on toxicity of substituted phenols on microbial cells affecting their enzyme systems, which may be due


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to their shared resonance electrons on their aromatic rings. Furthermore, substitution of halogens on the aromatic ring of the phenolic compounds enhances their recalcitrance properties due to their electron withdrawing effect as well as steric hindrance to the microbial enzymes sysrem (Copley 1997; Uberoi and Bhattacharya 1997). Hovander et al. (2005) reported that chronic exposure of fifty brominated and chlorinated phenols in plasma of Swedish blood donors causes damage of vital organs such as liver and bone marrow. Similarly, exposure to bromophenols causes multiple adverse effects on living organisms. Exposure of halogenated phenols causes serious endocrine-disrupting effects, particularly on thyroid hormone homeostasis due to their competition in binding to transthyretin (one of the human thyroid hormone binding transport protein) (Meerts et al. 1997). Pentachlorophenol (PCP) is a suspected carcinogen and teratogen. Grimwood and Mascarenhas (1997) reviewed the toxicity data of mono and dichlorophenols on algae, crustaceans and fish. They estimated the respective L(E) C50 (lethal concentration affecting 50% of the population, half maximum response) values ranging from 0.6-19.5, 2.55-29.7 and 5-7 mgl-1, respectively, which clearly reveals moderate to highly toxic nature of these phenolic compounds. 4-BP also uncouples the mitochondrial oxidative phosphorylation and aggravates liver damage by inducing oxidative stress and lipid peroxidation factors. Similarly, exposure to nitroaromatic compounds (NACs) particularly nitro-phenol and nitro-benzene are suspected to cause multiple disorder in human and animal models. Several literature reports are available on toxicity of these nitro-aromatic compounds on human health such as mutagenicity, urinary tract tumors and immuno-toxicity. Weihua et al. (2002) investigated EC-50 of activated sludge and reported a value of 10.70 mgl-1 for degradation of 4-NP. In another study, Razo-Flores et al. (1997) estimated the LC50 (concentration of an inhibitor where the response is reduced by half) values ranging from 4.91 to 9.96 mgl-1 for methanogens, which exhibited higher toxicity effects on the microbial cells.

BIOLOGICAL PROCESSES FOR SUBSTITUTED PHENOL DEGRADATION Biodegradation of Substituted Phenols by Aerobic Bacterial System Bioremediation of organic pollutants using bacterial systems is advantageous over other microbial systems owing to their ability to genetically evolve their degradation pathways to combat toxic xenobiotics present in contaminated environment. Bioremediation of phenolic compounds using pure culture of aerobic bacterial systems seems to be highly efficient. However, presence of aromatic ring and halogen



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substituent's on these phenolic compounds stabilizes the structure and thus offers resistance to enzymatic attack. Therefore, only selective microorganisms can metabolize substituted phenols as a sole source of carbon and energy. There are extensive evidence that, substituted phenols are completely degraded by bacterial systems which is well supported by stoichiometric release of inorganic halogen compounds (Yang et al. 2005), and equivalent biomass production with respect to the degradation of substituted phenols (Kumar et al. 2005) or the conversion of carbon 14 (14C) labeled chlorophenols to 14 CO2. The major approaches for degradation of substituted phenols by aerobic bacterial systems can be briefly described as follows. Initially the PCP undergoes hydroxylation by Pcp 4-monooxygenase enzyme (PcpB) of Sphingomonas chlorophenolica ATCC 39723 and forms tetra-chloro hydroquinone (TeCHQ) in presence of molecular oxygen, followed by dechlorination to 2,6-dichloro-1,4-hydroquinone (2,6-DCHQ) by a reductive dehalogenase of PcpC (Xun et al. 1992b; Tiirola et al. 2002; Thakur et al. 2002). Finally, in presence of oxygen 2,6-DCHQ is converted to 2-chloromaleylacetate by 2,6-DCHQ 1,2-dioxygenase (PcpA) with the liberation of chloride (Xun et al. 1999). Similarly, Matus et al. (2003) reported dechlorination of 2,4,6-TCP to 2,6-dichlorohydroquinone followed by 6-chlorohydroxyquinol by a monooxygenase enzyme system of Ralstoniaeutropha strain JMP134 (pJP4) prior to ring cleavage by a hydroxyquinol dioxygenase. Due to the special features of the monooxygenase enzyme, the second dechlorination reaction is catalyzed in absence of O2, where, dechlorination of 2,6-dichlorohydroquinone takes place by a hydrolytic enzyme (Xun and Webster 2004). Similar monooxygenases activities have been established in many other PCP-degrading aerobic bacterial species. Therefore, PCP-4-monooxygenase is known to be the major enzyme system responsible for degradation of chlorophenol by aerobic bacterial systems, which is also reported to be active towards degradation of various other halogenated phenols like 2,3,5,6 TeCP, 2,6 TCP, 2,3 DCP, penta-fluorophenol, tri-iodophenol, tribromophenol and di-bromophenol. PCP-4-monooxygenase enzyme also catalyzes para position of substituted phenols and liberated nitro group as nitrite, cyano group as cyanide, amino group as hydroxylamine and iodide as iodine (Xun et al. 1992b). Therefore, PCP-4 monooxygenase enzyme of aerobic bacteria plays a vital role in degradation of substituted phenols as well as several other toxic organic pollutants present in industrial wastewater. The PcpB gene of Pseudomonas species is shown to be homologous with PcpB gene of the actinomycetes species viz. Arthrobacter ATCC 33790 (Ferraroni et al. 2006). Thus, actinomycetes are potent high GC content aerobic bacteria for biodegradation of substituted phenols.


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Co-Metabolic Degradation of Substituted Phenol Co-metabolism is a simultaneous degradation process where, degradation of a second substrate depends on the availability of primary substrate such as sugar. For example, metabolism of primary substrate sugar, do not necessitate oxygenases enzyme which is normally employed by a bacterium in the degradation of 2-CP and PCP by cometabolism (Yang et al. 2005; Wang and Loh 1999). More particularly, induction of either dioxygenases or monooxygenases enzymes are found to be enhanced by presence of primary substrates and consequently facilitates the co-metabolic degradation of substituted phenols. Thus, the enhancement in removal of a co-pollutant might be attributed to the induction of enzymes and/or formation of biomass by the primary substrate required for the biodegradation process. Cobos-Vasconcelos et al. (2006) reported co-metabolic degradation of 2-CP by Burkholderia species using phenol as the primary substrate. In literature, various compounds analogous to phenol have been employed as the primary substrate by various bacterial species to support the cometabolic degradation of 2-CP, 4-CP and 2,4-DCP (Loh and Wu 2006; CobosVasconcelos et al. 2006; Kim et al. 2002). Similarly, 2,4,6-trichlorophenol was employed as the growth substrate for degradation of higher substituted chlorophenol by Azobacter species, Streptomyces rochei and Pseudomonas pickettii. In an another study, Goswami et al. (2002) reported the co-metabolic degradation of 2-CP, 4-CP and 2,4-DCP employing benzoate-induced Rhodococcus erythropolis M1 cells. The authors achieved complete degradation of 2-CP, 4-CP and 2,4-DCP up to a concentration of 300, 100 and 50 mgl-1, respectively. Goswami et al. (2005) investigated degradation of mixtures of 2chlorophenol, p-cresol and phenol employing a mixed culture system of R. erythropolis and Pseudomonas fluorescens in an activated sludge process. The presence of p-cresol prevented growth of these microorganisms in the mixed substrate system, whereas the pollutants degradation rates found to be superior than those using the pure cultures (Goswami et al. 2005). Cassidy et al. (1999) reported degradation of nitrophenols by pentachlorophenol induced hydroquinone dioxygenase enzyme of Sphingomonas species. Based on the structural analogue, other phenolic compounds have been frequently used as growth substrate for degradation of mononitrophenols. Similarly, co-metabolic transformations of nitroaromatic compounds using benzoic acids, peptone, yeast extract and urea have been reported (Qiu et al. 2007; Danuta et al. 2011). These results suggest the feasibility of chemicals such as benzoic acid, phenol, yeast extract, peptone and even other chlorophenols as co-substrates for co-metabolic degradation of substituted phenols present in contaminated wastewater.



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BIODEGRADATION OF SUBSTITUTED PHENOL BY ACTINOMYCETES Advantages of Actinomycetes over other Microorganisms for Degradation of Substituted Phenol The most commonly reported substituted phenols degrading actinomycetes (high GC content aerobic bacteria) are belong to Nocardia, Acaligenes, Rhodococcus and Arthrobacter species. These substituted phenols degrading actinomycetes are aerobic in nature and therefore, these species grow and degrade pollutants faster than anaerobic microorganisms. Further, complete mineralization of these substituted phenolic compounds can be achieved by actinomycetes species within a shorter time period rather than simple transformation as compared with anaerobic bacterial system (Kim et al. 2002). The salient feature of these actinomycetes species are due to their ability to synthesize a wide variety of enzymes mainly monooxygenase, dehalogenase, catechol 1,2dioxygenase, 1,2 hydroquinol dioxygenase and membrane bound enzyme e.g., cytochrome P450. which are responsible for the substituted phenols biodegradation. In general, actinomycetes produce both membrane bound and cytoplasmic enzyme systems that enhances the rate of phenolics biodegradation as compared with the low GC content aerobic and anaerobic bacterial systems which usually do not produce membrane bound enzymes. Uotila et al. (1991) reported metabolic activities of cytochrome P 450 enzyme of actinomycetes species like Mycobacterium and Rhodococcus species for PCP dechlorination and hydroxylation whereas, Sphingomonas species produces only PcpB monooxygenase enzymes for aerobic degradation of phenolic compounds. Although growth of these actinomycetes is found to be slow, yet single inoculum is sufficient enough to mineralize substituted phenols. Furthermore, in actinomycetes, the numbers of residues present in the active cleft of the enzymes are responsible for substrate selection and interaction that are altered mainly at leu 80, Asp 83, Val 107 amino acids positions. This reveals a broad specificity nature of the actinomycetes enzyme system, which enables the enzymes to interact and degrade a wide range of different toxic substrates present in real industrial wastewater (Ferraroni et al. 2006). There are several literature reports on single actinomycetes species capable of degrading different types of substituted phenols present in wastewater. For instance, Rhodococcus erythropolis CCM 2595 (ATCC 11048) is shown to degrade phenol, p-nitrophenol, p-chlorophenol, pyrocatechol, resorcinol, hydroquinone and hydroxybenzoate or as respective dual mixtures with phenol (Čejková et al. 2005). Rhodococcus species are also reported to degrade poly aromatic hydrocarbons (PAHs), phenol and a mixture of o-, m- and pcresols efficiently. Due to the broad substrates specificities nature of R. erythropolis CCM 2595 enzyme systems, the actinomycetes species can interact and degraded many


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para-substituted phenolic compounds such as cresol, chlorophenol, nitrophenol and resorcinol (Fialová et al. 2003). However, growth of R. erythropolis CCM 2595 do not supported by para-substituted phenolic compounds. Several Rhodococcus species, particularly, Rhodococcus chlorophenolicus PCP-1 has been shown to degrade various chlorinated phenols, chlorinated syringols, chlorinated guaiacols and o-methylate chlorinated p-hydroquinones efficiently. Arthrobacter chlorophenolicus A6 can degrade various substituted phenols such as 4-NP, 4-BP, 4-idophenol, 4-fluorophenol via the hydroquinone pathway (Wasteberg et al. 2000). Similarly, Navrátilová et al. (2005) reported the degradation of 3- nitrophenol, 4-nitroguaiacol and 4-nitrocatechol employing R. wratislaviensis J3. It is observed that some actinomycetes species are capable to degrade chlorophenol both under aerobic and anaerobic conditions. For example, in the presence of iodosobenzene the rate of anaerobic PCP degradation by R. chlorophenolicus PCP-1 found to be similar to that of aerobic PCP degradation. The growth rate of aerobic bacteria such as Pseudomonas species are reported to be decreased from 0.09 h-1 to 0.05 h-1 when the PCP concentration found to be increased from 40 mgl-1 to 150 mgl-1 (Radehaus and Schmidt 1992). On the other hand, actinomycetes develops hyphae and forms large micro colonies to protect the inner cells and hence shows very high toxicity tolerance even at a very high initial concentration of the toxic pollutants. Thus, actinomycetes are believed to be an important tool for biodegradation of many toxic organic pollutants. However, performance of these actinomycetes species for biodegradation of substituted phenols is not yet reported extensively in the literature. Golovleva et al. (1992) reported that Streptomyces rochei 303, immobilized in polycaporamide fibers, could degrade a combined mixture of 2,4,6-TCP, 2,4-DCP and 2,6-DCP up to a concentration of 205 mgl-1, 143 mgl-1 and 61 mgl-1, respectively, as the sole source of carbon and energy. Arthrobacter species ATCC 33790 mineralized PCP at high influent concentrations of 525 mgl-1 and 349 mgl-1 when operated in a chemostat and a fixed film bioreactor system, respectively (Edgehill 1994). Goswami et al. (2002) reported that benzoate induced R. erythropolis M1 completely mineralized 2-CP, 4-CP and 2,4 -DCP at an influent concentration of 300, 100, 50 mgl-1, respectively. Bae et al. (1996) established a new pathway for degradation of 4-CP by Arthrobacter ureafaciences CPR706 where, the chloro substituent eliminated in the first step and followed by production of hydroquinone as a transient intermediate. Strain CPR 706 demonstrated superior chlorophenol degradation rate and substrate tolerance capacity compared to other 4-CP degrading bacterial strains following the same hydroxylation pathway at the second carbon position to generate chlorocatechol. These aspects provide a strong support on the potential of actinomycetes for biodegradation of substituted phenols.



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Metabolic Pathway for Degradation of Chlorophenol by Actinomycetes In general, biodegradation of 4-CP by actinomycetes and other microorganisms is a multistep process. Initially, 4-CP is oxidized to 4-chlorocatechol and subsequently ortho cleavage of the aromatic ring take place as shown in Figure 1 (a) (Farrell and Quilty 2002). In few microorganisms, biodegradation of 4-CP is occurred by meta-cleavage pathways as shown in Figure 1 (a) (Farrell and Quilty 1999). Chlorocatechol (or catechol) 1,2-dioxygenases enzyme of actinomycetes species converted catechol and its derivatives to cis, cis-muconate and its derivatives, consequently it gets transformed to muconolactone catalyzed by cis,cis-muconate cycloisomerase enzyme system. Further, in presence of muconolactone isomerase enzyme the muconolactone is converted to 3-oxoadipate enol-lactone. Finally, 3-oxoadipate enol-lactone hydrolase (OELH) enzyme catalyses the 3-oxoadipate enol-lactone to 3-oxoadipate. The catechol (or chlorocatechol) 1,2-dioxygenases enzymes are categorized under intradiol 1,2-dioxygenases system, which is associated with the cleavage of carbon-carbon bond between catechol and hydroxyl groups by inserting an oxygen molecule (Broderick 1999). Following the ring cleavage step, the chlorine atom is removed and the carbon skeleton simplified and assimilated into the TCA cycle of the cell (central metabolism) (Figure 1a). However, few microorganisms do not exhibit degradation of 4-CP via cleavage of 4- chlorocatechol pathway. Arthrobacter ureafaciens CPR706 (Bae et al. 1996a) was found to convert 4-CP to hydroquinone (1,4-dihydroxybenzene); however, the pathway is not yet clearly established. Similarly, production of hydroquinone from 4-CP by Nocardioides sp. NSP41 was reported, but the complete pathway is not fully established. Nordin et al. (2005) reported biodegradation of 4-chlorophenol by A. chlorophenolicus A6 that generates 4-chlorocatechol before ring cleavage. The 4-chlorocatechol was further converted into hydroxyquinol and consequently cleaved to produce maleylacetate (Figure 1b). Nordin et al. (2005) also established a second pathway in which they observed conversion of 4-chlorophenol into hydroquinone followed by hydroxylation to produce hydroxyquinol (Figure 1b). In general, degradation of highly chlorinated compound yielded ring-cleavage intermediates such as hydroxyquinol and its chlorinated derivatives, however, in A. chlorophenolicus A6 degradation of 4-CP is occured via hydroxyquinol before cleavage of the phenolic ring. Generation of ring cleavage substrate (hydroxyquinol) is also reported for degradation of other phenolic compounds such as 4-aminophenol and 4-NP (Takenakaet al. 2003). Degradation pathway of 4-chlorophenol by A. chlorophenolicus A6 involves several enzymatic steps as shown in Figure 1 (b). In the first step, monooxygenase enzyme catalyzes the initial transformations of 4-CP. In the second step, an enzyme system is used for reductive dechlorination of 5-chlorohydroxyquinol. In the third step, hydroxyquinol dioxygenase enzyme is involved in the transformation of intermediate product. Lastly maleylacetate reductase is used for channeling the ring


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cleavage product into the central metabolism of the cell (Nordin et al. 2005). The authors further reported genes and open reading frames i.e., Cphgene cluster encoding these types of enzymes in A. chlorophenolicus A6. The gene cluster was reported to be harbor two putative monooxygenases, two hydroxyquinol 1,2-dioxygenases, two putative maleylacetate reductases along with two putative regulatory genes.

Figure 1. 4-chlorophenol degradation pathways. (a) Conventional pathway (The University of Minnesota Biocatalysis/Biodegradation Database, (Ellis and Wackett 1997). (b) Proposed 4-CP degradation pathway in A. chlorophenolicus A6 (Nordin et al. 2005). The transformation from 5chlorohydroxy quinone to 2-hydroxy 1,4 benzoquinone is theoretically due to reductive dechlorination. The presence of maleylacetate is inferred from genetic and biochemical evidence. Compounds shown in parenthesis indicates hypothetical intermediates supported, but not yet confirmed.



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Nordin et al. (2005) developed a mutant strain of A. chlorophenolicus A6 called T99, where one of the hydroxyquinol 1,2- dioxygenase genes is disrupted by the insertion of a transposon. The authors observed very poor growth of the mutant strain in 4-CP which strongly revealed a vital role of hydroxyquinol 1,2-dioxygenase gene (CphA-I) for effective growth of the microorganism in utilizing chlorophenol as substrate. The same enzyme system also used for degradation of 4-CP, 4-BP and 4-NP. However, the growth of the mutant T99 is improved in presence of unsubstituted phenol, which strongly revealing presence of an additional enzyme system engaged for metabolism of unsubstituted phenol. Thus, degradation process of 4-chlorophenol to hydroxyquinol by A. chlorophenolicus A6 exhibited in multiple pathway and which is further confirmed by the presence of two monooxygenases enzyme systems encoded by the Cph gene cluster (Nordin et al. 2005). They also reported that, one of the monooxygenase enzyme might have executed ortho-hydroxylations in both the branches, whereas, the other enzyme perform para-hydroxylations. Similar existences of two branches of metabolic pathways are also reported in literature for the transformation of toluene by Rhodococcus OFS and Burkholderiacepacia JS150. Häggblom et al. (1989) investigated the potential of Rhodococcus sp. CG-1 and Rhodococcus sp. CP-2 for degradation pentachlorophenol. They reported that degradation of PCP to tetrachlorohydroquinone occurred through para hydroxylation pathway followed by completely dechlorinated to trihydroxybenzene via three successive reductive dechlorinations and a hydroxylation steps. Bondar et al. (1999) investigated the ortho cleavage oxidative dechlorination of 2,3,5 trichlorophenol to 3,5-dichlorocatechol catalyzed by phenol hydroxylase enzyme of R. opacus 1G, but usually this phenol hydroxylase enzyme favors mono-oxygenation of the non-halogenated compounds. Similarly, R. percolatus strain is reported to be degraded mono- di- and trichlorophenols whereas, R. phenolicusa strain degraded chlorobenzene and dichlorobenzene as sole source of carbon and energy. Vogt et al. (2004) investigated the potential of several Rhodococcus strains in degradation of chlorophenol and chlorobenzene from contaminated ground water. The metabolic pathway of these strains are shown as follows: First chlorophenol and chlorobenzene under goes aerobic hydroxylation and produces chlorocatechols as the major intermediate product followed by completely degradation via modified ortho-cleavage pathway. König et al. (2004) investigated metabolic pathway of substituted phenol biodegradation by R. opacus1CP and reported that the degradation of 3,5-dichlorocatechol and 4-chlorocatechol are found to be regulated by clcBRAD operon, whereas, the degradation of 3-chloro-catechol regulated by the clcA2D2B2F operon. While, metabolism of protocatechuate and catechol are encoded by the pca and cat operons. They also reported that, degradation of chlorocatechol is encoded by the clc2 and clc operons located on the large linear plasmid of p1CP (740 kb). In the contrary, the mechanism of gene expression for enzyme associated with biodegradation of 2,4,6-TNP by R. opacus Hl PM-1 found to be a


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negative regulated. For example; IclR-type repressor designated as NpdR, appeared to be negatively regulated (Nga et al. 2004). Further, the 3-chlorocatechol biodegradation pathway of R. opacus 1CP as discussed above found to be differs from other reported literature. For example; the dechlorinating process of 3-chlorocatechol is found to be catalyzed by muconolactone isomerase enzyme, whereas, it is catalyzed by chloromuconate cycloisomerase enzyme in other pathways (Moiseeva et al. 2002). In fact, the two chlorocatechol degradation pathways of R. opacus 1CP found to be derived through independent but convergent evolution. Hydrolytic-dechlorination and oxidation of 2,4,6-trichlorophenol to 6-chloro-2-hydroxyquinone by Cupriavidusnecator JMP134 found to be catalyzed by phenol hydroxylases enzyme system (Xun and Webster 2004). In general, the phenol hydroxylases enzyme systems, the free FAD found to be reduced by the flavin reductase subunit and the reduced FAD utilized by the NADH dependent monooxygenase subunit during the hydroxylation process of phenolic compounds to catechol (van der Heuvel et al. 2004). Szokol et al. (2014) investigated the transcriptional regulation of phenol hydroxylase enzyme in Rhodococcus jostii RHA1. They reported differences in the transcriptional regulation of two clusters of the gene encoding paralogous components of the enzyme. In Rhodococci species the two-components of phenolhydroxylase gene expression is positively regulated by an AraC/XylS-type regulator systems (Szokol et al. 2014; Takeo et al. 2008). Similarly, transcription of phenolhydroxylase genes is activated by an over expressed AraC/XylS-type regulator in R. erythropolis CCM2595 and R. jostii RHA1 strains, which clearly indicated the role of phenol and other aromatic compounds as an inducer of FAD-dependent phenol hydroxylase enzyme system in the biodegradation process (Szokol et al. 2014; Fialova et al. 2003; Cejkova et al. 2005). In another study, Kolomytseva et al. (2007) reported the potential of R. opacus 1CP in biodegradation of p-cresol through the ortho pathway via 4methylcatechol. The R. opacus 1CP also produces catechol-1, 2-dioxygenase enzyme, together with dioxygenases activity for degradation of 4-chlorophenol and p-toluate (Kolomytseva et al. 2007). Similarly, acclimatized culture of R. opacus 1CP also reported to be degraded 4-CP as well as 2,4-dichlorophenol and produced several intermediate product such as 4-chlorocatechol, 3,5-dichlorocatechol, respectively (Eulberg et al.1998a).

Catabolism of Nitrophenols by Actinomycetes Several bacterial species are reported to be degraded nitro-aromatic compounds and use them as sole source of carbon and nitrogen. The biodegradation of nitrophenols is initiated by oxidative or reductive reactions. Generally, monooxygenase enzyme involved in removal of the nitro group from nitrophenols during the biodegradation process. Kitagawa et al. (2004) investigated the metabolic pathway of R. opacus SAO101 in



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degradation of 4-NP along with biphenyl and naphthalene. They reported conversion of 4-NP to 4-nitrocatechol, hydroxyquinol and ultimately to maleylacetate by hydroxyquinol pathway. Similar nitro phenol biodegradation pathway is also observed in most of the gram negative bacteria particularly in Burkholderia species. Similarly, R. wratislaviensis J3 shown to degraded 4-nitrocatechol, 3- nitrophenol and 4-nitroguaiacol (Navratilova et al. 2005). Sahoo et al. (2011) investigated the potential of Arthrobactor chlorophenolicus A6 in degradation of 4-nitrophenol employing a novel upflow packed bed reactor system. They achieved almost completely removal of 4-NP even at a very high influent concentration of 1000 mgl-1 within smaller HRT value of 7.5 hr. However, transient accumulation of nitrocatechol as an intermediate product found to be observed with increase in concentration of 4-NP or with decrease in HRT of the reactor (Sahoo et al. 2011). Rhodococcus and Nocardioides species degraded substituted phenol such as TNP and 2,4-DNP mediated by a hydride transferases enzyme system as shown in Figure 2.

Figure 2. Metabolic pathways of trinitrophenol biodegradation by Rhodococcus and Nocardioides strain.


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The hydride transferases enzyme involved in transfer of the hydride ions from NADPH to coenzyme F420 and generated its reduced form of F420H2. The reduced form of F420H2 is employed as a substrate for hydride transferases enzyme to transfer the hydride ions further to TNP or DNP, consequently generated hydride and dihydride Meisenheimer complexes of TNP or DNP as shown in Figure 2 (Ebert et al. 2001). Two successive hydrogenation reactions of TNP degradation resulted the formation of acinitro form of the dihydride Meisenheimer complex of TNP (2He-TNP). A nitrite group is removed from 2He-TNP, forming the hydride Meisenheimer complex of DNP, which further hydrogenated to produced 2, 4-dinitrocyclohexanone (DNCH). Consequently the DNCH is found to be converted to 4,6-dinitrohexanoate, whose two nitro groups are finally removed. It is also reported that P-nitrophenol monooxygenase enzyme involved in degradation of nitrophenol found to be clustered with the genes coding for hydroxyquinol 1,2dioxygenase, maleylacetate reductase and along with a regulatory genes of several Actinobacteria such as Rhodococcus species PN1(Yamamoto et al. 2011), R. opacus SAO101 (Kitagawa et al. 2004) and Arthrobacter species JS443 (Perry and Zylstra 2007). The expression of 4-nitrophenol hydroxylase enzyme gene of Rhodococcus species PN1 is found to be regulated by NphR activator where 4-nitrophenol played as role of an inducer (Takeo et al. 2008). Several authors reported the crystal structures of enzyme systems associated with degradation of substituted phenols for instance; catechol (3 or 4 chlorocatechol) 1,2-dioxygenases enzymes of the genus Acinetobacter, R. opacus 1CP and Rhodococcus opacus 1CP (Vetting and Ohlendorf 2000; Ferraroni et al. 2006; 2013; Matera et al. 2010).

Catabolism of Fluorophenol and Bromophenol by Actinomycetes Finkelstein et al. (2000) established the metabolic pathway of fluorophenols degradation by Rhodococci species. They reported biodegradation of fluorophenols occurred via ortho hydroxylation pathway and converted to fluoro-catechols. Whereas, biodegradation of mono-fluorophenols by Rhodococcus opacus 1CP occurres by another pathway, which generate trihydroxyfluorobenzenes as an intermediate product. Bondar et al. (1999) reported biodegradation of various di and tri-fluorophenols by Rhodococcus corallinus135. Though biodegradation of substituted phenol studied extensively in literature however, only few microorganisms are available for biodegradation of bromophenol, for example R. opacus GM-14 and Achromobacter piechaudii TBPZ (Ronen et al. 2005). Sahoo et al. (2013) investigated the potential of A. chlorophenolicus A6 for degradation of 4-bromophenol employing a newly designed packed bed reactor



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system. They achieved almost completely removal of 4-BP and more than of 98% toxicity even at a very high influent concentration of 1000 mgl-1 within smaller HRT value of 12.5 hr. Similarly, Rhodococcus species MS11 reported to be degraded dibromobenzenes as sole source of carbon and energy (Rapp and Gabriel-Jurgens 2003).

KINETICS OF MICROBIAL GROWTH AND SUBSTITUTED PHENOL DEGRADATION For the growth of any microorganism appreciably, the amount of substrate concentration should be adequately high relative to the number of microbial cells to permit many doublings of the initial population and vice versa. Therefore, it is apparent that the extent of microbial growth depends on the initial substrate concentration. In order to explain the microbial growth and substrate degradation kinetics, several growth kinetic models such as logistic, logarithmic, and Monod have been reported. Similarly, several non-growth associated models like zero order, first order, Monod based and three-half order models have been proposed in the literature. To determine the effect of substrate concentration on microbial growth, specific growth rate of the cultures at varying substrate concentrations can be computed as follow:



1 dx X dt

(1)

Where, the specific growth rate is denoted as µ(h-1), X denoted the biomass concentration (mgl-1). Inhibitory effect on the microbial growth can be modeled employing appropriate substrate inhibition models as reported in literature (Kumar et al. 2005; Nuhoglu and Yalcin 2005). In general, the microbial growth can be represented by a simple Monod equation as follows.



max S KS  S

(2)

Where, the maximum specific growth rate of the microorganism is denoted by μmax (h-1), S denoted the concentration of the limiting substrate (mgl-1), half saturation constant is denoted by Ks (mgl-1). However, Eqn. (2) is not capable of explaining inhibition of microbial growth at higher substrate concentrations. In such cases, Edward, Andrews, Web and Haldane models are usually employed to study the microbial growth kinetics. The Haldane model (Kumar et al. 2005) can be expressed as follow:



128



max S

(3)

S2 KS  S  Ki

Where, Ki is the inhibition coefficient (mgl-1). Yano and Koga (1969) suggested a model based on the dynamic behavior of microbial growth in a continuous bioreactor system. In their study, growth inhibition occurred when the rate limiting substrate is found to be high. The model form is given in Eqn. 4:



max

(4)

n

(KS / S )  1   (S / K j )

j

j 1

Where, Kj is a positive constant. Similarly, Mulchandani and Luong (1989) proposed a kinetic model as shown in Eqn. 5, which is the modified form of Haldane model.



max S S  K S  ( S 2 / Ki )(1  S / K )

(5)

Where, the substrate inhibition constant represented by Ki (mgl-1) and K denote for a positive constant. Luong (1987) proposed a substrate inhibition growth kinetic model as represented in Eqn. (6) which deals with substrate stimulation both at low and high substrate concentration. The, maximum substrate concentration Sm can be estimated using the model Eqn (6) as follow (Luong 1987).

max S 

S   1   K S  S  Sm 

n

(6)

Among these substituted phenols, degradation of chlorophenol under aerobic and anaerobic environments has been studied extensively. While, only a few reports are available on kinetics of nitrophenol biodegradation, whereas, biodegradation of bromophenol and fluorophenol found to be very scanty. The growth rate of aerobic bacteria is found to be high when chlorophenol used as a sole source of carbon and energy. A biomass growth rate of more than 2 day-1 was estimated using actinomycetes and Flavobacterium species for the degradation of 2,4-DCP, 2,4,6-TCP and PCP as sole



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source of carbon and energy (Golovleva et al. 1992; Hu et al. 1994). Usually the biomass yield of aerobic bacteria on chlorinated phenols found to be reduced with increase in number of chlorine substitutions, as chlorine is not incorporated into the biomass and partly due to the poor carbon conversion efficiencies. For example, the biomass yield of aerobic bacteria grown in TCP and DCP reported to be 0.132 g and 0.421 g dwt biomass g-1 chlorophenol consumed (Rutgers et al. 1997), whereas, the values ranged from 0.054 0.190 g dwt biomass g-1 chlorophenol consumed for aerobic bacteria grown on PCP (Hu et al. 1994; Rutgers et al. 1997). In general, the specific activities of aerobic bacteria grown on chlorophenols are estimated to be in the ranges of 29 and 851 mg chlorophenol bio-transformed g-1 dwt biomass day-1 (Kargi and Eker 2004; Sahinkaya and Dilek 2005). On the contrary, an anaerobic microorganism that involves reductive dechlorination of chlorophenols yielded a lower specific activity in the range of 0.37 and 12.9 mg chlorophenol bio-transformed g-1 dwt biomass day-1(Mohn et al. 1999; Ye et al. 2004). Therefore, it is clearly indicated that, the specific activities of aerobic chlorophenols degrading microorganisms found to be higher than that of anaerobic processes. In general, the lower the half saturation velocity coefficients higher the affinity of the substrate to the microbial cells (Tsai and Juang 2006). The Ks or Km values for most of the microbial biodegradation processes are reported to be in the range of 0.01 -11.7 mgl-1 and 13-35 mgl-1 for degradation of chlorophenol and nitrophenol, respectively (Kumaran and Paruchuri 1997; Bhatti et al. 2002; Kumar et al. 2005). On the contrary, Kargi and Eker (2004) reported a higher Ks value of 112 mgl-1 for co-metabolism degradation of 2-4 DCP by an aerobic bacterial system. Biodegradation of low concentration chlorinated phenol was investigated in a previously enriched microorganism in aerobic (Melin et al. 1997) and anaerobic bioreactor systems (Magar et al. 1999). They estimated half-velocity coefficients as low as 0.014–0.016 mgl-1 under both aerobic and anaerobic conditions which strongly revealing the affinity of the substrate for the biomass yield. The maximum specific growth rate (µmax) of aerobic bacteria utilizing chlorophenol, nitrophenol, cresol and phenol as growth substrates mostly in the range of 0.15 and 0.55 h-1 (Kumaran and Paruchuri 1997; Sahinkaya and Dilek 2005), which indicates that the microbial systems efficiently utilizes the substrates for their growth. The value of the inhibition constant (Ki) signify the degree of toxicity tolerance of the microorganism to the substrate. In general, larger the Ki value higher the microbial toxicity tolerance to the substrates (pollutants). The values of inhibition constant (Ki) mostly in the range of 145 to 516 mgl-1 for 4-CP, 4-NP, phenol and cresol biodegradation by mixed culture and pure culture of Pseudomonas species, C. tropicalis and Acinetobacter calcoaceticus (Kumaran and Paruchuri 1997; Sahinkaya and Dilek 2005) indicating a higher toxic tolerance of these microorganisms towards the different substrates.


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BIOREACTOR SYSTEMS FOR TREATING WASTEWATER CONTAINING SUBSTITUTED PHENOLS Most of the biological reactor systems can be classified into two main groups viz. suspended biomass growth process and immobilized biomass system. Stirred tank, air-lift and bubble column reactors can be classified under the former group. Whereas, immobilized bioreactors can be classified further into fixed bed (film) reactor and particle based reactor. For instance; among the fixed bed (film) reactors packed-bed bioreactor, membrane bioreactor and rotating biological contactor are very popular in treatment of industrial wastewater. Whereas, in case of particle-based reactors like air lift bioreactor, fluidized bed bioreactor are found to be trendy, where the microorganisms are immobilized onto a solid inert biomass supporting moving carriers. The pollutants and oxygen mass transfer coefficient in particle based bioreactor systems are found to be superior than fixed film reactor system. On the contrary, the biomass profile in the fixed film reactor is found to be better than particle based bioreactor system. There are several reports on treatment of wastewater contaminated with substituted phenols using different bioreactor systems and are discussed further. Quan et al. (2004) investigated the biodegradation of mixture of 2,4-DCP and phenol by Achromobacter species immobilized on to honeycomb-like ceramic as carrier employing an internal loop airlift bioreactor system. They reported decrease in removal efficiency of 2,4-DCP from 100 to 87.9% with increase in phenol loading rates, whereas, phenol removal efficiency remained at about 99.6%, Which clearly indicated that the presence of phenol inhibited biodegradation of 2,4–DCP consequently the sole source of carbon and energy found to be shifted from 2,4-DCP to phenol. Jajuee et al. (2007) studied the biodegradation of p-xylene and naphthalene by an acclimatized microbial culture in a pilot-scale novel airlift immobilized bioreactor system. They achieved completely biodegradation of pxylene and naphthalene at an initial loading rate of 81 mgl-1h-1 and 40 mgl-1h-1 respectively. In most of the studies Monod kinetic model is found to be well described the experimental findings obtained during the course of batch and continuous operations. Table 1 presents the performance of various bioreactor systems reported in the literature in treatment of substituted phenols contaminated wastewater operated under continuous mode.


Table 1. Bioreactor systems employed for treatment of wastewater containing substituted phenols investigated under continuous mode of operation Bioreactor system

Packed bed biofilm reactor

Microorganism used

Pollutant treated

Influent conc. (mg l-1)

Effluent conc. (mg l-1) 1.39

Removal efficiency (%) 99% TCP COD (≈93%)

Burkholderia kururiensis and Stenotrophomonas sp -

2,4,6-TCP

139 VLR= 46.83 mgl-1day-1

4-NP

5.28

Two-phase SBR

Mixed culture

4-NP

528.73 VLR=1490 mgl1 day-1 450

SBR

TM7, Proteobacteria

2 stage RBC

Pseudomonas stutzeri

(HAIB)and Fixed film reactor PUF Batch and SBR Reactor (EFBAB)GAC fluidized bed Membrane bioreactor acrylic sheet Hollow fiber polyethylene

Methanosarcina and Methanosaeta Pseudomonas stutzeri

4-NP, 2,4-DNP, aniline 4-CP 2,4DCP peptone PCP 4-CP, 2,4DCP 4-CP and Phenol PCP

Aerobic stirred tank bioreactor

Pseudomonas putida ATCC 49451 Mixed culture

HRT(h)

Reference

-

Gomez-De Jesus et al. 2009

99

-

Rezouga et al. 2009

BDL

100

-

50 - 180

1-2

99

12

Tomei et al. 2008 Xing-yu et al. 2007

220 and110

>98

16.8

Sahinkaya et al. 2006b

21

BDL 2nd stage BDL

100

-

Saiaet al. 2007

220 and 110

0.05

100

24

200 4-CP and 600 phenol LR of 12–40 mg m-3d-1

BDL

-

10–15

BDL

99.9

12

Sahinkaya et al. 2006a Loh and Ranganath 2005 Visvanathanet al. 2005


Table 1. (Continued) Bioreactor system

Microorganism used

PBR

Mixed culture

SBR

Mixed culture

Honeycomb-like ceramic column packed ALR. PBR GAC- FBR

Pollutant treated

Influent conc. (mg l-1)

Effluent conc. (mg l-1) BDL

Removal efficiency (%) 99.7

HRT(h)

8.7

Zilouei et al. 2006

BDL

100

16

Tomei et al. 2004 Xiangchun et al. 2003

2-CP, 4CP, 2,4 DCP 2,4,6- TCP 4-NP

100 LR of 264 mgl−1day−1 300-400

Achromobacter sp.

2,4-DCP and phenol

49.96

-

94–99

36

Mixed culture Mixed culture

4-CP 4-CP

20 20-50

BDL 0-14

100 69-100

2.8 0.28

UASB

Mixed culture

2-30

-

90-99

12-30

Batch-fed Air-lift percolator

Mixed culture

2NP, 4-NP, and 2,4 DNP 2,4,6-TCP, 2,3,4,6-Te CP and PCP PCP

16.8 TCP, 28.8 TeCP,9.1 PCP

0.04

99.9

64

Reference

Kim et al. 2002 Carvalho et al. 2001 Karim and Gupta 2001 Langwaldt et al.1998

FBR Celite ROligotrophic enriched 0.727 0.006 99.79 4.45 Melin et al. 1997 633diatomaceous culture earth #1 Reactor Type: UASB, upflow anaerobic sludge bed; HAIB Horizontal flow anaerobic immobilized biomass reactor; EFBAB, External-loop fluidized bed airlift bioreactor; PBR, packed bed reactor; FBR, Fluidized bed reactor; SBR, Sequencing batch reactor; RBC, Rotating biological contractor; ALR, air-lift bioreactor.



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ANAEROBIC BIODEGRADATION Chlorinated phenols are readily under goes anaerobic biodegradation by reductive dechlorination process, where, chloro-groups of the phenolic compounds are displaced by hydrogen atoms. Reductive dechlorination usually occurs in presence of electron donors for example, recovery of ferrous iron (Fe2+) from supplemented ferric iron (Fe3+) revealing a strong evidence of chlorophenol biodegradation linked to iron reduction. Similarly, Bae et al. (2002) demonstrated degradation of 2-chlorophenol in an activated sludge process. They observed degradation of 2-CP increases with presence of nitrate and decreases in absence of nitrate. Genthner et al. (1989) investigated the degradation of 2CP or 3-CP by enrichment cultures from freshwater sediments under methanogenic conditions. They reported generation of methane along with accumulation of intermediate product like phenol and benzoate. Anaerobic degradation of phenol found to be initiated with carboxylation at the para position of the phenol ring followed by dehydroxylation to yield benzoate as an intermediate product. In fact, biodegradation of 2-CP occurs with carboxylation and dehydroxylation processes prior to under goes reductive dechlorination, consequently accumulation of 3-chlorobenzoate as an intermediate product (Becker et al.1999). Similarly, several existing literature reported the incomplete anaerobic degradation of nitrophenol by methanogenic consortia and accumulation of their respective toxic metabolic intermediate viz, aminophenols (Karim and Gupta 2003). In another study, Haghighi-Poden and Bhattacharya (1996) reported that, 16 mgl -1 of 3-NP and 8 mgl-1 of 2, 4-DNP causes inhibition to methanogenesis. Moreover, biodegradation of substituted phenols by anaerobic bacteria species not found to be a promising method due the following limitations; (i) They could not use these phenolics as a sole source of carbon and needs external carbon sources to serve as electron donor for their growth and pollutants degradation activities (Delia et al. 2005). (ii) Anaerobic biodegradation of halogenated phenol occurs by reductive dechlorination, which is partially or completely inhibited by the presence of other electron acceptors such as sulfate, sulfite and thiosulfite, nitrate, O2 and CO2 that are commonly found in most of the contaminated wastewater (van Briesen et al. 2004). (iii) Simultaneous degradation of chlorophenol (reductive dehalogenation) and methane production is not favorable in anaerobic bacteria (Saia et al. 2007) as hydrogen demand increases and consequently a strong competition for hydrogen takes place between the two processes. (iv) Incomplete mineralization usually occurs as a result of which intermediates of lower substituted phenols such as tetrachlorophenols, trichlorophenols, tribromophenols and other recalcitrant compounds are found to be accumulated.


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COMBINED OR INTEGRATED ANAEROBIC AND AEROBIC SYSTEM FOR EFFECTIVE BIOREMEDIATION OF TOXIC POLLUTANTS Usually low strength wastewater with COD concentration less than 1000 mgl-1can be treated efficiently using aerobic microbial system. Recently, considerable progress has been made on anaerobic systems based on the concept of resource recovery viz. generation of bio-fuel and biogas even at a very high strength wastewater with COD concentration more than 4000 mgl-1. In addition, production of minimum sludge volume with consumption of less energy and release of less volatile pollutants in to the environment are some of the major advantages of anaerobic system. However, as discussed above anaerobic treatment experiences poor biomass yield, biomass settling rate and very often process instabilities. Moreover, anaerobic process, necessitate for post treatment of its effluent which often contains various lethal intermediate pollutants such as un-degraded polyaromatic organic compounds, hydrogen sulfide and even ammonium ion (NH4 +) (Supaka et al. 2004). Although anaerobic process resulted higher pollutant removal efficiency, however, stabilization of organic pollutant is not feasible completely. Therefore, aerobic post-treatment is indispensable to polish the anaerobic effluents as a result a very high overall treatment efficiency can be achieved using combination of anaerobic–aerobic systems (Gray 2005). Combination of anaerobic–aerobic systems have been successfully employed for treatment of a wide range of industrial wastewater such as textile, leachate, pulp and paper, pharmaceutical and primary municipal wastewater. Table 2 presents the lists of the anaerobic–aerobic high rate bioreactors systems. Higher COD removal efficiency can be achieved by employing high rate anaerobic–aerobic bioreactors system even at a shorter HRT ranging from few hours to few days. Several authors evaluated the potential of UASB/aerobic CSTR system for treatment of a wide range of real industrial wastewater contaminated with BOD/COD ratio in the range of 0.17–0.74 (Sponza and Demirden 2007; Agdag and Sponza 2005; Tezel et al. 2001). Table 2, present that the higher COD removal efficiency can be achieved in the range of 50–100% (BDL) at an initial COD concentration in the range of 2.89-8,150 mgl-1 employing the combined anaerobic and aerobic systems even at lower HRT varies from 2.2 hr to 150 hr. Lerner et al. (2007) reported a higher level of COD removal rate using an anaerobic– aerobic treatment system, for example, the effluent COD concentration of a combined anaerobic–aerobic reactor system found to be dropdown from 80-120 mgl-1 with compared to that of an aerobic activated sludge process (220–250 mgl-1). Similarly, the volume of sludge can be minimized to 45% by using a combined UASB–AFB system compared to that of an aerobic system (Yu et al. 2000). Ahn et al. (2007) investigated removal of COD from high strength wastewater; they achieved about 99% COD removal with an initial COD concentration varying from 6000-14,500 mgl-1 even at a relatively short HRT of 24 hr.


Table 2. Combined or integrated anaerobic–aerobic high rate bioreactor systems for degradation of industrial and synthetic wastewater Type of waste water

Type of reactor

Pulp and Paper industry Pharmaceutical industry Synthetic waste water contain phenol and thiocyanate Complex industrial effluent

UASB and CSTR UASB and CSTR SAAAMB reactor

Synthetic Textile ww PTA effluent Synthetic waste water 2-CP, 2,4DCP

AnFB and Air lift suspension reactor UAST and AFB AFFFBR and AS UASB and RBC

Influent COD/Pollutant (mgl-1) 5500-6600

Total COD removal%

Aerobic COD removal% -

Anaerobic HRT

Aerobic/total HRT(h)

91

Anaerobic COD removal% 85

5h

6.54

Tezel et al. 2001

3000

97

68-89

71-85

-

-

4,200–8,150

-

23–53%

Anoxic 90%

5 to 10 days

30–60

Sponza and Demirden 2007 Sahariah and Chakraborty 2013

3800

-

60-65

-

1.4-1.8h

2.2-2.5

Heijnen et al.1991

2700

80

50

60

10h

10

Yu et al. 2000

5000

96.4

62-64

90

1-1.12h

22-26

30 mgl-1 2-CP, 2,4-DCP

100

≥ 99

BDL

12h each UASB

23-28.8

Pophali et al. 2007 Majumder and Gupta et al. 2007


References

Table 2. (Continued) Type of waste water

Textile industry ww

Type of reactor

Influent COD/Pollutant (mgl-1)

Total COD removal %

Anaerobic COD removal %

Aerobic COD removal % 40-90

Anaerobic HRT

Aerobic/total HRT(h)

References

Packed 800-1200 50-85 30-65 12-72h 10 Kapdan and column Alparslan 2005 reactor and AS Diluted land fill SAA 1000-3300 85-95 Yang and Zhou leachate bioreactor 2008 Degradation of IAAGB 2.89–3.75 Total HRT Tartakovsky et al. Aroclor 1242 reactor 50.4 2001 Synthetic IAAGB OLR 2.89 3.75 95-98 62-95 0-33 Total HRT Shen and Guiot -1 -1 wastewater reactor mgl day 48h 1996 #2 Reactor Type: UASB, upflow anaerobic sludge bed; CSTR, continuously stirred tank reactor; AS, activated sludge; AFB, aerobic fluidized bed; SBR, sequencing batch reactor; FFB, fixed film bioreactor; AFFFBR, anaerobic fixed film fixed bed reactor; AnFB, anaerobic fluidized bed; Type of waste water (ww); HRT, Hydraulic retention time: h, hour; IALR, internal airlift loop reactor OLR, organic loading rate; SAA, Simultaneous aerobic and anaerobic bioreactor; IAAGB, Integrated anaerobic–aerobic granular biofilm reactor; SAAAMB, Sequential anaerobic –anoxic –aerobic moving-bed reactor; BDL below detection limit; Gac FBR, granulated fluidized bed reactor.



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Garbossa et al. (2005) developed a bench-scale radial anaerobic/aerobic immobilized biomass (RAAIB) reactor, where polyurethane foam cubes are used as bio-supporting materials. They reported 84% COD removal efficiency from an influent COD of 345 mgl-1 at HRT values in the range of 1.2-15.5 hr. Similarly, Yang and Zhou (2008) investigated the treatment of diluted landfill leachate wastewater with an initial COD concentration in the range 1000 to 3300 mgl-1 by employing a simultaneous aerobic– anaerobic (SAA) bioreactor system. They achieved an average COD removal efficiency of 94%. Shen and Guiot (1996) used an anaerobic–aerobic granular biofilm bioreactor and reported that the methanogenic activities of the anaerobic granular sludge remained unaltered even in presence of dissolved oxygen (DO) in the re-circulated fluid. However, increase in influent DO concentration the yield of methane found to be decreases from 64 to 42% with respect to the influent COD concentration. In the contrary, the rate of CO 2 production found to be increases from 0.23 to 0.39 liter (CO2) g-1 COD, which strongly revealing the aerobic degradation of the organic substrate. Despite of significant amount of COD removal achieved in aerobic process of the coupled reactor more than 62% of the initial COD found to be removed by anaerobic process. Pasukphun and Vinitnantharat (2002) investigated the treatment of textile wastewater by a combined anaerobic/aerobic bioreactor system and they achieved higher COD and color removal efficiencies when the duration of the anaerobic biodegradation process found to be higher than that of the aerobic biodegradation process. For instance; a higher COD removal can be achieved (90%) when the ratio of anaerobic/aerobic cycle is maintained at 17.5/2.5 hr with compared to COD removal of 87% when the ratio is maintained at 14/6 hr (Pasukphun and Vinitnantharat 2002). A similar result also reported by Kapdan and Oztekin (2006) in the treatment of synthetic textile wastewater. They achieved more than 85% COD removal at an optimum HRT of 12 and 11 hr., respectively for anaerobic and aerobic biodegradation process. However, with an increase in aerobic HRT from 19 to 20hr, 80% of the COD removal can be achieved due to aerobic phase whereas, COD removal under anaerobic conditions found to be only 50%. On the contrary, with increase in anaerobic HRT, the COD removal due to the aerobic phase found to be negligible. This insignificance COD removal performance at lower HRT of aerobic phase may be due to the influx of toxic metabolic end product of the anaerobic phase to the aerobic phase which inhibits the metabolic activities of the aerobic culture. In literature, treatment of polycyclic aromatic hydrocarbons and highly chlorinated solvents using combined anaerobic and aerobic free or co-immobilized cultures have been applied successfully by many investigators. Zheng and Li (2009) investigated removal of phenol and COD in an anaerobic–anoxic–aerobic system and reported that with increase in HRT the phenol and COD removal efficiencies of anaerobic phase increases from 29 to 38 and 27 to 38%, respectively. Anaerobic–aerobic granular biofilm bioreactor has been implemented for removal of varieties of chlorinated pollutants such as trichloroethylene (Tartakovsky et al. 2005) and polychlorinated biphenyl (PCB) (Tartakovsky et al. 2001). The


138


biodegradation of chlorinated pollutants under oxygen-limited conditions in a biofilm bioreactor take place based on the co-existence of aerobic methanotrophic and anaerobic methanogenic bacteria. The sequence of biodegradation pathway is given as follows: for chlorinated aromatic hydrocarbons dechlorinations is occurred in anaerobic phase followed by aerobic ring cleavage (Supaka et al. 2004). Similarly, removal of nitrogen comprises nitrification in aerobic phase followed by denitrification in anaerobic phase (Liu et al. 2008). Uygur and Kargi (2004) evaluated the performance of phenol removal using a four-step sequencing batch reactor with combined anaerobic/aerobic process. They achieved about 80% COD removal efficiency with an initial phenol concentration of 600 mgl-1. However, combined or integrated anaerobic and aerobic systems have some limitations such as effluents of an active anaerobic digester enter directly into aerobic reactor system as a result huge quantities of obligate anaerobes along with facultative microorganisms enter into the aerobic reactor systems which are not adopted properly with the aerobic conditions. Consequently, these active anaerobes very often distress microbial population of aerobic reactor and lead to a mixed microbial population with poor oxygen consumption as well as COD removal activity. On the other hand, it increases the suspended solid concentration or the turbidity of effluent wastewater. Therefore, optimization of the anaerobic biomass to aerobic system is very crucial. However, the design, operation and process development of combined anaerobic–aerobic bioreactors are more complicated and expensive. Therefore, the combined anaerobic– aerobic bioreactors are still in its infancy and only few reports are available in literature. Proper design of integrated bioreactor system, optimization of process parameters and installation of methane gas capturing system, the combined or integrated anaerobic– aerobic bioreactor system can be appear as a feasible technology for treatment of high strength industrial wastewater, in addition, greenhouse gas emissions can be minimized.

CONCLUSION AND FUTURE PERSPECTIVES Aerobic microbial treatment plays a major role in the treatment of wastewater containing substituted phenols. In order to meet strict compliance for the pollutants removal from wastewater, considerable attention is focused on combined anaerobicaerobic bioreactor technology. Among the different microorganisms, bacteria belong to the species actinomycetes are very efficient for degradation of substituted phenols from contaminated wastewater. Furthermore, the actinomycetes degrade pollutants both aerobically and anaerobically, and therefore can be employed in combined or in integrated reactor systems for the treatment of phenolic wastewater. However, most of the combined anaerobic and aerobic bioreactors need to be tested at large-scale and its performance evaluated by relevant industries for establishing its full potential.



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Chapter 6

THE POTENTIAL OF BIOSORPTION TECHNIQUES IN HEAVY METALS MITIGATION Akhilesh Bind, Sushma Ahlawat*, Veeru Prakash and Somya Agarwal Department of Biochemistry and Biochemical Engineering, Jacob Institute of Biotechnology and Bioengineering, Sam Higginbottom University of Agriculture, Technology and Sciences, Allahabad, India

ABSTRACT Biosorption is a new and economically well-defined process for the mitigation of heavy metal ions from the sources like waste water, basically based on the property of diffusion/intraparticle diffusion metal ions through the given porous structure of sorbent and effects arise due to the resistance and electro-repulsive interaction of ions within the sorbent which utilizes dead biomass for removing heavy metals effective even at trace level. Rapid industrialization, urbanization and technological development contribute in more waste generation and pollution of natural resources. Heavy metals including Ni, Co, Pb, Cd, Zn, Ar, Cu, Cr, Fe, Hg, Sr, and many more are carcinogenic in nature when consumed beyond their optimum limits needed by the body of humans, animals, and even by plants. Diverse categories of biosorbents like plant waste materials (lignocellulosic), microbes, algae, fungi, animal’s components, resins, activated carbon compounds etc. has been exploited and found to be very promising. Biorsorbent characterization utilizes Fourier Transform Infrared Spectroscopy (FTIR), Transfer Electron Microscopy (TEM), Scanning Electron Microscopy (SEM), Spectrophotometer, Thermo-gravimetric analysis, Surface titrations, Adsorbent Response Surface Methodology, X-ray Photoelectron Spectroscopy (XPS), Electron Diffraction X-Ray (EDX), X-Ray Fluorescence, Response Surface Methodology (RSM), Confocal Laser Scanning Microscopy (CLSM). There are *



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Akhilesh Bind, Sushma Ahlawat, Veeru Prakash et al. various mathematical models like Ideal Adsorption Solution Theory (IATST), Bangham’s model, Elovich model, Temkin, Langmuir and Freundlich isotherm models, Dubinin-Radskenich model, Redlich-Peterson and Sips model etc. are used for predicting optimized condition of heavy metal removal.

Keywords: biosorption, heavy metals, mitigation, industrialization, FTIR, SEM

INTRODUCTION Water is one of the most valuable natural resource available to human being but in limited resources. Water covers 71% of the Earth’s surface and is important to all the forms of life. Fresh water content is about 2.5% of the Earth’s surface and 98.8% of this water is locked in ice and ground water. Only 0.3% or less than this fresh water is in rivers, lakes and in atmosphere. Being recognized as vital source of existence, its pollution is common problem faced today which needs to be addressed through researchers for human wellbeing and environment. Heavy metal pollution in environment is a global problem because these cannot be deteriorated or destroyed and hence tends to gather in aquatic ecosystems, soils and sediments. The highly electronegative metals with a density greater than 5g/m3 are known as heavy metal. Naturally occurring heavy metals in the Earth are becoming concentrated as a result of human activities due to wastes produced from industries, mining, automobile emissions, batteries containing lead, fertilizers, pesticides, paints, woods, manufacturing of electric goods, treating of metal surface etc. Pollutants are being discharged into the atmosphere causing hazardous environmental pollution and are also fatal. Rapid urbanization and industrialization substantially increases pressure on the quality of available water, thereby threatening the usability of these supplies. Today industry uses more than 40% of total water withdrawals in developed countries; the comparable figure in developing countries is less than 10%. Thus, industrial development will also lead to more water quality problems. With such a high demand for the limited quantity of potable water it is necessary to prevent or at least limit its contamination with pollutants (Harvey et al., 2015; Volesky, 1990; Wang, 2002; Prabhakar, 2001; Vijayaraghavan and Yun, 2008). Amongst substances which are reaching toxic and hazardous levels are heavy metals (Vieira et al., 2000). In general, water pollutants are classified as organic and inorganic. Organic water pollutants include solvents from industries, organic volatile chemicals and wastes from food industries. Inorganic water pollutants are like metals, fertilizers, industrial discharges responsible for acidity etc. Heavy metals are of interest which fall under inorganic group and are mostly produced through anthropogenic source through industrialization (Volesky, 2007). Over the past decade, heavy metal pollution through anthropogenic activities and industrialization has caused great concern about environmental protection. Heavy metal accumulation through soil, aquatic bodies and industrial effluent poses a serious health


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problems due to their toxicity at even trace level. Recently a newspaper “THE HINDU” reported ‘a dip in Ganga can cause cancer’ which has been declared by the department of Atomic Energy’s Centre for Compositional Characterization of Materials (NCCM) which collected and tested water samples from Kumbhmela in January 2013. Of the essential metals like lead, mercury, chromium, cadmium, arsenic are poisonous in nature whereas large scale usage and more concentration of nickel, copper, cobalt, are also destructive to the environment. Radionuclides like Thorium, Uranium and Strontium exhibit threats (Moore, 1990; Gadd, 2009). Kofi Annan, Former Secretary-General of the United Nations, in 2001 said, “Fierce competition for fresh water may well become a source of conflict and wars in the future”. Therefore, there is an urgent need of nascent methods for removal of metals from waste-water and bring it down to regulatory level, especially from the environmental pollution control point of view. Intensive research have already been undertaken which reflect biomass dead as biosorbent is more effective rather than the living active one in mitigating heavy metals. Past two decades had been focused on identifying available non-living biomass which has the potential to remove heavy metals. Biosorbents including algae (Davis et al., 2003; Figueira et al., 2000), fungi (Gao et al., 2009), bacteria (Volesky and Holan, 1995) and waste or unused materials form cultivable land (Sudet al., 2008) are well studied and reported. The study of biosorbent’s metal uptake efficiency is important as an application of biosorption to industries which is influenced by various parameters like solution’s pH, dose of biosorbent, initial ion concentration, size of the particle and their concentration, temperature, contact time, speed of agitation and competing ions (Hasan et al., 2008; Singh et al., 2005; Wang et al., 2008).

MECHANISM OF BIOSORPTION Biosorption process involve bonding with chemical species to solid phase (biosorbent) and a liquid phase (solvent generally water) which contains dissolved species, that has to be sorbed (sorbate metal ions) and because of higher affinity of sorbent (Das et al., 2008). The metal ions get removed not based on only through one mechanism but also due to several mechanisms which are frequently complex and independent of metabolism. Various mechanisms involved differ qualitatively and quantitatively according to the used species, biomass origin, and its processing metal sequestration mainly follows mechanisms like ion exchange (Tenorion et al., 2001), chelation, adsorption by physical forces; and ion entrapment in inter- and intra-fibrillar capillaries and spaces present in the structural polysaccharide network which is done by diffusion the cell wall and membranes and the concentration gradient (Volesky and Holan, 1995, Febrianto et al., 2009). Biosorption or metal isolation can occurs through following approaches.


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Adsorption Adsorption is a mass transfer process which involves the accumulation of substances at the interface of two phases, such as, liquid–liquid, gas–liquid, gas–solid or liquid–solid interface. The substance being adsorbed is the adsorbate and the adsorbing material is termed the adsorbent. Interaction between solid surface and adsorbed molecules is called physisorption and chemisorption in case of having physical nature and chemical nature respectively. In physisorption, the attraction interactions are van der Waals forces and, as they are weak the process results are reversible. Contrary to physisorption, chemisorption occurs only as a monolayer and, furthermore, substances chemisorbed on solid surface are hardly removed because of stronger forces at stake. Under favorable conditions, both processes can occur simultaneously or alternatively (De Gisi et al., 2016). It is a new and economically well-defined method for the removal of heavy metal ions from the sources like waste water, basically depending on the property of diffusion of metal ions through the given porous structure of sorbent and effects arise due to the resistance and electrorepulsive interaction of ions within the sorbent which utilizes dead biomass for removing heavy metals effective even at trace level. Individual or mixture of metal binding mechanism is functional in many ways in immobilizing metal species on the sorbent. Various chemicals functional groups among biosorbents are considered for eliminating metal solutions. Mostly sorbents possess different forms of ligands that participate actively in metal binding through carboxyl, phosphate, sulfhydryl, amine, hydroxyl groups on their surfaces. Biosorbents application in reducing the recalcitrant toxic pollutants and for the recuperation of high value heavy metals from aqueous wastewaters is latest development in environmental or bioresource technology (Vijayaraghavan and Yun, 2008; Volesky, 2007; Aksu, 2005). Most important advantage of adsorption technology as compared to conventional methods is low cost with high efficiency, less byproduct formation in form of chemical or biological sludge, and its reusability and regeneration or metal recovery (Volesky, 2007).

Complexation Biosorption can be done through interaction of various metal ions and functional groups followed by coordination complex formation. Various researchers have shown that heavy metal adsorption by lignocellulosic biomass is due to ion exchange and process complexation. Coordination complex is formed by the metal and nitrogen or oxygen of chitin. Biosorption of radioactive elements like Thorium and Uranium involve adsorption as well as complexation as similar hypothesis is given by Aksu et al., (1992) on the biosorption of copper by C. vulgaris and Z. ramigera takes place by adsorption and complex formation.



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Ion Exchange The replacement of an anion in solid phase in contact with a cationic solution by another ion is called ion exchange. The basic units of the cell wall of the natural biopolymers and divalent metal ions i.e., different polysaccharides get replaced with the contradictory polysaccharide ions (Tsezos and Volesky, 1981).

STUDY OF BIOSORPTION MECHANISM Biosorption mechanism depends on the concentration of metallic ions and biosorbent cell surface properties which involve different functional group elucidated by Fourier Transform Infrared Radiation spectroscopy. It has already been attempted by various researchers to characterize and report about different functional groups present in biomass responsible removal or so called mitigation of heavy metal and illuminate the various mechanisms involved in metal biosorption.

Chemical Composition Studies of Biosorbents Different types of active and inactive organisms have been used as biosorbents to remove heavy metal ions from aqueous solutions. Biosorbents are rich in organic ligands or the functional groups and thus play a major role in removing various heavy metal contaminants. Carboxyl, hydroxyl, sulfate, phosphate, and amine groups are some of the major functional groups present on participating biosorbent surface. Synergism, antagonism and non-interaction are the phenomenon which influences the adsorption of heavy metals due to presence of other ions and organic materials in wastewater. This study showed that copper was well adsorbed from wastewater with multiple heavy metal ions in comparison with wastewater containing single Cu (II) ion, whereas cadmium adsorption was inhibited by other metal ions and zinc removal remains unaffected (Abdolali et al., 2014). Numerous researchers on biosorption have expounded on the chemical composition of biosorbents, which suggests the major role of these components in metallic bio-sequestration. For example, the main components of coir having lignin, pectin, waxes and hemi-celluloses, celluloses etc. with variation in the composition as mentioned in literature. This reveals about the network of main component, as the main binding sites, showing the presence of low or high covalent character (Drake et al., 1996).


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Heavy Metal Quantification The heavy metals quantification after the batch shake experiments is reported in the form of percentage removal as per equation given below. Heavy metal concentrations can be analyzed as per the American Public Health Association standards (APHA, 2005). 𝐻𝑒𝑎𝑣𝑦 𝑚𝑒𝑡𝑎𝑙 𝑟𝑒𝑚𝑜𝑣𝑎𝑙 (%) =

(𝐶𝑜 − 𝐶𝑒 ) × 100 𝐶𝑜

where C0 and Ce are the initial and final heavy metal concentrations (mg L-1) present in the solutions, respectively.

Regeneration of the Exhausted Biosorbents and Cost Estimation For determining a process to be cost-effective, regeneration of heavy metals study should be undertaken with various acid, base for example by using 0.5 N HCl and 0.5 N NaOH solutions as the stripping agents. Metal loaded biomass after the biosorption are further transferred to the flasks and kept in an orbital shaker for 24 h. Following this, filtrates are further analyzed to examine the percentage fraction of desorbed Cu (II), Cr (VI), As (III) and Pb (II) metallic ions to examine the percentage fraction of desorbed. Successive biosorption–desorption cycles repetitions were performed for the optimal use of biosorbents (Kaur et al., 2013). Desorption efficiency for both the biosorbents are further evaluated from the amount of metal ions desorbed to the metal ions adsorbed in the desorption medium, as per the following expression (Yoonaiwong et al., 2011). 𝐷𝑒𝑠𝑜𝑟𝑝𝑡𝑖𝑜𝑛 𝑒𝑓𝑓𝑖𝑐𝑖𝑒𝑛𝑐𝑦 (%) =

𝐴𝑚𝑜𝑢𝑛𝑡 𝑜𝑓 𝑚𝑒𝑡𝑎𝑙 𝑖𝑜𝑛 𝑑𝑒𝑠𝑜𝑟𝑏𝑒𝑑 × 100 𝐴𝑚𝑜𝑢𝑛𝑡 𝑜𝑓 𝑚𝑒𝑡𝑎𝑙 𝑖𝑜𝑛 𝑎𝑑𝑠𝑜𝑟𝑏𝑒𝑑

Sorption Isothermal Modelling Biosorption system uses various isotherm equations for the equilibrium modelling like Langmuir, Freundlich, Dubinn-Radushkenich and Temkin isotherms are mostly adopted for data analysis for predicting the validity of process on scale up. As the initial metal ion concentration increases, the equilibrium distribution of metal ions of the aqueous and solid phases changes and the relationship between sorbed (qe) and remaining aqueous metal ion concentration (Ce) is evaluated. Langmuir equation is based on the following points (Langmuir, 1918): 1. Finite number of energetically uniform identical sites are presented by the solid surface



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2. No interaction is seen among the adsorbed species i.e., rate of adsorption does not get influenced by the amount of adsorbed species 3. Formation of a monolayer after the reaching of the solid surface at the saturation point 4. Langmuir equation refers to (Langmuir, 1918) qe = (qmax b Ce)/(1+b Ce )

(1)

where qe (mg g-1) is defined as the equilibrium sorption capacity and Ce (mg L-1) is stated as the equilibrium concentration and qmax(mg g-1) is called the maximum amount of metal ion per unit weight of adsorbent and b is a constant (L mg-1) in the given above equation. Langmuir equation can be expressed as: Ce/qe = 1/qmaxb + Ce/qmax

(2)

The values of qmax and b can be calculated by the help of plot of Ce/qe versus Ce. Now, Separation factor for the equation will be as follows:

RL 

1 1  bCo

(3)

As per Langmuir isotherm condition, when RL is less than 1, the data for favourable adsorption would be 0 < RL < 1. The Freundlich equation efficiently defines the adsorption on heterogeneous surface (Freundlich, 1906) and can be given as: qe = KF.Ce1/n

(4)

where KF and 1/n are called as Freundlich constants. Equation (3) can be modified as:

log qe = log KF +

slope

1 logCe n

(5)

1 and intercept ln KF can be determined from straight line in the plot. n

Assumptions in Temkin isotherm clearly signifies that the heat of adsorption of all the molecules present in a layer decreases linearly with coverage and their adsorption is



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characterized by the distribution of maximum binding energy (Temkin, 1934; Temkin and Pyzhev, 1940). Commonly, it is represented as:

qe 

RT RT ln( AT )  ln(Ce) bT bT

(6)

Linear form of Temkin isotherm is linearized as:

qe  B ln( AT )  B ln(Ce)

(7)

The isotherm constants AT and B can be computed from plot of qe versus lnCe, where B = RT/ b represents heat of adsorption of the adsorbed molecules, T is absolute temperature (in Kelvin) and R is the universal gas constant, 1/b and AT (L mg-1) are the constants. Dubnin-Radushkevich Isotherm is used to describe the characteristic porosity of the biomass and the apparent energy of adsorption of various adsorbed molecules (Dubinin and Radushkevich, 1947). The Dubnin-Radushkevich isotherm equation is expressed as: qe = qm e-2

(8)

where q is mg sorbate g-1 sorbent, qm (mg g-1) is theoretical sorption capacity of adsorbed molecules. The Polanyi sorption potential  is equal to  = RT ln(1+1/Ce)

(9)

R is the gas constant (8.314 J mol-1 K-1) T(K) is absolute temperature, Ce is the equilibrium concentration of sorbate in solution (mg L-1). The constant  is related to mean free energy, E (KJ mol-1) of sorption per mole of the sorbate (Hobson, 1969; Dubey and Gupta, 2005): E = √2

(10)

This signifies the idea about the adsorption mechanism of the heavy metals adsorption with the help of this equation. The value of E between 8 and 16 KJ mol -1, represents chemical ion exchange between the molecules while for E< 8 Kjmol-1, the biosorption process is physical in nature as per the assumption of Dubinin and Radushkevich (1947) and Onyango et al., (2004).



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Kinetic Modeling Various kinetic models, able to describe the kinetics of heavy metals biosorption, are basically the batch systems. Lagergren’s pseudo first order equation is used to understand the regulation and mechanism of biosorption of metals through various biosorbents and many equations too. Some of them are the pseudo- second order equation given by Ho and Mckay (2000), intra particle diffusion form of Weber- Moris relation (Weber and Moris, 1963) and second order rate of equation described by Nabizadeh et al., (2005). The pseudo first order Lagergren model is expressed as: dqt/dt = K1(qe-qt)

(11)

where K1 (min-1) is constant of pseudo-first-order adsorption model, qe and qt (mg g-1) gives the amounts of meta cations sorbed on biosorbents at equilibrium and time t (min). Integration of above equation at, qt = 0 at t = 0 and qt = qt at t = t formulized as: log (qe- qt ) = log qe – K1t/2.303

(12)

Constant K1 and qe can be observed from the slope and intercept by plotting log (qeq) versus time t plot. Pseudo second order rate expression can be represented as follows: dqt/dt = K2 (qe-qt)2

(13)

where K2 is the rate constant (g mg-1 min-1). Integrating the equation (13) and applying qt = 0 when t = 0 and qt = qt at t = t, leads to the equation given below: t/qt = 1/ K2qe2 + t/qe

(14)

The values of qe and K2 can be obtained from the slope and intercepts of plots of t/qt versus time t. The kinetic information provided by this model is approachable (Ho and Mckay, 1998; Deng et al., 2006; Loukidor et al., 2004) to search important parameters regarding the design of bioreactor (Cruz et al., 2004). Weber and Morris (1962) proposed the intra particle diffusion model by the help of which initial rate of intra particle diffusion is calculated which is represented as: qt = Kid t 1/2 + C


(15)

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where C is intercept and Kid is the intra-particle transport rate constant. Value of qt is sorbed concentration (mg g-1) at time t (Ayyappan et al., 2005). Second-order rate equation is linearized as:

1 1   k .t qe  qt qe

(16)

where qe and qt are metal cations adsorbed (mg g-1) at equilibrium and time t (min) whereas k◦ is constant of adsorption (g mg-1 min-1).

Thermodynamics of Adsorption Gibbs free energy, enthalpy and entropy are the concerned thermodynamic parameters which show the temperature dependence of the biosorption process.

CONTRIBUTORY TECHNIQUES USED IN BIOSORPTION STUDIES Several sensitive analytical techniques play an important role in the study of metal cell interaction for the identification of metal biosorption mechanisms. Biosorption process encompasses the interaction between the biosorbent (solid phase) and a solvent (liquid phase) containing the dissolved species. The mechanism of biosorption is complex in nature which can be explained by several mechanisms viz. physiosorption, chemisorption, micro-precipitation, ion exchange and chelation. The efficiency of metal ion binding is highly dependent on the constituents of adsorbent. The outer surface of biomass contains polysaccharides and proteins that offer active sites for binding of metal ions in aqueous media. In metal binding mechanisms, functional groups enhance the affinity of metal ion and form weak or strong bonds with metal ions. The mechanism was verified by the presence of different functional groups such as ketones, aromatic amines, carboxylic group, ethers, aldehydes and phenolic groups (Ahluwaliya and Goyal, 2007). Cellulose, hemicelluloses, lignin and pectin with small amount of protein are also present in plant biomass. These materials help in the binding of heavy metals on the surface of biosorbents. The mechanism of biosorption can also be revealed by the study of Fourier Transform Infra-Red (FTIR) spectroscopy, X-Ray Photo Electron spectroscopy (XPS), Scanning Electron Microscopy (SEM), Energy Dispersive X-ray (EDX) Fluorescence spectroscopy and elemental analysis (Kaur et al., 2012, Arief et al., 2008). Some of the techniques efficiently used in biosorption of heavy metals are described below.



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Fourier Transform Infrared (FTIR) Spectroscopy Understanding the mechanisms of the solute adsorption onto the solid surface is essential for the certain removal of contaminants from aqueous solution. The adsorption mechanism involves oppositely charged ionic interaction such as dipole– dipole, dipoleinduced dipole and induced dipole-induced dipole, hydrogen bonding, chemical bonding, and ion exchange. The surface chemistry of the adsorbent and its effect on the adsorption process is generally investigated in order to interpret the solute adsorption. Fourier transform infrared spectroscopy (FTIR) spectroscopy is a useful tool for studying the interaction between an adsorbate and the active sites on the surface of the adsorbent. The interpretation of the FTIR is based on the chemical structure of the raw date pits (RDP). The RDP consists of three main components, namely cellulose, hemicellulose, and lignin, besides other minor constituents such as oil, protein, etc. Both cellulose and hemicelluloses contain majority of oxygen functional groups which are present in the lignocellulosic material such as hydroxyl, ether and carbonyl, while lignin is a complex, systematically polymerized, highly aromatic substance acting as a cementing matrix that holds between and within both cellulose and hemicellulose units. The presence of various functional groups and their complexation with heavy metals during biosorption process is studied by using spectroscopic techniques, such as FTIR and XPS. These are widely used in the area of natural products, organic synthesis and transformations to study the qualitative analysis of organic compounds. It is also used to search the functional groups present in the substances (Kaavathy et al., 2012) particularly with the availability of the main groups which are involved in adsorption. FTIR spectroscopy is used to study the nature of binding sites and their contribution. Jaafarzadeh et al., (2015) used chitin shrimp shells for the removal of arsenic (V) and zinc (II) ions as low cost biosorbent. They confirmed the presence of different groups on the cell surface of chitin shrimp shells for arsenic (V) and zinc (II) ions biosorption using FTIR spectra. Das et al., (2014) obtained FTIR spectra of Pleurotus platypus for Zn (II) adsorption were to gain an insight on the functional groups involved which was found to be amines (N-H stretch: 3441.01– 3410.01 cm−1), alcohols (O-H stretch: 1402.25 cm−1; 1382.96 cm−1 and C-O stretch:1068.56 cm−1 and 1043.49 cm−1), carboxylic acids (C = O stretch: 1662.04 cm−1) and ketones (C- CO- C bend: 542.00–602.58 cm−1) which played a major role in the biosorption process. Pilli et al., (2010) studied biosorption Cr(VI) onto Hydrilla verticillata weed and conducted FTIR spectrum analysis for the characterization of functional group responsible for heavy metal removal and found that Cr(VI) loaded biosorbent revealed that O–H, N–H and C–O groups were the leading Cr(VI) binding groups available in biosorbent. The functional groups responsible for biosorption of heavy metal ion on to chlorella species were confirmed by FTIR spectra. The presence of amino, carboxylic, hydroxyl and carbonyl groups of Chlorella are depicted using this technique only by Kanchana et al., (2011). FTIR spectra also showed that chemical


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reactions did not include gold biosorption by Cladosporium cladosporiodides. For this, gold anions (AuCl4) were bound to the protonated carbonyl and carboxyl groups by electrostatic bond interactions in between them. The C-N, N-H, OH and NH3+ moieties are involved in Ag, Cu, Cd and Pb biosorption.

X-Ray Photoelectron Spectroscopy (XPS) It is used to investigate the surface chemistry, due to the presence of functional groups, of solids in a nondestructive manner and is considered to be one of the most powerful techniques. It also provides oxidation state information of the atoms (Gupta et al., 2000). The two oxidation states of iron, Fe2+ and Fe3+, were observed by XPS (Figueira et al., 1999). FTIR spectroscopic analysis proved the requirement of carbonyl groups for uptake of both iron species whereas sulphonates were found involved in the ferric ion biosorption. Krishnani et al., (2008) used this technique to estimate the quantification of metals sorbed in the given samples having a large amount of absorbed metals. Relative abundances determination and species recognition is done by this technique only. Surface chemistry of lignocellulosic biomatrix of rice is determined by the XPS technique. The spectra of lignocellulosic biomass was found to consist of C and O as its major components while other elements were Ca, Mg, Na, and Si. In other study, reduction of Chromium (6) biosorbed by chromium treated biomass on to the biomaterial is done in its trivalent form. Studies done by XPS on biomass washed with acetone of S. unarum by Ashkenazy et al., (1997) reports the involvement of oxygen and nitrogen containing functional groups in lead biosorption.

Scanning Electron Microscopy (SEM) Scanning electron microscope produces the sample image by scanning the electron beam which is read along with the signals to draw the image of specimen. This technique can provide resolution better than 1 nanometer. Tarley and Arruda (2004) described the morphological facts of rice husk using the technique SEM to show how uneven surface of rice husk helps in sorption of metals. Oliveira et al., (2008) produced the images of coffee husks before and after the biosorption of metal cations to explain important morphological differences. SEM micrograph of native coffee husks is shown to have rough patches indicating oddly disseminated abrasions in the husk skin converting to smooth surface after Cu2+ sorption. Whereas as morphology shows an interrelated network of veins regarding adsorption of chromium. Differences in ionic states and sorption mechanisms were attributed by the authors. Simple cationic exchange mechanisms for deposited metal ions and chromium oxyanions were proposed.



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Transmission Electron Microscopy (TEM) This technique gives an idea about localization of sorbent in cell providing important information about sorption mechanism of metals. Various researchers have used TEM in combination with other advanced techniques to explain this mechanism of adsorption. Zhou et al., (2005) produced the TEM images of Pb2+adsorption of before and after the process which very clearly reveals the lead sorption on beads occurs on the whole area of beads. He used these advanced techniques to explain three concurrent mechanistic processes including complex formation among Pb2+ and nitrogen, followed by its adsorption with hydrolytic product precipitation in the form of crystalline beads very clearly. Suh et al., (1998) studied the accumulation of Pb2+ in S. cervisiae with the help of TEM after varying exposure intervals to sorbate lead solution, in which first step is a quick binding to the cell wall. Passive transportation of Pb2+ by the cell wall for a short time duration 3-5 minutes whereas the second step is the penetration through the cell membrane into the cell’s cytoplasm.

FUTURE PROSPECTS AND CONCLUSION OF TECHNOLOGIES USED IN BIOSORPTION OF HEAVY METALS Instead of being at developmental stage, Biosorption technology seems to be promising in heavy metal mitigation. It is anticipated to develop common-purpose biosorbents that can remove a variety of pollutants. One such possibility would be the use of ‘combo’ biosorbents consisting more than one type of biomass. Biosorption technology is at nascent stage and work in this area is necessary to validate its possibilities on an industrial scale. So far, biosorption research has mainly focused on removal of pollutants such as heavy metals and organic compounds. In current scenario, due to limited deposit and high cost, precious metals are explored by biosorption process. For metals such as gold, platinum, palladium, ruthenium, etc. recovery, biosorbent properties and its performance-effectiveness considered to be paramount important than cost-effectiveness. Recovery efficiency and purity of final recuperated products would be auxiliary criteria for evaluating biosorbents and related processes. It is expected that in the near future high-performance biosorbents will be used for the analytical purposes. Biosorbents can also be also used for the purification of ionic pharmaceuticals like proteins, antibodies, and peptides etc. Column chromatography would be more effective approach for the high-purity products than fixed or moving bed adsorption of heavy metals for such purposes. In future, therefore, more attention should be given to these areas. Although biosorption is one of the main components of environmental biotechnology, to the best of


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our knowledge, there is neither an international conference nor a global network for researchers in this area. Needless to say, a weak relationship between researchers slows down development of biosorption technology and delays its commercialization. Further technological upscale feasibility has to be investigated for performance enhancement and cost reduction. Industrial pollution consists of more than one metal which is in contrast to laboratory solution, hence simultaneous removal of metal of interest should be developed.

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Chapter 7

AZO DYE REMOVAL TECHNOLOGIES Maulin P. Shah* Industrial Waste Water Research Lab Division of Applied and Environmental Microbiology Enviro Technology Limited, Ankleshwar, India

ABSTRACT The management of large amounts of wastewater effluents containing industrial textile coloring agents is a challenge as it deteriorates water quality that disrupt aquatic life and constitute a serious threat to the public health. Furthermore, most of the dyes and their metabolic intermediates are mutagenic and carcinogenic. The synthetic dyes are chemically synthesized contaminants in the biosphere that inhibit mineralization of wastewater. The azo dyes are the most important class and the largest commercial dyes that account for nearly 75% of all products of textile dyes. Physicochemical methods used to treat the wastewater of dying process have disadvantages such as excessive use of chemicals and the production of sludge which disposal obviously cause problems. Bioremediation for the removal of dyes is gaining interest because it is obviously profitable, environmental friendly and produces much less sludge. Therefore, for the successful implementation of a process of biological treatment of wastewater to remove azo dyes have been developed and deeply discussed in this chapter.

Keywords: Azo dye, textile due, decolorization, eutrophication, bioremediation, biological treatment

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INTRODUCTION Since the beginning of civilization, dyes have been used by the peoples for the painting and dyeing of their surroundings, skins and clothes. All the colorants used before the middle of 19th century were of natural origin. Inorganic pigments such as soot, manganese oxide, hematite and ocher were used in living memory [1]. Paleolithic rock paintings, such as the 30,000 years old drawings recently discovered in the caves of chauvet, France provides an ancient testimony of their applications. Natural organic dyes also have a timeless history of application, in particular as textile dyes. These all dyes are aromatic compounds, usually derived from the plants, but also from the insects, mushrooms and lichens. The manufacturing of synthetic dyes began in 1856, when the English chemist W. H. Perkin was trying to synthesize quinine and get bluish substance with excellent dyeing properties that was later recognised as violet aniline, purple tyrant or mauveine [2]. Perkin patented this invention and set up a production line. This concept of research and development was soon to be followed by others and new dyes began to appear in the market. At the beginning of 20th century, almost all the synthetic dyes are completely replaced by the natural dyes [3].

CLASSIFICATION OF DYES Dyes are molecules containing chromophore and auxochrome. The chromophore is a group of atoms, which controls the color of the dye, and it is generally an electron withdrawing group [4]. The most important chromophores are -C = C-, -C = N-, -C = O, N = N-, -NO2 and -NO. The auxochrome is an electron donor substituent which can intensify the color of the chromophore by modifying the overall energy of the electronic system and providing the solubility and the adhesion of the dye to the fiber. The most important auochromes are - NH2, -NR2, -NHR, -COOH, -SO3H, -OH and -OCH3 [5]. On the basis of the chemical structure or the chromophore, 20 to 30 different dye groups can be identified. Azo, anthraquinone, phthalocyanine and triarylmethane dyes are quantitatively the most important chromophores (Figure 1).

Figure 1. Most important chromophores.


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DYE REMOVAL TECHNOLOGIES The textile industry has been subjected to immense pressure to reduce the use of harmful substances, in particular mutagenic, carcinogenic and allergenic chemicals and textile dyes. The sanitation of textile’s dye sewage relies not only on the elimination of colors but also in the degradation and mineralization of dye molecules. A wide range of technologies have been developed for the removal of synthetic dyes from wastewater to reduce their environmental impact [6]. These technologies include physical methods (membrane filtration processes and sorption techniques), chemical methods (coagulation or flocculation and conventional oxidation processes), and biological methods (microbial and enzymatic degradation). The above mentioned technologies having their own merits and demerits [7]. The colors are the first pollutants to be recognized in waste water and must be removed before discharged into the watercourses. Even the presence of very small amount of dyes in water affects visual quality of water. In comparison to colorless organic compound, removal of color compound is more important that burdens a larger portion of biochemical oxygen demand. The methods for removing biochemical oxygen demands from most waste water are relatively well established [8]. To remove synthetic dyes from wastewater, a wide range of methods have been developed as described in the following sections. Table 1. Advantages and disadvantages of dye removal techniques Physical/ Chemical Methods Electro kinetic coagulation Electrochemical destruction Ozonation

Advantages

Disadvantages

Economically feasible

High sludge production High cost of electricity

Membrane filtration

Breakdown compounds are nonhazardous Applied in gaseous state; no alteration of volume No sludge production Effective decolourisation of both soluble and insoluble dye Initiates and accelerates azo bond cleavage Good removal of a variety of dyes Removes all dye types

Ion exchange Irradiation

Regeneration; no adsorbent loss Effective oxidation at lab scale

Photochemical Fenton’s reagent NaOCl Activated carbon

Short half life Formation of bi-products Sludge generation Release of aromatic amines Very expensive to operate Concentrated sludge production Not effective for all dyes Requires a lot of dissolved O2


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The effluents from textile industries are the most expressive of an ecological and physiological point of view [9]. To achieve satisfactory and acceptable quality levels that allow the recycling of wastewater from textile, elimination of dyes and related compounds are very crucial. Based on the fact that azo dyes constitute the largest percentage of textile dyes, most treatment methods are based on the discoloration of azo dyes [10]. Currently, the main operational methods used in the treatment of wastewater involve physical and textiles chemical processes [11, 12]. There are several factors that determine each technical and economic feasibility of dye removal technology. These include dyes, composition of waste water, doses, running costs, environmental fate and handling [13]. Each dye removal technique has its own limitations and one single process may not be sufficient to achieve complete decolorization (Table 1). To overcome these problems, dye removal strategies have a combination of different techniques. Dye removal strategies therefore essentially consist of a combination of different techniques [14].

PHYSICO-CHEMICAL METHODS Membrane Filtration The increased cost of water consumption and wasteful require treatment process that is integrated with the circuits of water on the ground rather than as a subsequent treatment [15]. From this point of view, membrane filtration offers a potential application. The processes that use membranes offer exciting possibilities for separation of dyestuffs and dyeing auxiliaries that reduce simultaneously hydrolysed color and biochemical oxygen demand /chemical oxygen demand of wastewater [16]. The advantages of membrane filtration are its fast process with low space requirement. The disadvantage includes limited lifetime and high cost of membrane [17]. The choice of the membrane process, in the case of reverse osmosis, nanofiltration, ultra filtration or microfiltration, should be guided by the quality of the final products.

Flocculation and Coagulation These techniques are repeatedly used for the separation of suspended solids portion in the wastewater [18]. Suspended solids are remain in the wastewater and will settle out due to same type of surface charge on the particles that repel each other when they come close together. To overcome this problem, proper coagulation and flocculation techniques are used. In coagulation process, chemical coagulants bearing opposite charge, those of



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the suspended solids, are added to the suspension that neutralize the charges. Now the small suspended particles are capable of sticking together. Flocculation is a gentle mixing stage in which unstable particles bind with larger hairs to form a floc that can be easily removed from the suspension [19, 20]. Floc size continues to build with additional collisions and interaction with added inorganic polymers (coagulant) or organic polymers. Inorganic coagulants such as aluminum and iron salts are the most commonly used coagulants. Synthetic polyelectrolytes, fly ash and clay are also used as a thickening agents [21]. However, inorganic coagulants are not very much effective to remove highly soluble dyes [22]. The greatest disadvantage of using this process is the possibility of secondary pollution [23]. Recently, some organic polymers are reported that have good coagulation property [24].

Ion Exchange Standard ion exchange systems are not extensively compatible to utilize dye of wastewater because the ion exchangers cannot catch wide range of dyes due to presence of various additives in the wastewater [25]. In this technique, the waste water becomes an ion exchange resin until all available exchange sites are saturated. Both anionic and cationic dyestuffs were efficiently removed in this process. The major disadvantage of the process includes high cost of regenerating organic solvents from the ion exchanger [26]. Reverse Osmosis In reverse osmosis, membranes have retention rate of 90% or more types of ionic compounds and produces high quality permeate [27]. Degradation and elimination of chemical additives in dye waste water can be carried out in one step reverse osmosis. Reverse osmosis makes it possible to hydrolyze all mineral salts and reactive dyes and chemical aids. It should be noted that higher the concentration of dissolved salt, the more important the osmotic pressure becomes; therefore, the greater the energy required for the separation process. Nanofiltration Nanofiltration has been used to treat colored wastewater from the treatment of textile dye drainage water. The adsorption step precedes nanofiltration because this sequence reduces the concentration polarization during the filtration process, which increases process output [28]. Harmful effects of high concentrations of dyes and salts in dye house outflows have often been reported [29, 30, 31]. In most published studies on dye waste water, the concentration of mineral salts is not more than 20 g/l and the concentration of dye does not exceed 1.5 g/l [32]. In general, the wastewater is reconstituted with only one dye [33], and the volume studied is also low. The treatment of waste water by nanofiltration is one of the rare applications that are possible for the treatment of


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solutions with highly concentrated and complex solutions [34]. An important problem is the accumulation of dissolved solids, which eliminates the treated discharge outflows in water currents impossible.

Ultra Filtration Ultra filtration allows the removal of macromolecules and particles, but the elimination of pollutants such as dyes is never complete [35]. It is only between 31% and 76%. Even in the best cases, the quality of treated wastewater does not allow reuse of sensitive processes such as textile dyeing. Ultra filtration cannot be used as a pretreatment for reverse osmosis [36] or in combination with a biological reactor [37]. Microfiltration Microfiltration is suitable for the treatment of coloring pigments containing pigment dyes [38] as well as subsequent rinsing baths. Chemicals used in the dyeing bath, which is not filtered by microfiltration, will remain in the bath. Microfiltration can also be used as a pre-treatment of nanofiltration or reverse osmosis [39]. Adsorption It is the processes in which ions or molecules present in one phase are inclined to accumulate and concentrate on the surface of another phase. Physical adsorption occurs when weak intermediate bonds exist between the adsorbate and the adsorbent. Examples of such bindings are Van der Walls interactions, hydrogen bonding and dipole-dipole interactions. In most cases, physical adsorption is easily reversible. Chemical adsorption takes place when strong interspecies bindings are present between the adsorbate and the adsorbent by an electron exchange. Examples of such bonds are covalent and ionic bonds. Chemisorption is considered irreversible in most cases. Suzuki [40] discussed the role of adsorption in water environment processes and also evaluated the development of newer adsorbents to modernize the treatment and systems. Most adsorbents are very porous materials. Since the pores are generally very small, the internal surface is in the order size greater than the external area. Among the numerous techniques of dye removal, this technique gives the most important results because it can be used to remove others types of dyes also [40]. Waste water adsorption techniques have become more popular in recent years due to their efficiency in the removal of pollutants that are not readily biodegradable. Adsorption can produce high quality of water, while also a process that is cost-effective. Decoloriztion is a consequence of two mechanisms - adsorption and ion exchange and is influenced by many factors including dye / adsorbent interaction, adsorbent surface, particle size, temperature, pH and contact time. If the adsorption system is properly designed, it will produce a high quality treated waste water.



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Electrolysis Electrochemical technology is a simple process to remove wide range of dyes and pigments. This technique is based on the use of electric power in wastewater using sacrificial iron electrode to produce iron hydroxide in solution which leads to remove soluble and insoluble acid dyes from wastewater. In addition, Fe(II) may also reduce azo dyes into aryl amines. The electrochemical system can eliminate 90% of impurities. However, the process is expensive due to high energy requirements and uncontrolled radical reactions [41]. Activated Carbon Activated carbon is a preferred adsorbent, which is widely used for wastewater treatment, and contains a variety of toner. However, the disadvantage associated with this is its high cost [42, 43, 44]. Regeneration of saturated carbon is also expensive, not simple, and leads to loss of adsorbent. Use of coal at relatively low prices as starting materials is also unfounded for most of the anti-pollution applications [45]. Various carbonaceous materials such as coal, lignite, coconuts, wood and turf are used for commercially activated carbon [46]. However, the abundance and availability of agricultural by-products makes them good activated carbon raw materials. The agricultural products are renewable raw materials for the production of activated carbon since methods for reuse of waste materials are highly desirable [47]. The destruction of agricultural by-products is currently a major economic and ecological issue, conversion of by-products to adsorbents, such as activated carbon, represents a possible outlet. Fenton The oxidation system is based on fenton’s reagent (mixture of hydrogen peroxide and iron salt) that has been used for the treatment of both organic matter and inorganic substances. The process is based on the formation of reactive oxidation species that are capable of effectively disintegrating contaminants on wastewaters [48]. It is accepted that both hydroxyl and ferrylic complexes occur in fenton reaction depends on the mechanism and conditions of use of which one is predominant [49]. Oxidation system can be used effectively for the destruction of non-biodegradable toxic waste water [50]. Fenton's oxidation process can stain a large range of dyes and relative to ozonization. The process is relatively cheap and leads, in general, in a higher chemical oxygen demand reduction [51]. Fenton oxidation is limited to that in the textile process; where waste water is generally high at pH, while the Fenton process requires low pH. At higher pH, deposition produces large amounts of waste liquor iron salts and process loss its effectiveness. Ozonation Ozone is a very powerful and fast oxidizing agent that can react with most of the chemicals and with simple oxidizing agent ions, such as S-2, forming oxyanions such as


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SO3-2 and SO42 [52]. Ozone quickly decolorizes water-soluble dyestuffs, but with nonsoluble dyestuffs react much slower. In addition, the wastewater generated by the treatment of textiles generally includes other refractory components that react with ozone, which increases its demand [53, 54, 55]. Ozone degradation requires a high pH (pH> 10). In alkaline solutions, ozone reacts almost inseparably to all the present compounds [56], transforming organic compounds smaller and more biodegradable molecules [57]. Thus, ozone therapy seems logical to use biological methods to achieve complete mineralization [58]. An important limitation of ozonation is the high cost of ozone production process with it short half-life [59].

Advance Oxidation Process An advanced oxidation process can be used to remove dyes from the waste water to produce a highly reactive radical which may be reacts a wide variety of compounds that are difficult to break down. This process includes chlorination, bleaching and photocatalytic oxidation [60].

BIOLOGICAL METHODS Polymers and colorants are generally difficult to biodegrade and many substances are totally unsuitable for conventional biological treatments. For textiles dyes, in particular, more emphasis is given on biological treatment systems compared to physical and chemical methods. Biodegradation methods are commonly applied in the treatment of industrial effluents due to many microorganisms such as bacteria, yeasts, algae and fungi are able to accumulate and degrading different pollutants [61], and all biological systems require continuous effluents. Biological treatment requires a large land area and is limited by sensitivity to the diurnal variation and the toxicity of some chemicals, and less flexibility in design and operation. Biological treatment is unable to obtain satisfactory color removal with current conventional biodegradation processes.

Decolourization by Prokaryotes

The ability of bacteria to metabolize azo dyes has been investigated by a number of research groups [62]. Under aerobic conditions, azo dyes are not readily metabolized, though the ability of bacteria with specialized reductive enzymes to aerobically degrade certain azo compounds [63]. In contrast, many bacteria are anaerobic which reduces azo dyes by non-specific, soluble, cytoplasmic reductase activity. Anaerobic reduction of degraded azo dyes may convert into aromatic amines [64], which may be toxic, mutagenic, and possibly carcinogenic for mammals [65]. Therefore, in order to achieve



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complete degradation of azo dyes or azo compounds, it requires aerobic biodegradation of the produced materials [66]. In phthalocyanine colors, reversible reduction and discoloration are present under anaerobic conditions [67].

Decolorization by Fungi

The most studied fungi with respect to the degradation of dyes are ligninolytic fungi that produces enzymes such as lignin peroxidase, manganese peroxidase and laccase [68]. There are much literatures on the potential use of these fungi to oxidize phenolic, nonphenolic, soluble and insoluble dyes [69]. Species of Pleurotus ostreatus, Schizophyllum, Sclerotium rolfsii and Neurospora crassa appeared to increase up to 25% the degree of discoloration of particular textile dyes like triaryl methane, anthraquinone and indigoide using enzymatic preparations [70]. In contrast, manganese peroxidase has been reported as the main enzyme involved in color discoloration by Phanerochaete Chrysosporium [71] and lignin peroxidase for Bjerkandera adusta [72]. Some non-white-rot mushrooms that can successfully discolor dyes also reported by the researchers [73].

Decolorization by Yeast

In the literature, the ability to degrade azo dyes by yeasts has been described only in some respects. The first two reports describe the use of Ascomycete and Candida zeylanoides yeast isolated from contaminated soils to reduce model azo dyes [74]. Characterization of enzymatic activity is described in other studies with yeast Issatchenkia occidentalis [75], and the enzymatic system involved in a work with Saccharomyces cerevisiae [76].

Decolorization by Algae

Only few algae are reported for the degradation of dyes such as Chlorella, Oscillateria [77] and Spirogyra [78] (Figure 2). Jinqi and Houtian [77] also indicate that certain azo compounds tested could be used as sources of carbon and nitrogen. This could mean that algae can play an important role in the elimination of azo dyes and aromatic amines in stabilization ponds [78].


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Figure 2. Chlorella vulgaris and Spirogyra sp.

CONCLUSION This opinion makes a simple comparison between different physicochemical methods namely photocatalysis, electrochemical adsorption, hydrolysis and biological methods; also discussed the advantages and disadvantages of these methods involved in azo dye decolorization. The main disadvantages of physical methods such as adsorption, ion exchange and membrane filtration were simply transferred dye molecule to another phase rather than destroy them and are only effective when the volume of effluents are small. The main disadvantage of chemical methods such as chemical oxidation, electrochemical degradation and ozonation were the requirements of an effective pretreatment sludge production. Biological degradation, bioaccumulation and biosorption are the three most important biological technologies used in the dye removal process. These techniques have a good potential to replace conventional methods used for the removal of dyes from the industrial wastewater. Biological processes can be applied in situ at the infected site, these were usually environmental friendly, that is, no secondary pollution and they were cost effective. These were the principle benefits of biological technologies for treatment of industrial waste. Therefore, in the recent years, research is strongly focused on biological methods for the treatment of waste water. The downside of the degradation process is its low degradation efficiency for some dyes and practical difficulties in continuous method. The major disadvantage of bioaccumulation process is using living organisms that is not advisable for the continuous treatment of highly toxic wastewater. This problem can be overcome with biosorption by using dead biomass. In this context, the literatures suggested that biosorption is an inoffensive and cost-effective method; and also does not produce any secondary pollutants.



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[71] Aguilar MI, Saez J, Llorens M, Soler A, and Ortuno JF (2002). Nutrient removal and sludge production in the coagulation-flocculation process. Water Res 36: 2910-9. [72] Aguilar MI, Saez J, Llorens M, Soler A, Ortuno JF, Meseguer V, and Fuentes A (2005). Improvement of coagulation-flocculation process using anionic polyacrylamide as coagulant aid. Chemosphere 58: 47-56. [73] Aksu Z (2005). Application of biosorption for the removal of organic pollutants: a review. Process Biochemistry 40: 997-1026. [74] Allegre C, Maisseu M, Charbit F, and Moulin P (2004). Coagulation-flocculationdecantation of dye house effluents: concentrated effluents. J Hazard Mater 116: 57-64. [75] Ambrósio ST, and Campos-Takaki GM (2004). Decolorization of reactive azo dyes by Cunninghamella elegans UCP 542 under co-metabolic conditions. Bioresour Technol 91: 69-75. [76] Angelini LG, Bertoli A, Rolandelli S, and Pistelli L (2003). Agronomic potential of Reseda luteola L. as new crop for natural dyes in textile production. Industrial Crops and Products 17: 199-207. [77] Jinqi L, and Houtian L (1992). Degradation of azo dyes by algae. Environ Pollut 75: 273-8. [78] Kamel MM, El-Shishtawy RM, Yussef BM, and Mashaly H (2005). Ultrasonic assisted dyeing III. Dyeing of wool with lac as natural dye. Dyes and Pigments 65: 103-110.




Chapter 8

PLANT GROWTH PROMOTING BACTERIA IN HEAVY METALS BIOREMEDIATION Nezha Tahri Joutey, Nabil Tirry, Wifak Bahafid, Hanane Sayel and Naïma El Ghachtouli* Microbial Biotechnology Laboratory, Faculty of Sciences and Techniques, Sidi Mohamed Ben Abdellah University, Route Immouzer, Morocco

ABSTRACT Pollution of soils by heavy metals is a considerable environmental concern. It has a harmful impact on human health due to its accumulation in the food chain. In fact, microorganisms and plants employ different mechanisms for the bioremediation of polluted soils. Microbial populations, particularly Plant Growth Promoting Bacteria (PGPB), are known to affect heavy metal mobility and availability to the plant through their plant growth promoting traits, release of metal chelating agents, antioxidative enzyme activities, metal mobilization/immobilization and metal transformation. Therefore, they have the potential to enhance phytoremediation processes. An approach that combines both microorganisms and plants should ensure an efficient method to remediate metal polluted soils. However, the species of organisms involved in the process are the key factors in this approach’s success. This chapter highlights the different mechanisms of plant growth promotion and of metal remediation by metal-detoxifying PGPB as well as the recent progress in the exploitation of these bacteria (either natural or genetically engineered) alone or in combination with mycorrhizal fungi to support phytoremediation in the prospect of environmental restoration.

Keywords: heavy metals, bacteria, bioremediation, plant growth promotion mechanisms, mycorrhizal fungi *



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ABBREVIATIONS ACC AMF HCN IAA ISR PGP PGPB Al As Cu Cd Mn Ni Zn U Pb Cr(VI) Cr(III)

1-AminoCyclopropane-1-Carboxylate Arbuscular Mycorrhizal Fungi Hydrogen Cyanide Indole-3-Acetic Acid Induced Systemic Resistance Plant Growth Promotion Plant Growth Promoting Bacteria Aluminum Arsenic Copper Cadmium Manganese Nickel Zinc Uranium lead hexavalent chromium trivalent chromium

INTRODUCTION Plant Growth Promoting Bacteria (PGPB) are free living or symbiotic bacteria colonizing various habitats in, on or near parts of plants that affect their growth by providing some benefit to plants using one or both direct and indirect mechanisms. Direct promotion of plant growth by bacteria may be by providing substances that are usually in short supply. These bacteria have the ability to solubilize phosphorus and iron, to atmospheric nitrogen and to produce plant hormones, such as cytokinins, gibberelins, auxins, and ethylene. Furthermore, they improve plant tolerance to stresse, such as high salinity, drought, metal toxicity and pesticide load (Bashan and De-Bashan 2005). Indirect mechanisms may occur by preventing the deleterious effects of phytopathogenic microorganisms. They produce substances that harm or inhibit other microbes, but not plants. They alter the metabolism of the host plant, to increase its resistance to pathogen infection or they limit the availability of iron to pathogens (Bashan and De-Bashan 2005). Recently, in hopes of using PGPB on a large scale in sustainable agricultural practice and environmental cleanup strategies, attention has been paid to the understanding of many of the fundamental biochemical and genetic mechanisms that are operative in plant– bacterial interactions (Ahemad 2014, 2015). PGPB will begin to replace the excessive use


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of chemicals in agriculture as biofertilizers. Furthermore, the addition of PGPB to plants that are used in phytoremediation protocols typically makes the entire remediation process much more efficacious. During the last two decades, bioremediation has emerged as a potential tool to remediate the metal contaminated environment (Ma et al. 2011). Both plants and soil microorganisms have some constraints in their individual abilities to remove pollutants. An approach that combines microorganisms and plants (mainly PGPB) may promote a synergistic action providing a powerful strategy for strengthening the cleanup of contaminated environments (Adediran et al. 2015; Ontañon et al. 2014) (Figure 1). The use of transgenic plants and of bacteria that are genetically modified to inoculate phytoremediating plants are active areas of bioremediation research (Zhang et al. 2013). Unfortunately, the controversies surrounding genetically modified organisms and the fact that heavy metals remain in the soil are major limitations to this approach of bioremediation (Singh et al. 2011). Therefore, in this chapterwe have taken into consideration descriptions made by other authors to give an in-depth definition of PGPB in the hope of using it appropriately by researchers. Furthermore, it gives recent advances in plant growth promotion mechanisms. The chapter will focus afterwards on describing how the beneficial partnerships between plants and PGPB (either natural or genetically engineered) can be exploited as a strategy to accelerate metal bioremediation through their PGP mechanisms and their bioremediation abilities. Also, it will raise a crucial point about approaches combining PGPB and Arbuscular Mycorrhizal Fungi (AMF) together in the aim of increasing their bioremediation abilities.

PLANT GROWTH PROMOTING BACTERIA (PGPB): DEFINITION AND CLASSIFICATION Plant growth-promoting rhizobacteria (PGPR) was coined for specific strains of rhizosphere bacteria which stimulate plant growth (Kloepper and Schroth 1978). Also, rhizobacteria have been reported to describe rhizosphere bacteria that exhibit root colonization (Schroth and Hancock 1982). In 1991, the definition of PGPR was revised and theywere reported as a diverse subgroup of rhizosphere-colonizing bacteria, with a larger ecological niche encompassing root colonization including area in close proximity of the root (Zablotowicz et al. 1991). So, to be more precise, the term PGPR should only refer to plant growth promoting root-bacteria (that may exist in the internal area of root, outside of root surface and in close proximity of the root). Since some bacteria may have a beneficial effect on the plant even though they are outside the rhizosphere environment, Bashan and Holguin (1998) proposed a revision of the original definition of the term PGPR. Hence, they suggested that the old term PGPR, should be abandoned and that


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“rhizobacteria” will be replaced by “bacteria,” creating the modified term plant growth promoting bacteria (PGPB). Thus, PGPB were defined by the same author as free-living soil, rhizosphere, rhizoplane, and phyllosphere bacteria that, under some conditions, are beneficial for plants. Endophytic PGPB that live inside the plant have also been reported (Bashan and De-Bashan 2005). Nevertheless, a literature review of published papers in this field of scientific research has shown that researchers still confuse between PGPR and PGPB and often use the term PGPR inappropriately. Several researchers tried to classify PGPR or PGPB. Bashan and Holguin (1998) proposed two classes: biocontrol-PGPB and PGPB. Somers et al. (2004) classified PGPR, based on their activities, as biofertilizers (increasing the availability of nutrients to plant), phytostimulators (plant growth-promoting, usually by the production of phytohormones), biopesticides (controlling diseases, mainly by the production of antibiotics and antifungal metabolites) and rhizoremediators (degrading organic pollutants). While, PGPR were separated, by Gray and Smith (2005), into extracellular PGPR (ePGPR), occuring in the rhizosphere, on the rhizoplane or in the spaces between cells of the root cortex, and intracellular PGPR (iPGPR), which exist inside root cells, mainly in nodular structures. However, the classification into the two classes proposed by Bashan and Holguin (1998), and those proposed by Gray and smith (2005) does not seem to have been largely accepted. Since, in most studied cases, a single strain will often present multiple mechanisms of action, including biological control, that fulfill various functions. For example, Bacillus sp. BPR7 solubilized various sources of phosphates, potassium and zinc and produced siderophores, IAA, organic acid, phytase, lytic enzymes, oxalate oxidase, ACC deaminase and cyanogens. It also showed negative effect on the growth of several phytopathogens such as Fusarium solani, F. oxysporum, Macrophomina phaseolina, Rhizoctonia solani, Sclerotinia sclerotiorum and Colletotricum sp. (Kumar et al. 2012). In fact, in accordance with the types of plant-bacteria interactions, PGPB can be divided into two types: free living and symbiotic. Based on the bacterial localization, PGPB may be divided into several types: rhizospheric (living near, on, or even inside roots), endophytic (that can colonize some or a portion of the interior tissues of plants), phyllosphere (aerial parts of plants dominated by the leaves) and cyanobacteria able to facilitate the growth of aquatic plants, which may not bind to plants surfaces. Based on PGP mechanisms, they can be divided into three types, bacteria with indirect-PGP mechanisms, bacteria with direct-PGP mechanisms, and bacteria that use both indirect and direct mechanisms of PGP (Sivasakthi et al. 2014). Then, it is so difficult to reunite all these traits in one classification. Finally, to gather diverging ideas concerning PGPB and their classification, the more general term plant-growth promoting bacteria (PGPB), used in this manuscript, has been utilized to describe free living or symbiotic bacteria colonizing various habitats in, on or near part of plants that affect their growth by



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providing some benefit to plants using one or both direct and indirect mechanisms. Thus, in what follows we used only the term PGPB.

MECHANISMS OF PLANT GROWTH PROMOTION Promotion of plant growth by bacteria may be direct, by providing substances that are usually in short supply (atmospheric nitrogen fixation, phosphorus and iron solubilization, and plant hormones production), and by improving plant's tolerance to stresses (Richardson et al. 2009). Indirect PGP occurs by protecting the plants from the deleterious effects of phytopathogenic microorganisms (Bashan and de-Bashan 2005).

Direct PGP Mechanisms Different bacteria enhance plant growth directly by supplying nutrients or through synthesis and transport of plant growth hormones that have direct effects on the regulation of plant physiological processes.

Supply Nutrients In the absence of potentially pathogenic microorganisms, some bacteria have the ability to enhance plant growth by solubilizing poorly soluble nutrients with either bacteria siderophores or their capacity to reduce pH by the excretion of organic acids. Iron is an essential nutrient, but it is scarce in soil. PGPB, can produce siderophores which are high-affinity iron-chelating compounds that scavenge iron from mineral phases by formation of soluble Fe3+ complexes that can be taken up by root cells by active transport mechanisms (Bouizgarne 2013). Phosphorus (P) is another essential plant nutrient with low availability in many agricultural soils because it’s precipitate by reacting with iron, aluminum, and calcium. Some PGPB can play a key role in phosphorus nutrition by increasing its availability to plants by solubilization and mineralization by solubilizing phosphate (Martinez-Viveros et al. 2010). Phytohormones Production Phytohormones are plant growth regulators which have stimulatory effects on plant growth (Megala and Elango 2013). The commonly recognized classes of phytohormones, regarded as the “classical five,” are: abscisic acid, auxins, cytokinins, gibberellins, and ethylene (Baca and Elmerich 2007). Besides, polyamines that have been associated to many physiological processes and biotic and abiotic plant stress reactions (Eyidogan


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et al. 2012), may also play an important role in the bacterial tolerance to heavy metals, in particular Cr(VI) (Tahri Joutey et al. 2014).

Indirect-PGP Mechanisms The indirect promotion of plant growth by PGPB are carried through lessening or preventing the deleterious effects of phytopathogenic organisms, by producing antagonistic substances, by inducing systemic resistance to pathogens (Beneduzi et al. 2012), or also by competition for a substrate or an ecological niche (Ahmad et al. 2011). Suppression of Plant Pathogens The mechanisms underlying bacterial antagonisms for plant pathogens, involve the production of a wide variety of compounds with antimicrobial activity (antibiotics, lytic agents such as lysozymes, exotoxins and bacteriocins,…) (Beneduzi et al. 2012). Besides, inorganic volatiles compounds (hydrogen cyanide, ammonia) and some volatile organic compounds (long-chain ketones, sulphuric compounds), when applied in biologically relevant concentrations, may inhibit the growth of phytopathogenic fungi (Hunziker et al. 2015). Moreover, metal chelating agents named siderophores have received much attention in recent years. Because, besides their function as biocontrol agents by limiting the iron nutrition of pathogens (Sayyed et al. 2013), they have potential values in environmental research applications (Ahmed and Holmström 2014) as it will be discussed later. Induction of Systemic Resistance The induced systemic resistance (ISR) is an indirect mechanism that implicates the activation of plant defense state in response to infection by a pathogenic agent (Glick 2012). ISR involves a series of defense reactions that propagates from the site of induction to distant parts of the plant. It include signal transduction mediated by phytohormones, synthesis of phytoalexins, oxidative stress protection, production of enzymes related to plant defense, and formation of structural barriers such as callose deposition, accumulation of phenolics and wall thickening (Figueredo et al. 2014). Competition Some PGPB may compete phytopathogens for nutrients and suitable niches on the root surface. This competition may limit plant disease severity and incidence. In some cases, PGPB may rapidly colonize root surfaces and use most of the available nutrients, thus reducing the growth of pathogens (Glick 2015). For example, iron is important for plant health and metabolism. The scarcity of bioavailable iron in soil and on plant surfaces is responsible of competition. Under iron-limiting conditions, PGPB produce siderophores to competitively acquire ferric ion. The production of siderophores, in



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sufficient quantities, limited Fe3+ availability to the pathogen and lead to the induction of host resistance against the pathogen (Saraf et al. 2014).

PGPB ASSISTED PHYTOREMEDIATION OF HEAVY METALS Among all pollutants, heavy metals are of concern to human health due to their mutagenicity, cytotoxicity, and carcinogenicity. Furthermore, heavy metals cannot be degraded during bioremediation but can only be transformed from one organic complex or oxidation state to another. Due to a change in their oxidation state, heavy metals can be transformed to become either less toxic, easily volatilized, more water soluble (and thus can be removed through leaching), less water soluble (which allows them to precipitate and become easily removed from the environment) or less bioavailable (Chibuike and Obiora 2014). Bioremediation of heavy metals can be achieved via the use of microorganisms, plants, or the combination of both organisms.

Metal Detoxification by PGPB Although PGPB have been used widely as growth promoting agents in agronomic practices, special attention is being given to the exploitation of their metal detoxifying potential in the remediation of metal contaminated soils (Ahemad 2014). The detoxifying potential of heavy metals by PGPB was associated with many characteristics including PGP traits, induction of antioxidative enzyme activities, metal mobilization/ immobilization and metal reduction.

PGP Traits for Metal Detoxification Siderophores: The production of siderophores by some PGPB may have facilitated the extraction of metals from the soil; this is because heavy metals have been reported to simulate the production of siderophores and this consequently affects their bioavailability (Chibuike and Obiora, 2014). Hence, heavy metals influence the activities of siderophore-producing bacteria which in turn increases mobility and extraction of these metals in soil. Otherwise, Dimkpa et al. (2009) found that siderophores secreted by PGPB strains decreased the formation of free radicals, so that it allows to protect microbial auxins from degradation and enable them to enhance plant growth, which permit the augmentation of metal uptake. Siderophores are active in solubilizing and increasing the mobility of a wide range of metals. This ability of siderophores mainly depends on the stability constants of metal–siderophore complex which depends on ligand functionalities of siderophores that may have an affinity for a particular metal other than Fe (Ahmed and Holmström 2014).


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1-aminocyclopropane-1-carboxylate (ACC) deaminase: Some PGPB strains contain ACC deaminase, which can cleave ACC, the precursor ethylene in the plant. This action decrease the level of ethylene in stressed plants grown in metal contaminated soil, and lead to the prevention of root growth inhibition by ethylene (Barnawal et al. 2012). Thus, the heavy metal-mobilizing and metal–resistant strain Phyllobacterium myrsinacearum has the capacity to use ACC as the sole source of nitrogen by producing the enzyme ACC deaminase (Ma et al. 2013). Similarly, Achromobacter xylosoxidans, a Cu-resistant strain, used ACC as the sole nitrogen source and enhanced Cu accumulation in Indian mustard (Ma et al. 2009a). Phytohormones: The phytohormone production by PGPB is shown to play a key role in plant–bacterial interactions and plant growth contaminated soils by heavy metals (Kuklinsky–Sobral et al. 2004). Thus, phytostimulation efficiency and therefore the phytoremediation potential may be influenced by phytohormone production by bacteria. The observed plant growth promotion under Pb stress after inoculation of plant with P. fluorescens is thought to be the consequence of bacterial IAA production and excretion (Sheng et al. 2008). Madhaiyan et al. (2007) reported the greater potential of the endophytic bacteria, Burkholderia sp. and Methylobacterium oryzae to enhance growth of Lycopersicon esculentum under Ni and Cd stress. This effect was attributed to the ability of the bacteria to lower the level of stress ethylene induced by Ni and Cd or to the possible contribution of the endophytes in reducing the phytotoxic effects of the metals through their capability of biosorption and bioaccumulation.

Metal Chelating Agents Besides siderophores, organic chelators as biosurfactants, organic acid anions, and metallophores are known as metal-binding compounds. Organic acids may influence metal solubility, mobility and bioavailability in the soil by binding metal ions. Panhwar et al. (2015) showed that organic acids produced by a mixture of PGPB composed of Burkholderia thailandensis, B. seminalis, Stenotrophomonas maltophila, and Bacillus sp. were able to reduce Al toxicity through its chelation. Biosurfactants are amphiphilic compounds that reduce surface and interfacial tensions. They may be produced either on microbial cell surfaces or excreted extracellularly. Heavy metal ions can be captured by anionic biosurfactants through electrostatic or complexation techniques. Many yeast and bacterial species are able to produce exocellular polymeric surfactants in the form of polysaccharides, proteins, lipopolysaccharides, lipoproteins or complex mixtures (Mosa et al. 2016). The microorganisms are also able to release another group of compounds that contribute to regulate the intracellular metal concentrations, metallophores (e.g., protochelin and azotochelin produced by Azotobacter vinelandii) (Deicke et al. 2013). These compounds allow to avoid metal toxicity and maintain appropriate concentrations for microorganisms growth.



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Antioxidative Enzyme Activities Antioxidant compounds have an important effect on plants protection against oxidative damage induced by heavy metals (Adrees et al. 2015). The effect of Reactive Oxygen Species (ROS) can be prevented by enzymatic and non-enzymatic antioxidants (e.g., carotenoids, vitamin C and E and flavonoids). These antioxidants may chelate the metal ions responsible for generating ROS. They may also inactivate ROS or disturb the radical chain reaction and oxidative processes (Flora 2009). Increased biomass production under Zn stress due to P. aeruginosa inoculation was associated with superoxide dismutase (SOD), antioxidative enzyme activities, peroxydase (POD) and catalase (CAT), and non-enzymatic components such as ascorbic acid and total phenolics (Islam et al. 2014). Similarly, increased CAT and SOD activities and reduced malondialdehyde content by Pseudomonas koreensis were reported in Miscanthus sinensis grown in heavy metal–contaminated soil (Babu et al. 2015). It was also suggested that the alleviation of Cu stress of Vicia faba under hydroponic conditions due to inoculation with Rhizobium and PGPR Enterobacter clocae and Pseudomonas sp. was related to the antioxidant activities of SOD and CAT (Fatnassi et al. 2015). Metal Mobilization/Immobilization The phytoremediation of heavy metals might be limited by their low bioavailability due to their insolubility and soil-bound properties (Dal Corso et al. 2013). Indeed, successful phytoremediation depends on the bioavailability of heavy metals in soils and also on the interaction of plant roots with bacteria associated to root plants, which contribute to mobilize metal ions, increasing their bioavailable fraction. (Li et al. 2010). Therefore, it should be possible to improve the potential of metal extraction by hyperaccumulator plants through plant inoculation with selected metal resistant PGPB. Thus, the water-soluble Pb in soil was significantly increased by the bacterial strains P. fluorescens G10 and Microbacterium sp. G16. The availability of Pb to Brassica napus was enhanced (Sheng et al. 2008). The insoluble Pb and Cd fractions were gradually solubilized during the growth of Burkholderia sp. strain J62 (Jiang et al. 2008). Various sources of Zn such as ZnO, ZnCO3, or Zn3(PO4)2 were dissolved by Gluconacetobacter diazotrophicus through its production of 5-ketogluconic acid and Zn was made available for plant uptake (Saravanan et al. 2007). Mobilization of chromium from the soil, and enhanced availability to the plant by Pseudomonas sp. VRK3 was also reported (Hemambika et al. 2013). However, in some cases it has been reported that the inoculation with metal resistant bacteria decreased the uptake of metals by the plants and thereby increased plant biomass. For example, Madhaiyan et al. (2007) reported that availability of Ni and Cd in soil and their uptake in roots and shoots of tomato were reduced by Burkholderia sp inoculation. These effects were attributed to metals immobilization in the rhizosphere.


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Metal Transformation Due to a change in their oxidation state, heavy metals can be transformed to become either less toxic, easily volatilized, more water soluble (and thus can be removed through leaching), less water soluble (which allows them to precipitate and become easily removed from the environment) or less bioavailable (Tahri Joutey et al. 2013; Chibuike and Obiora 2014). Studies on metals biotransformation by PGPB are very scant. For example, Sayel et al. (2014) reported that P. fluorescens PF28 that promote plant growth, was able to reduce Cr(VI) to the less toxic Cr(III), showing the potential utilization of this strain for plant growth improvement as well as for Cr(VI) bioremediation. Similarly, Mathew et al. (2015) reported the enzymatic reduction of metallic mercury Hg2 + into less toxic volatile Hg°, with the mercury reductase MerA produced by the bacterium Photobacterium spp. MELD1. Recently, Xu et al. (2016) showed AsV reduction by Asresistant bacteria isolated from Pteris vittata and showing different PGP. Phytoremediation Phytoremediation is the use of plants to remediate polluted soils. It is considered as an eco-friendly and cost effective technology, especially for treatment of large contaminated areas with diffuse pollution. This technology is presently receiving high attention (Ma et al. 2011). There are various ways used by plants to interact with heavy metals including phytoextraction (extraction, transport and accumulation of metals in plant tissues), phytostabilization (stabilization of the metal in soil by reducing its bioavailability) and phytovolatilization (transformation of metals into volatile forms) (Pilon-Smits 2005). Recently, much more attention has been paid to the use of trees to clean up polluted soil or the dendroremediation. This technology present a great potential for metal phytoextraction, especially when using fast-growing tree species, for example, Salicaceae species, poplars (Populus sp. pl.) and willows (Salix sp. pl.) (Pajević et al. 2016). The plants selected to be used in the phytoremediation must have high metal absorption and accumulation (hyperaccumulators) aptitude and rapid growth rate (Mudgal et al. 2010). These criteria are important to decrease the treatment time and to avoid the disadvantages of phytoremediationthat requires a long-term economic commitment and patience by project managers. Otherwise, the adjunction of PGPB in phytoremediation strategy is of great importance in metal contaminated soils (Glick 2015). It is noteworthy that phytoextraction is the most widely used method for treatment of metals contaminated environment and this process may be greatly enhanced by PGPB (Glick 2010).

Enhancing Phytoremediation by PGPB PGPB can enhance phytoremediation of various pollutants, which offers many advantages: (i) the stimulation of plant growth due to the plant growth-promoting traits of



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PGPB, (ii) the phytostabilization and phytoextraction of heavy metals by the combined actions of microorganisms and plants, (iii) the effect on heavy metal mobility and phytoavailability, (iv) the enhancement of antioxidative enzyme activities, and v) the interaction with other microflora mainly AMF, which enhance phytoextraction/ phytostabilization of the plants (Figure 1).

Figure 1. Mechanisms used by plants and PGPB for cleanup of metals. Plants-microbes interactions enhance metals detoxification.

The growth and metal-extraction efficiency of plants exposed to toxic metals is widely documented. Recent advances in PGPB-bioaugmentation assisted phytoremediation showed that metal chelation by bacteria is the key mechanism of PGPB in bacterial-assisted phytoremediation by B. juncea inoculated with R. leguminosarum and P. brassicacearum (Adediran et al. 2015). Furthermore, Chiboub et al. (2016) selected a consortium of four isolates belonging to Pseudomonas sp. and the Rhizobium sullae that has the ability to produce PGP substances such as IAA and siderophore that enhance the growth of Sulla coronaria under Cd stress. The authors showed intracellular Cd accumulation and confirmed the existence of a cadmium-resistant gene by polymerase chain reaction. Whereas, Juncus acutus was able to rhizofiltrate Cr(VI) from contaminated water. Dimitroula et al. (2015) showed that isolated endophytic bacteria from J. acutus have a high potential to reduce Cr(VI) to Cr(III). Sobariu et al. (2016) reported that Azotobacter sp. bacteria could indeed stimulate the average germination efficiency of Lepidium sativum. The growth of L. sativum has been affected to a greater extent in Cd(II) solutions due its higher toxicity compared to that of Cr(VI). The reduced tolerance index (TI, %) indicated that plant growth in symbiosis with PGPB was however


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affected by heavy metal toxicity, while the tolerance of the plant to heavy metals was enhanced in the bacteria-plant system. Kamran et al. (2016) reported that the inoculation with P. putida increased plant growth and Ni uptake by Eruca sativa. These effects were attributed to IAA, siderophore and ACC deaminase. Some other examples are shown in Table 1.

PGPB FOR STRENGTHENING MYCORRHIZO-REMEDIATION Besides bacteria, mycorrhizal fungi can also be used in phytoremediation, this technology is named mycorrhizoremediation. Establishing symbiotic relationships with 80–90% land plants (Amora-Lazcano et al. 2010), these fungi are known to improve plant health and growth by improving mineral nutrition, in particular phosphorus and increasing tolerance or resistance to abiotic and biotic stresses (Ismail et al. 2013). The two most common types of mycorrhiza are ectomycorrhiza and arbuscular mycorrhiza (endomycorrhiza), differentiated for lack or occurrence of root cell invasion by the fungus, respectively. In both mycorrhiza types, heavy metals can be bound and sequestered by the main fungal cell wall components (Phieler et al. 2014). Indeed, ectomycorrhizal associations can display considerable resistance against toxicity in soil polluted with metals. The density and surface area of the mycelium and structure of the fungal sheath are likely to be important characteristics determining the efficiency of an ectomycorrhizal association to endure metal toxicity and to defend the host plant from pollutant contact (Colpaert et al. 2011). Besides, several studies revealed that arbuscular mycorrhizal fungi (AMF) enhance plant tolerance to metals and organic pollutants and promote phytoremediation process (Wu et al. 2016). Many concrete examples of how mycorrhizae can overcome some of the effects of heavy metal(loid) stress in plants are reviewed by Gamalero et al. (2009) and can be summarized as follow : heavy metals may be (i) attached to the fungal cell wall and subsequently accumulated in the vacuoles, (ii) sequestred by siderophores and deposited into the root apoplasm or into the soil, and possibly taken up by plant ferrisiderophore receptors, (iii) complexed by metallothioneins or phytochelatins synthesized by the fungus or the plant, (iv) extruded from the cytosol by specific transporters located on plant membranes, (v) sequestred by glomalin, a glycoprotein produced by AMF, (vi) stabilized by AMF in the soil, thus reducing their availability and decreasing the risk of toxicity to other soil microorganisms and plants growing in the immediate vicinity, (vii) altered metabolism of heavy metals by extensive changes in gene expression as well as protein synthesis induced by the AM symbiosis. Otherwise, in natural ecosystems, numerous bacterial taxa are closely associated with AMF where they colonize the surface of extraradical hyphae and spores on which they can form biofilm-like structures (Cruz and Ishii 2012). The metal resistant bacteria and AMF are often used as bioinoculants to enhance the establishment, growth and



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development of remediating plants in metal contaminated soils. They may act as: (i) biofertilizers, by facilitating mineral phytoavailability, (ii) phytostimulators, by modulating phytohormones balance, (iii) bioalleviators, by reducing ethylene synthesis, (iv) biopesticides, by preventing deleterious effects of phytopathogens via production of antifungals and ISR, and (v) biomodifiers, by modifying root biomass and morphology (Ma et al. 2016). In addition, Gamalero et al. (2009) reported one more mechanism involved in the bacteria-AMF-plant interactions, is that the bacterial strain may behave as a mycorrhiza helper bacterium and then stimulate fungal development. Thus, it is hypothesized that plants, bacteria and AMF form a tripartite association that result in a promoting plant growth consortium (Bonfante and Anca 2009). However, it is noteworthy that the literature regarding the use of a combination of bacteria and AMF to improve phytoremediation efficiency and plant growth in heavy metal polluted sites is relatively scant. Recently, Dhawi et al. (2016) showed changed levels of metabolomics compounds in plants treated by Pseudomonas sp. and the mycorrhizal mix and suggested that PGPB enhanced metabolic activities which resulted in increased element uptake and sorghum root biomass. An in situ experimental study conducted by Guarino and Sciarrillo (2017) through an integrated phytoremediation system (IPS), revealed that inoculation of rhizosphere microbes (bacteria and AM fungi) enabled to increase the biomass of Acacia saligna and Eucalyptus camaldulensis. Consequently, it promotes the phytoremediation of potential toxic elements which are kept in plant root system, with a restricted translocation to the aerial parts. In the rhizosphere level, the bioavailability of the contaminants was considerably reduced by the “plant-fungi-bacteria” system. Thus, it was concluded that this strategy allows a socially, economically, and environmentally sustainable costs for soil remediation in few years with. Other examples of plant-AMFbacteria interactions in heavy metals bioremediation are given in Table 2. Recent approaches combining PGPB and AMF together aims to increase their bioremediation abilities. Nevertheless, further investigation about interaction between fungi and bacteria may bring more detailed knowledge of molecular and biochemical mechanisms implicated, which will help to improve quality of co-inoculants in bioremediation strategies.

IMPORTANCE OF BIOENGINEERING IN BIOREMEDIATION The use of transgenic plants and inoculation of phytoremediating plants with bacteria that are genetically modified are emerging areas of phytoremediation research (Zhang et al. 2013). The recently coined term “genoremediation” is used in this context.


Table 1. PGPB assisted phytoremediation of metal-contaminated soils PGPB Bacillus sp. SN9

Heavy metals Ni

Plants Brassica juncea and B. oxyrrhina Sedum plumbizincicola

Effect of PGPB On metals Solubilize Ni in soil

Phyllobacterium myrsinacearum RC6b Bacillus PSB10

Cd, Zn and Pb Cr

Chickpea (Cicer arietinum L.)

Cr tolerance Cr detoxification

Phyllobacterium myrsinacearum strain RC6b Rhizobacterial strain RZB-03 Pseudomonas P35

Pb, Cd, Zn,

Sedum alfredii and Medicago sativa L.

Metal resistance

Cr(VI)

Scirpus lacustris L.

Chromate tolerance

Ni and Cd Cr(VI)

Withania somnifera.

Multiple heavy metals tolerance (As, Pb, Ni Cd) Chromate tolerance and reduction

Cr(VI)

Brassica napus

Brucella sp. K12.

Cr(VI)

Hibiscus esculentus L.

Psychrobacter sp. SRS8

Ni

Ricinus communis and Helianthus annuus

Pseudomonas aeruginosa strain OSG41 Pantoea sp. FC 1

Chickpea

Resistance and metal mobilization

Reduction of Cr(VI) to Cr(III) Decrease of Cr(VI) concentration in soil Ni accumulation

References On plants Increase phytoextraction: by enhancing Ni- accumulation and shoot and root biomass. Increase plant growth and organ metal concentrations, except Pb.

Ma et al. (2009b) Ma et al. (2013)

Provide protection to the plant by mitigating the toxic effects of Cr. Reduce Cr uptake by plant, increase in growth, yield and quality of plant. Increase the phytoextraction efficiency of Pb, Cd, and Zn by shoots. Increase shoot biomass yields of plants. Increase metal concentrations in shoots. Increase biomass, total chlorophyll and protein content. Increase Cr accumulation in plants. Enhance seed germination and stimulation of root and shoot growth. Increase biomass, nodule formation, grain yield and protein of plants.

Wani and Khan (2010)

Production of siderophores, IAA and solubilization of inorganic phosphates. Decrease of Cr(VI) concentration in plant.

Ontañon et al. (2014) Maqbool et al. (2015) Ma et al. (2011)

Plant growth.


Liu et al. (2015) Singh et al. (2010) Rathaur et al. (2012) Oves et al. (2013)

Table 2. Plant-fungi-bacteria synergic effect on heavy metals bioremediation PGPB Brevibacillus sp.

AM fungi Indigenous Cd-tolerant and Cd-sensitive G. mosseae Autochthonous Glomus mosseae

Heavy metals Cd

Plants Trifolium repens L.

Cd

Trifolium repens L.

Pb

Glycine max

Brevibacillus sp.

Glomus macrocarpum Tul. and Tul. Glomus mosseae

Cd

Trifolium repens L.

Brevibacillus sp.

Glomus Mosseae

Zn

Pseudomonas monteillii

Indigenous AMF

Cd

Microbacterium sp. SUCR140

G. intraradices G. mosseae, G. fasciculatum, G.aggregatum Glomus, Scutellospora, Acaulospora

Cr(VI)

Brevibacillus sp.

Bradyrhizobium sp.

Streptomyces, Azotobacter, Pseudomonas and Paenibacillus

Fe

Sorghum bicolor Zea mays

Pennisetm glaucum and S. bicolor

Effects on Plant/bioremediating effects Increase nutrient and root development. Decrease Cd availability and uptake by the plant. Enhance extent of root biomass and symbiotic structures. Enhance plant nutrient acquisition (N and P). Reduce Cd transfer from soil to plants. Improve uptake of nutrients (mainly P) and Pb. Increase P and K and decrease concentration of heavy metals in plant tissue. Maximize enzyme activities. Enhance IAA accumulation and soil enzymatic activities Increase the total biomass of sorghum plants. Increase AM colonization in soil. Reduce chromate toxicity and improve growth and yields of plants.

Reference Vivas et al. (2003a)

Increase the extent of iron absorption.

Mishra et al. (2015)


(Vivas, et al. 2003b)

Andrade et al. (2004) Vivas et al. (2005) Vivas et al. (2006) Duponnois et al. (2006) Soni et al. (2014)

Table 2. (Continued) PGPB Pseudomonas sp. TLC 6-6.5-4

Pseudomonas sp. TLC 6-6.5-4

Bacillus licheniformis, B. megateriu, B. polymyxa, B. subtilis, B. thuringiensis

AM fungi G. intraradices, G. mosseae, G. aggregatum, and G. etunicatum G. intraradices, G. mosseae, G. aggregatum, and G. etunicatum Pisolithus tinc-torius, E. colombiana, G. clarum, G. intraradices, G. etunicatum

Heavy metals Cu, Zn, Fe

Plants Zea mays

Effects on Plant/bioremediating effects Increase maize nutrient uptake and growth. Change in metabolic pathways.

Reference Dhawi et al. (2015)

Cu, Fe, Zn

Sorghum

Increase root biomass and uptake of most of the elements, changed levels of metabolomics in roots.

Dhawi et al. (2016)

As, Cd, Pb and Zn

Acacia saligna and E. camaldulensis

Retention of potential toxic elements in plant roots and restricted translocation to the aerial parts.

Guarino and Sciarrillo (2017)



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Transgenic Plants Several biotechnological approaches are involved in genoremediation, such as metal homeostasis genes, genes for abiotic and biotic stresses, metals chelators and transporter genes, biodegradative enzymatic genes, risk mitigating genes and metal uptake regulator genes (Mani and Kumar 2014). In the context of transgenic plants, to enhance metal uptake, three different engineering approaches can be conceived, which include enhancing the number of uptake sites, alteration of specificity of uptake system to reduce competition by unwanted cations and increasing intracellular binding sites (Mello-Farias et al. 2011). Hairy roots have shown undeniable potential regarding the use of plants for in vitro phytoremediation studies. For example, hairy roots obtained from transgenic Nicotiana tabacum were very efficient in copper phytoremediation, and hairy roots from transgenic A. thaliana expressing the Cu-binding periplasmic protein CopC were very efficient regarding Cu accumulation (Pérez-Palacios 2015). However, some limitations exist, such as the difficulty of hairy roots to comply with constant environmental fluctuations, metal loads, the presence of undesirable microorganisms and hydraulic conditions (Khandare and Govindwar 2015).

Genetically Modified PGPB The understanding of the metal-microorganisms interactions and their application for metal bioremediation experienced a significant evolution with the recent advances in the field of molecular biology. The use of genetically engineered bacteria for bioremediation of heavy metals is of special interest due to its selective nature of transforming the toxicants. Huang et al. (2004) recommended three criteria to select a suitable strain for gene recombination and inoculation into the rhizosphere, (i) the strain should be stable after cloning and the target gene should have a high expression, (ii) the strain should be tolerant or insensitive to the contaminant, and (iii) some strains can survive only in several specific plant rhizosphere. Many genetic engineering bacteria that can be adopted in microbe assisted remediation of heavy metal polluted soils are reported in the literature. For instance, Valls et al. (2000) reported that genetically engineered Ralstonia eutropha can be used to sequester metals (such as Cd) in polluted soils. This is made possible by the introduction of metallothionein (cysteine rich metal binding protein) from mouse on the cell surface of this organism. Although the sequestered metals remain in the soil, they are made less bioavailable and hence less harmful. Patel et al. (2010) showed that the insertion of Hexa-histidine peptide in a permissive site of the surface layer (Slayer) protein RsaA of Caulobacter crescentus enhances the binding capacity of Cd. The controversies surrounding genetically modified organisms and the fact that the heavy metal remains in the soil are major limitations to this approach of bioremediation. For


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more information regarding genetic engineered bacteria, authors are referred to Singh et al. (2011) minireview, which explores genetic engineered bacteria in the environmental bioremediation and also highlights the challenges and limitations associated with the use of genetically modified bacteria in field conditions.

CONCLUSION This chapter presents recent scientific advancement in usefully applying PGPB as bioremediation means for environmental management of heavy metals (assisted with plants and microorganisms) to diminish their toxicity around the world. PGPB possessing single or multiple traits such as alteration of metal availability, alleviation of metal toxicity, production of siderophores, phytohormones and biochelators, fixation of nitrogen, and solubilization of mineral nutrients have been extensively suggested as efficient bioinoculants for PGPB-assisted phytoremediation. Deciphering cooperation, integration and assimilation of such biotechnological advances along with traditional and ethical understanding are required for the suitable remediation of polluted soils.

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Chapter 9

EX SITU STIMULATED BIOREMEDIATION OF A SOIL CONTAMINATED WITH OIL POLLUTANTS: THE DYNAMICS AND THE EFFICIENCY OF BIODEGRADATION OF SATURATED AND AROMATIC HYDROCARBONS Tatjana Šolević Knudsen1,*, Mila Ilić1, Jelena Milić1, Gordana Gojgić-Cvijović1, Srđan Miletić1, Vladimir Beškoski2 and Miroslav M. Vrvić2 1

Institute of Chemistry, Technology and Metallurgy, University of Belgrade, Serbia 2 Faculty of Chemistry, University of Belgrade, Serbia

ABSTRACT The aim of this research was to investigate the overall dynamics and the efficiency of biodegradation of saturated and aromatic compounds – constituents of the oil pollutant, during ex situ bioremediation of the polluted soil. The effect of natural biodegradation was investigated in a parallel set of samples of the same soil which was not subjected to the processes of stimulation (bioaugmentation, biostimulation and aeration). During the period of six months, the biodegradation process caused a gradual and prominent degradation of phenanthrene, dibenzothiophene, fluoranthene/pyrene and their alkyl homologues. Among the polycyclic aromatic hydrocarbons (PAHs), distributions and abundances of benz(a)anthracene/chrysenes and triaromatic steroids were the least *



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Tatjana Šolević Knudsen, Mila Ilić, Jelena Milić et al. affected by biodegradation. n-Alkanes and isoprenoids remained low in abundances during and until the end of the experiment, indicating that the availability, rather than the biodegradability, was responsible for removal of these compounds. In the set of the samples exposed to the stimulated bioremediation, within the homologue series of phenanthrenes, dibenzothiophenes and fluoranthene/pyrenes, higher alkylated homologues were more degradable than the lower ones. Additionally, lower alkylated homologues were more degradable than their parent molecules. These results are opposite to those described in the literature and indicate that a proper choice of bioremediation conditions can influence biodegradation sequences of the compounds present in the pollutant. Furthermore, these results showed that the extent of biodegradation in the samples exposed to stimulation was much higher than in the samples where only natural biodegradation occurred, pointing that the stimulated biodegradation is much more efficient process than the natural biodegradation.

Keywords: ex situ bioremediation, soil, oil pollutant, polycyclic aromatic hydrocarbons

1. INTRODUCTION Contamination of soils with crude oil and its derivatives is nowadays one of the major environmental problems. In the myriad of available remediation technologies for treatment of soils polluted with oil pollutants, bioremediation is usually a method of choice because it is cost effective and environmentally friendly (Reis and Miertus 2008). Furthermore, it relies on the natural ability of soil microorganisms to degrade organic compounds, which can ultimately result in a complete mineralization of the pollutant (Sims et al. 1989). Biodegradation of crude oil in reservoir rocks is a well investigated process. It is generally accepted that this process proceeds according to the quasi-sequentional biodegradation scale after Peters and Moldowan (1993) which was amended later by Head et al. (2003). On the other hand, the sequence of compounds to be degraded during the biodegradation of oil pollutant in the environment might be different from the one during the biodegradation of crude oil in reservoir rocks. Discordance in the biodegradation sequences of these two processes is usually a consequence of a specific microbial community employed (Van Hamme et al. 2003) or specific conditions in the environment (Margesin and Schinner 2001). However, even if these differences are taken into account, it is a general rule that the order of the most biodegradable compounds in oil pollutants in the environment is the same as the order of the most biodegradable compounds in crude oils in their reservoir rocks. In the group of saturated hydrocarbons the commonly accepted sequence of removal of different compound classess during biodegradation is: n-alkanes > alkylcyclohexanes > acyclic isoprenoid alkanes > bicyclic alkanes, steranes, hopanes (Peters and Moldowan, 1993). In the group of aromatic hydrocarbons, biodegradation usually follows the sequence: BTEX (benzene, toluene, xylene, ethylbenzene) aromatics


Ex Situ Stimulated Bioremediation of a Soil Contaminated with Oil Pollutants 213 > naphthalenes > phenanthrenes, dibenzothiophenes > chrysenes > mono- and triaromatic steroids (Wenger et al. 2001). Among all of these compounds, in oil spill pollution studies the greatest concerns attract polycyclic aromatic hydrocarbons, due to their environmental persistence but also due to their toxicity and carcinogenicity. Their alkyl homologues are even more abundant in crude oils and in their derivatives. Furthermore, they persist in the environment longer than their parent compounds, and some of them are considered more toxic. However, many environmental studies are dealing with parent PAH molecules only, while the reports on their alkyl homologues are still scarce and usually limited to selected isomers (Fingas 2015). Our previous research in this field revealed that zymogenous microorganisms isolated from crude oil polluted soils had significant bioremediation potential and that they could be successfully used in biodegradation of both, saturated and aromatic hydrocarbons (Gojgic-Cvijovic et al. 2012; Ilic et al. 2011; Milic et al. 2015; Solevic et al. 2011). We also concluded that bioremediation of the environment polluted with crude oil and its derivatives can efficiently remove petroleum hydrocarbons from both, polluted soil (Avdalovic et al. 2016a; Beskoski et al. 2010; 2012; Jednak et al. 2017; Jovancicevic et al. 2008; Milic et al. 2009) and ground water (Avdalovic et al. 2016b; Beskoski et al. 2016; Maric et al. 2015). The aim of our present study was to explore the overall dynamics and the efficiency of biodegradation of saturated and aromatic compounds constituents of the oil pollutant during the ex situ bioremediation of the polluted soil. In the class of aromatic hydrocarbons both PAH molecules and their methylated homologues were thoroughly investigated. The results of the bioremediation experiment of the soil that was regularly aerated and treated with biomass of zymogenous microbial consortia isolated from the oil–contaminated soil (bioaugmentation) and nutritive substances (biostimulation), were compared with the results of biodegradation of the soil that was not exposed to these processes of stimulation.

2. EXPERIMENTAL STRATEGY, MATERIALS AND METHODS 2.1. The Oil Polluted Soil and Formation of the Pile The soil used in this research was excavated soil from surrounding areas of an energy power plant in New Belgrade (Serbia). Due to a break-down of the energy power plant facilities, this soil had been polluted for a year with fuel oil and sediment from a fuel oil tank, but also with some other crude oil derivatives. Because of that, this soil was considered that might contain the appropriate amount of oil pollutants as well as a bioremediation potential high enough to satisfy the goals set by the aim of this research.


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In this study, a biopile bioremediation cleanup technology was applied. This approach has been proven effective in reducing concentrations of nearly all of constituents of petroleum products typically found in oil polluted soils (US EPA 2016). The biopile for bioremediation was prepared on a waterproof asphalt surface. The surface area used was approximately 1500 m2 with a 1% slope gradient. The polluted soil (approximately 150 t; 210 m3) was uniformly distributed over 300 m3 of ungraded not rinsed sand from the Sava River (settlement Ostruznica, Serbia). The river sand was added as a bulking agent but also as a porosity increasing material. Approximately 60 m3 of the sawdust from poplar, beech, and oak, originating from wood processing industry, was added to the mixture of the oil-polluted soil and the river sand. The sawdust role was twofold: as an alternative additional source of carbon and as a material which should increase the water retention capacity. All these components were mixed several times with a front-end loader in order to ensure homogeneity. The entire homogenized material, defined as a substrate for bioremediation, was then formed into a biopile shape with bulldozers. The final volume of the biopile was approximately 600 m3 with dimensions of 75 × 20 × 0.4 m (length, width, height). The entire treatment area was enclosed with a perimeter drain. All leachates and runoffs were directed to a joint vessel, from which they were pumped back onto the biopile. Immediately after mixing, but before the addition of sawdust, approximately 10 m3 of the biopile material was set aside on the same waterproof asphalt surface, to be used as a control pile. During the experimental period, the control pile was left untreated and used to monitor the effect of natural biodegradation in it. The complete analytical procedure that was applied to the biopile samples was also applied to the control pile samples.

2.2. Production of the Zymogenous Microbial Consortium Used for Reinoculation and Treatment of the Biopile A consortium of microorganisms was isolated from the polluted soil, by enrichment in 200 mL volumes of mineral medium (10 vol.%) (Loser et al. 1998), having a heavy fuel oil (2 g/L) as the only energy and carbon source in Erlenmeyer flasks (1 L). Suspensions of the microbial consortium were used to seed four Erlenmeyer flasks (5 L). Each of them contained 2000 mL of the medium with 23 g of nutrient broth (Torlak, Belgrade, Serbia), 100 mL of soil extract (http://www.ccap.ac.uk/media/recipes/SE.htm) and 20 g of the heavy fuel oil. Commercial non-toxic and readily biodegradable surfactants, BioSolve CLEAR supplied by The Westford Chemical Corporation (Westford, MA, USA) were used as


Ex Situ Stimulated Bioremediation of a Soil Contaminated with Oil Pollutants 215 surface active agents to solubilize the oil pollutants. The original solution supplied by the manufacturer was used at a concentration of 1 mL/L. The growth conditions were as follows: temperature, 28ºC, constant agitation speed (120 rpm), pH 7.0 (adjusted with 1 M HCl or NaOH) and duration of growth 96 h. The microbial population from all four flasks was used to inoculate (approx. 1 vol.%) a bioreactor designed by our group (total volume 1000 L) with a working volume of 800 L, producing the microbial consortium. The medium used was (per L): 12 g meat peptone (Torlak, Belgrade, Serbia), 0.2 g (NH4)2HPO4, 25 g of autoclave-sterilized soil sampled from undisturbed deciduous woodland, 1 mL Bio-Solve CLEAR original solution, and 10 g the heavy fuel oil. The growth conditions were: non sterile, 25ºC, aeration and agitation 0.70 volume of air/volume of medium min-1, pH 7.0 (adjusted with 10 M HCl or NaOH), duration 48 h and sunflower oil (1 mL/L) used as an antifoam. During the experiment the biopile was continuously sprayed with biomass of microbial consortia isolated from the crude oil–contaminated soil (bio augmentation) and nutritive substances (bio stimulation). Each 2 weeks aeration and mixing were performed by turning and mixing the biopile. After each treatment the biopile was covered with polyethylene foil in order to prevent reduction of the temperature but also to prevent the influence of the weather conditions on the biodegradation efficiency.

Figure 1. Phases in preparation of the ex situ bioremediation biopile: A) Mixing of the polluted soil with sawdust and sand; B) Formation into a biopile shape; C) Spraying of biostumulation solution and watering; D) Final bioremediation biopile (adapted from Beskoski et al. 2012).


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2.3. Sampling Composite samples for analyses were taken from the biopile and the control pile by “zig-zag” sampling from 30 random places. An Eijkelkamp auger soil sampler with appropriate augers was used to sample the soil. The composite samples were sieved through a 1 mm grid, collected in stopped glass jars, stored at 4ºC and analyzed within 12 - 24 h after sampling (Paetz and Wilke, 2005). During the interval of 6 months, the samples were taken five times (07. September 2009; 06. October 2009; 09. November 2009; 12. January 2010. and 18. March 2010.). The samples taken from the BioPile were marked BP1–BP5. The samples taken from the Control Pile were marked CP1–CP5. During the six month experiment, daily temperature was in the range from -2.3 to 23.5°C, averaging at 7.6 ± 6.3°C. The temperature of the soil was stable during the whole period, above 25°C, due to the intensive microbiological activity.

2.4. Microbiological Analyses The number of microorganisms was determined by plating appropriate serial dilutions on agar plates incubated at 28°C. The medium used for determination of total chemoorganoheterotrophs was the nutrient agar (Torlak, Belgrade, Serbia). The medium used for determination of hydrocarbon degraders was the mineral medium (Loser et al. 1998) containing 2 g/L of standard D2 diesel fuel (Bossert et al. 2002). The medium used for determination of yeast and moulds was the malt agar (Torlak, Belgrade, Serbia). Analytical Profile Index (API) strip tests were used as a method for manual microorganism identification to the species level. Initially, in order to choose the adequate API kit, orientation tests were performed, such as identification using Gram’s stain, morphological features and other simple tests like oxidase test, catalase test, fermentation test, and growth in a CO2 atmosphere. According to the results obtained from the orientation tests, a suitable API test was chosen for microorganism identification (Table 1). These tests were inoculated with the bacterial suspension which reconstitutes media. During incubation, metabolism produces colour changes that are either spontaneous or revealed by the addition of reagents. The reactions are read according to the Reading Table and Identification is obtained by referring to the Analytical Profile Index or using the identification software, according to the manufacturer instructions (bioMerieux, Marcy l’Etoile, France).


Ex Situ Stimulated Bioremediation of a Soil Contaminated with Oil Pollutants 217 Table 1. Criteria for selection of suitable API tests Presumptive Organism ID Gram negative bacilli  Oxidase positive  Non-fermenters  Non-Enterobacteriaceae Gram negative bacilli  Oxidase negative  Fermenters  Enterobacteriaceae Gram positive bacilli  Catalase positive  Small bacilli of characteristic shape  Partial discolouring (granulation) Gram positive bacilli  Catalase positive  Short bacilli with rounded ends  Mobile at 20-25C  Immobile at 37C

Which API strip to use API 20 NE

API Rapid 20 E

API Coryne

API 50CH/B

For identification of gram-negative non-Enterobacteriaceae API 20NE tests were used. For identification of Enterobacteriaceae API Rapid 20E tests were used. For identification of Corynebacteria and coryne-like organisms API Coryne tests were used. For identification of Bacillus API Bacillus 50 CHB/E were used. All tests were performed according to the manufacturer instructions (bioMerieux, Marcy l’Etoile, France).

2.5. Analyses of Saturated and Aromatic Compounds For detailed analyses of saturated and aromatic compounds, organic substance from all soil samples was extracted in Soxhlet apparatus with dichloromethane (HPLC, J. T., USA). From these extracts, saturated and aromatic hydrocarbons were isolated by alkaline digestion followed by column chromatography: the extracts were saponified with a 5% solution of KOH in methanol and, after standing overnight, neutralized with 10% hydrochloric acid. The products were dissolved in a mixture of dichloromethane (containing 1% methanol) and hexane (1:40). This mixture was separated by column chromatography on alumina (MERCK 90 Å, 70-230 mesh; 5% deactivated) and silica gel (MERCK 60 Å, 70-230 mesh; 10% deactivated). The saturated hydrocarbons were eluted with hexane. The aromatic hydrocarbons were eluted with dichloromethane. Detailed


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description of the analytical procedure and separation of other compound classes (alcohols and fatty acid methyl esters) was discussed in previous papers (Jovancicevic et al. 2003; 2005). Saturated and aromatic hydrocarbons in these extracts were analyzed by gas chromatography - mass spectrometry (GC-MS) techniques. For these analyses an Agilent 7890N gas chromatograph with a HP5-MS capillary column (30 m length; 0.25 mm inner diameter; 0.25 μm film) was used. The following temperature program was employed: 80ºC for 0 min; then 2ºC min-1 to 300ºC and then held for 20 min. Helium was used as the carrier gas with the flow rate 1 cm3 min-1. The GC was coupled to a Hewlett-Packard 5972 MSD operated at 70 eV and scanning masses in the 45 - 550 range. Preliminary analyses of the investigated samples were conducted in the full-scan mode. Detailed analyses of the target compounds were conducted in the single-ion monitoring mode (SIM), comprising the following ions: m/z = 71 (n-alkanes), m/z = 183 (isoprenoids), m/z = 178 (phenanthrene), m/z = 192 (methylphenanthrenes), m/z = 206 (dimethylphenanthrenes), m/z = 220. (trimethylphenanthrenes), m/z = 184 (dibenzothiophene), m/z = 198 (methyldibenzothiophenes), m/z = 212 (dimethyldibenzothiophenes), m/z = 202 (fluoranthene/pyrene), m/z = 216 (methyl-fluoranthenes/pyrenes), m/z = 228 (benz(a)anthracene/chrysene), m/z = 242 (methyl-benz(a)anthracenes/chrysenes) and m/z = 231 (triaromatic steroids). The peaks were identified according to organic geochemical literature data (e.g., Peters et al. 2005 and the references therein), or based on the total mass spectra, using mass spectra databases (NIST/EPA/NIH mass spectral library NIST2000, Wiley/NBS registry of mass spectral data, 7th ed., electronic versions). The mixture of saturated hydrocarbons was separated into fraction of n-alkanes and fraction of branched and cyclic alkanes by urea adducts (Evans et al. 1957). The nalkanes in the urea adducts were analyzed by GC. The gas chromatographic analyses were conducted on a GC 8000 instrument (Fisons Instruments, Italy) with a ZB1 capillary column (Phenomenex, Germany; 30 m length; 0.25 mm inner diameter; 0.25 μm film). The carrier gas was hydrogen at a velocity of 40 cm s-1. 1 μL of samples was injected in a splitless mode (injector temperature 270ºC). The following temperature program was used: 80ºC at 3 min hold, then programmed at 5ºC min-1 to 300ºC.

3. CHARACTERISTICS OF THE PETROLEUM POLLUTED SOIL AND THE SUBSTRATE FOR BIOREMEDIATION Intensity of biodegradation of oil pollutants in soils is influenced by a large number of factors. The most significant of them are: presence of microorganisms capable of degrading hydrocarbons; quality, quantity and bioavailability of the oil pollutant;


Ex Situ Stimulated Bioremediation of a Soil Contaminated with Oil Pollutants 219 presence of macro- and micronutrients; presence and availability of electron acceptors; type and structure of soil; content of water; temperature and pH value (Langwaldt and Puhakka 2000; Bossert and Bartha 1984). The optimum conditions required for microbial activity in soil are: low clay or silt content; favourable structure of the soil (indicated as high porosity and low bulk density); soil moisture in the range of 30 - 90% of water holding capacity (minimum value is 25 28%); soil pH value within the 6.5 - 8.0 range (most microorganisms can tolerate wider pH range between 5.5 - 8.8); the ratio of basic nutrients, carbon, nitrogen, phosphorus and potassium (C/N/P/K ratio) 100:10:1:0.1; absence of heavy metals and other contaminants which might act as growth and activity inhibitors (Vidali 2001). The basic chemical and physico-chemical characteristics of the polluted soil used in this research and in the substrate for bioremediation were analyzed and monitored in a parallel independent research (Beskoski et al. 2011). These analyses comprised determination of: content of clay and sand, bulk density, water holding capacity, moisture (for original material), pH, organic and inorganic carbon, nitrogen and available phosphorus and potassium using standard methods (Wilke 2005; Pansu and Gautheyrou 2006). The content of oil pollutants was measured as a content of Total Petroleum Hydrocarbons (TPH; ISO 16703 and DIN EN 14345). Concentration of sand and clay in this soil was (61 ± 35) % and, accordingly, the soil was classified as a clayified sand. The moisture of the soil investigated was (17.8 ± 0.3) % of a water holding capacity. The pH value was 7.3 - 7.5. The measured value for TPH was (12.4  0.5) g kg-1. The concentration of heavy metals in the polluted soil was below that of reference values (Dutch Standards 2000), indicating that this soil did not contain these microorganism growth and activity inhibitors (Beskoski et al. 2011). From a total of 9 types of microorganisms identified in the zymogen consortium, 6 of them belong to the group of efficient petroleum hydrocarbons’ degraders. These are species from the genera Pseudomonas, Rhodococcus, Achromobacter and Stenotrophomonas (Varjani 2017; Bossert and Bartha 1984; Singh and Ward 2004). Concentration of total chemoorganoheterotrophs in the soil investigated was 1.2 × 6 10 CFU g-1. The proportion of hydrocarbons degrading bacteria comprised approximately 20% of the total number of bacteria in this soil (Beskoski et al. 2011), indicating presence of an intensive biodegradation processes and, additionally, large bioremediation potential of the investigated soil. All these results demonstrated that the chosen petroleum polluted soil could be a good substrate for monitoring the bioremediation of petroleum pollutant: the soil pH was in the optimum range, the TPH content was high enough with presence of an intensive biodegradative processes and, accordingly, large bioremediation potential. Additionally, the content of heavy metals, which are microorganism growth and activity inhibitors, in this polluted soil was below that of reference values. However, some parameters of this soil were not satisfactory for bioremediation purposes. The conditions for microbial


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activity in this soil would be better if the content of clay was lower. The moisture of the soil investigated was below the threshold value required for microbial activity in soils and the ratio of C/N/P/K in this soil was far from optimum. Accordingly, it was concluded that an improvement of the soil structure, bio stimulation, addition of water and addition of nutrients could make this soil an appropriate medium for the efficient biodegradation of the oil pollutant. In order to improve some of the aforementioned characteristics, the polluted soil was mixed with river sand and sawdust. As a result, during formation of the substrate for bioremediation, water holding capacity, porosity and ability to be mixed increased. pH value of the substrate for bioremediation remained in the optimum range for oil degradation (6.5 - 8.0). As a consequence of dilution, TPH value of the substrate for bioremediation (5.2  0.2 g kg-1) decreased in comparison with the polluted soil (12.4  0.5 g kg-1). Other parameters which are important for successful bioremediation (water content, aeration level, concentration of hydrocarbon degrading microorganisms and C/N/P/K ratio) were during the experiment kept at the optimum levels by regular watering, mixing, re-inoculation and spraying with a solution of dissolved ammonium nitrate, diammonium phosphate and potassium chloride (Beskoski et al. 2011).

4. THE DYNAMICS AND THE EFFICIENCY IN BIODEGRADATION OF SATURATED HYDROCARBONS Saturated hydrocarbons, and among them n-alkanes, are the most abundant compounds in most crude oils (Killops and Killops 2005). At the same time, these compounds are the most biodegradable components in both, crude oils in their source rocks (Peters et al. 2005; Head et al. 2003) and oil pollutants in the environment (Atlas 1981). Due to their biodegradability, n-alkanes are very useful indicators of progress of biodegradation process, especially at the beginning, when other compounds from the oil pollutant are still not affected (Peters et al. 2005). Early effects of the microbial degradation can be successfully monitored through the changes of the ratios: nC17/pristane, n-C18/phytane and n-alkanes/isoalkanes (Wang et al. 1998). Because of that, saturated hydrocarbons are usually the first ones to be analyzed during crude oil biodegradation studies. Total Ion Chromatograms (TICs) of saturated hydrocarbon fractions isolated from the control pile extracts at the beginning, in the middle, and at the end of the experiment, are shown in Figure 2. Monitoring of the changes in the distribution and the abundances of different compounds in the control pile actually gives an estimate of the transformations that occur during the natural microbial degradation of oil pollutant.


Ex Situ Stimulated Bioremediation of a Soil Contaminated with Oil Pollutants 221

Figure 2. Total ion chromatograms (TICs) of the saturated hydrocarbon fractions isolated from the control pile extracts at the beginning (CP1), in the middle (CP2) and at the end of the experiment (CP3; adapted from Ali Ramadan et al. 2013).

In all these samples, the fractions of total saturated hydrocarbons are characterized by a broad and prominent “hump” of an unresolved complex mixture (UCM; Figure 2). This shape of chromatogram is typical of oils altered by biodegradation (Gough and Rowland 1990; Killops and Al-Juboori 1990). Although n-alkanes are dominant compounds in most crude oils, as a result of biodegradation, UCMs of biodegraded oils contain a low amount of n-alkanes. The most dominant compounds in UCMs are usually alkyl-substituted and non substituted monoand polycyclic alkanes, isoprenoids and highly branched alkanes (Frysinger et al. 2003; Ventura et al. 2008). In the chromatograms shown in Figure 2, the peaks originating from n-alkanes are very low in intensity. Detailed analysis of these chromatograms revealed that the most abundant compounds in these UCMs were alkyl-substituted cyclic alkanes and branched alkanes. These results are in agreement with the current knowledge on the composition of the UCMs of biodegraded oils. However, based on these TICs, a precise conclusion about the intensity of microbial degradation of the investigated oil pollutant cannot be made.


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With the aim of more detailed analysis of the distribution and abundance of the individual n-alkanes in these samples, these compounds were concentrated by urea adduction technique, i.e., separated from branched, cyclic and polycyclic alkanes in a mixture of saturated hydrocarbons (Evans et al. 1957). Gas-chromatograms of the urea adducts (n-alkanes) of the saturated fractions, isolated from the control pile extracts at the beginning, in the middle, and at the end of the experiment are shown in Figure 3. In the gas chromatograms of all these samples, a homologous series of n-alkanes in the range from C16 to C31 was identified. Small differences in the shape and intensity of the UCMs in these chromatograms are result of non-complete separation of n-alkanes from branched and cyclic alkanes, which were dominant compounds in saturated fractions of these samples. Nevertheless, it is noticeable that the abundance and the distribution of the n-alkanes did not change during the six-month long biodegradation of the oil pollutant. Considering the fact that n-alkanes are usually completely removed from crude oil polluted soils after several weeks of biodegradation (Wang et al. 1998), these results might be surprising. However, as stated earlier, this oil pollutant had already been in this soil for more than a year before the beginning of the bioremediation process. Taking this fact into the account, degradation resistance of n-alkanes in this soil can be explained by a process termed ‘aging.’ This process refers to a gradual increase in desorption resistance of organic pollutant compounds which are in contact with soils and sediments over longer period of time and it does not include reactions that alter the structure of the molecule but only sorption of hydrophobic organic molecules to the soil matrix (Alexander 2000; Hatzinger and Alexander 1995). According to our results, it can be concluded that small amounts of n-alkanes, although being the most biodegradable compounds in crude oils, can survive in the soil environment for several months even in presence of the efficient native hydrocarbonbiodegradable microbial community. Total Ion Chromatograms (TICs) of saturated hydrocarbon fractions isolated from the biopile extracts at the beginning, in the middle, and at the end of the experiment, are shown in Figure 4. Comparing with the results from the parallel non-biostimulated biodegradation experiment (Figure 2), a similarity can be noticed. The TICs of saturated hydrocarbon fractions isolated from the oil pollutant at the beginning, in the middle, and at the end of the experiment are quite similar to the TICs corresponding to the samples analyzed in the present bioremediation experiment: all of them are dominated by the unimodal UCM maximizing at C20 while the peaks originating from n-alkanes are very low in intensity. However, based on these TICs only, a precise comparison between the microbial degradation intensity of oil pollutant investigated in these two experiments (bio stimulated and non-bio stimulated) cannot be made. Accordingly, a precise conclusion about the influence of bio stimulation factors on degradation of n-alkanes in the fraction of total saturated hydrocarbons cannot be drawn.



Figure 3. Gas-chromatograms (GC) of the urea adducts (n-alkanes) of the saturated fractions isolated from the control pile extracts at the beginning (CP1), in the middle (CP2) and at the end of the experiment (CP3; adapted from Ali Ramadan et al. 2013).

In order to investigate if in the conditions of stimulated bioremediation biodegradation of n-alkanes, which were found in a very low amount, and remained intact after six months of natural biodegradation, could be stimulated as well, these compounds were analyzed in more details. Aiming at facilitating their analysis, these compounds were isolated by urea adduction technique (Evans et al. 1957) and further analyzed by gas chromatography. Corresponding gas chromatograms are shown in Figure 5. In the gas chromatograms of n-alkanes isolated from all investigated samples, a series of homologues in the C16 to C31 range was identified. All these gas chromatograms are characterized by unimodal distribution of n-alkanes with n-C20 as the most abundant homologue. These results clearly show that the abundance and the distribution of these compounds remained unchanged during the stimulated bioremediation of the oil pollutant.


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Figure 4. Total ion chromatograms (TICs) of the saturated hydrocarbon fractions isolated from the biopile extracts at the beginning (BP1), in the middle (BP2) and at the end of the experiment (BP3).

After comparison with the results from the parallel non-bio stimulated biodegradation experiment, a great similarity between these results (as shown in gas chromatograms) can be noticed in n-alkane range as well as in the abundance and the distribution of the individual homologues. These compounds were identified in the same range in all samples investigated, and they remained at the same low level until the end of both, nonbio stimulated and bio stimulated biodegradation experiment. According to these results, it can be concluded that a low amount of n-alkanes from oil pollutant can persist in soil over six months, even in conditions of stimulated bioremediation. This persistence can be explained as a consequence of strong interactions of n-alkanes with the soil matrix (Alexander 2000; Hat zinger and Alexander 1995) which can further reduce their solubility, mobility and biodegradability in the environment even when these processes are intensively stimulated.



Figure 5. Gas-chromatograms (GC) of the urea adducts (n-alkanes) of the saturated fractions isolated from the biopile extracts at the beginning (BP1), in the middle (BP2) and at the end of the experiment (BP3).

5. THE DYNAMICS AND THE EFFICIENCY IN BIODEGRADATION OF AROMATIC HYDROCARBONS Aromatic hydrocarbons are one of the major components of crude oils, and accordingly, oil pollutants. In crude oils they typically represent 20 – 45% of the total hydrocarbons (Tissot and Welte 1978).


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Figure 6. Reconstructed ion chromatograms of phenanthrene (P; m/z = 178), methyl-phenanthrenes (MP; m/z = 192), dimethyl-phenanthrenes (DMP; m/z = 206) and trimethyl-phenanthrenes (TMP; m/z = 220), obtained by GC-MS analysis (using the single ion monitoring, SIM method) of the aromatic fractions isolated from the control pile extracts at the beginning (CP1), in the middle (CP2) and at the end of the experiment (CP3; adapted from Novakovic et al. 2011).

Aromatic fraction of crude oil is usually a complex mixture comprised from mono-, bi-, and polycyclic aromatic hydrocarbons with three and more condensed aromatic rings and their alkyl substituted derivatives (Radke 1987). Polycyclic aromatic sulphur heterocycles are often found in crude oil aromatic fractions as well (Peters et al. 2005). Distribution of all of these compounds in different oil derivatives depends on their production processes and the respective boiling ranges (Kaplan et al. 1997). Considering the fact that the primary mechanisms for removal of mono- and bicyclic aromatic hydrocarbons from oil pollution spills in soils are evaporative losses (Wang et al. 1998), these compounds will not be discussed in this study.


Ex Situ Stimulated Bioremediation of a Soil Contaminated with Oil Pollutants 227 Aromatic hydrocarbons with three and more condensed aromatic rings are considered less biodegradable than saturated hydrocarbons with the same number of carbon atoms (Peters and Moldowan 1993; Head et al. 2003). Furthermore, in the class of PAHs, susceptibility of these compounds to biodegradation decreases with increasing number of fused rings (Fingas 2015). These differences in biodegradability are even more pronounced in soil environment as a result of increased hydrophobicity and sorption capacity of higher-weight PAH molecules (Leahy and Colwell 1990).

Figure 7. Reconstructed ion chromatograms of dibenzotiophene (DBT; m/z = 184), methyldibenzotiophenes (MDBT; m/z = 198) and dimethyl- dibenzotiophenes (DMDBT; m/z = 212) obtained by GC-MS analysis (using the single ion monitoring, SIM method) of the aromatic fractions isolated from the control pile extracts at the beginning (CP1), in the middle (CP2) and at the end of the experiment (CP3).


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Figure 8. Reconstructed ion chromatograms of fluoranthene/pyrene (F/Py, m/z = 202), methylfluoranthenes/pyrenes (MF + MPy, m/z = 216) and dimethyl- fluoranthenes/pyrenes (DMF + DMPy, m/z = 230) obtained by GC-MS analysis (using the single ion monitoring, SIM method) of the aromatic fractions isolated from the control pile extracts at the beginning (CP1), in the middle (CP2) and at the end of the experiment (CP3).

In the samples from CP1 (the beginning of the experiment) to CP3 (the end of the experiment) a gradual degradation of phenanthrene and its methyl-, dimethyl- and trimethyl-isomers is noticeable (Figure 6). In this group of compounds, the effect of biodegradation is the most intensive for phenanthrene and then proportionally less intensive for its methyl isomers. The least affected are the homologues with the highest number of methyl groups, trimethyl-phenanthrenes. On the classification scale of oils according to their biodegradation level (Peters and Moldowan 1993; Head et al. 2003), the sample CP1 belongs to the level 3 (moderate biodegradation), the sample CP2 to the level 4 to 5 (heavy biodegradation) and the sample CP3 can be classified as level > 6 (severe biodegradation).



Figure 9. Reconstructed ion chromatograms of phenanthrene (P; m/z = 178), methyl-phenanthrenes (MP; m/z = 192), dimethyl-phenanthrenes (DMP; m/z = 206) and trimethyl-phenanthrenes (TMP; m/z = 220) obtained by GC-MS analysis (using the single ion monitoring, SIM method) of the aromatic fractions isolated from the biopile extracts at the beginning (BP1), in the middle (BP2) and at the end of the experiment (BP3; adapted from Novakovic et al. 2011).

A similar observation regarding the biodegradation dynamics and the efficiency can be made for dibenzotiophene and its methyl-isomers (Figure 7) and for fluoranthene/ pyrene and their methyl-isomers (Figure 8). In all these cases, parent (non-substituted) molecules are more degradable than methyl-substituted. Among the methylated homologues susceptibility to biodegradation gradually decreases with increase in the number of methyl groups in the molecules. Differences in the biodegradation efficiency between different polycyclic hydrocarbons investigated in this non-stimulated biodegradation experiment can be


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summarized as: phenanthrenes (most degradable) ≈ dibenzotiophenes > fluoranthenes/ pyrenes > chrysenes and triaromatic steroids (the least degradable; results not shown). Differences in the biodegradation efficiency between polycyclic hydrocarbons which belong to the same class of compounds but differ in number of methyl groups can be summarized as: parent molecule > methyl-isomers > dimethyl-isomers > trimethylisomers. All these changes during biodegradation can be characterized as typical, and they are in agreement with the published literature data (Peters et al. 2005; Fingas 2015).

Figure 10. Reconstructed ion chromatograms of dibenzotiophene (DBT; m/z = 184), methyldibenzotiophenes (MDBT; m/z = 198) and dimethyl- dibenzotiophenes (DMDBT; m/z = 212) obtained by GC-MS analysis (using the single ion monitoring, SIM method) of the aromatic fractions isolated from the biopile extracts at the beginning (BP1), in the middle (BP2) and at the end of the experiment (BP3).



Figure 11. Reconstructed ion chromatograms of fluoranthene/pyrene (F/Py, m/z = 202), methylfluoranthenes/pyrenes (MF + MPy, m/z = 216) and dimethyl-fluoranthenes/pyrenes (DMF + DMPy, m/z = 230) obtained by GC-MS analysis (using the single ion monitoring, SIM method) of the aromatic fractions isolated from the biopile extracts at the beginning (BP1), in the middle (BP2) and at the end of the experiment (BP3).

On the other hand, the results of the biodegradation resistance of n-alkanes obtained in this research are contrary to the current knowledge on the biodegradation susceptibility of these compounds. n-Alkanes are significantly more biodegradable compounds than aromatic hydrocarbons with three and more condensed aromatic rings. At the biodegradation level > 6, attained at the end of this biodegradation study, n-alkanes should have been almost completely degraded. On the contrary, they remained present in low concentrations and with unchanged distribution until the end of the experiment even when some of the aromatic hydrocarbons were severely affected by biodegradation. Accordingly, it can be concluded that under conditions of a reduced availability of certain


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classes of compounds, microorganisms opt for those that are, although less biodegradable, more accessible i.e., those that are present in higher amounts. Reconstructed ion chromatograms of phenanthrene, dibenzothiophene, fluoranthene/pyrene and their alkyl homologues obtained by GC-MS analysis of aromatic fractions isolated from the biopile extracts during the experiment are shown in Figures 9-11. According to these results, the differences in the biodegradation efficiency between different polycyclic hydrocarbons investigated in the biopile samples can be summarized as: phenanthrenes (most degradable) ≈ dibenzotiophenes > fluoranthenes/pyrenes > chrysenes and triaromatic steroids (least degradable; results not shown). In comparison with the parallel non-stimulated biodegradation experiment it can be noticed that this biodegradation sequence remained the same during the stimulated biodegradation experiment. However, the analysis of changes in TPH contents during this experiment in the control pile (natural biodegradation) and biopile (biostimulated process; Beskoski et al. 2011) revealed a significant difference between these two sets of samples. In the control soil, reduction in the TPH content during the experiment was only 10%. On the contrary, in the biopile at the end of the experimental period 94% of the TPH were biodegraded. These results demonstrate that although the biodegradation sequence of the main groups of aromatic compounds remained the same in the control- and the biopile, the biostimulated process was much more efficient. A detailed analysis of the hydrocarbons which belong to the same class of compounds but differ in number of methyl groups revealed another difference between non-stimulated and stimulated biodegradation experiments. In the samples from BP1 (the beginning of the experiment) to BP3 (the end of the experiment), a gradual degradation of phenanthrene and its methyl-, dimethyl- and trimethyl-isomers is noticeable (Figure 9). However, contrary to the results from the parallel non-stimulated experiment, in this group of compounds the effect of biodegradation is the most intensive for trimethylphenanthrenes and then proportionally less intensive for methyl isomers with lower number of methyl groups. The least affected is the non-substituted parent molecule phenanthrene. A similar observation regarding the biodegradation dynamics and efficiency can be made for dibenzotiophene and its methyl-isomers (Figure 10) and for fluoranthene/pyrene and their methyl-isomers (Figure 11). In all these cases, parent (non-substituted) molecules are less degradable than methyl-substituted. Among the methylated homologues, susceptibility to biodegradation gradually increases with increase in the number of methyl groups in the molecules. For the biostimulated process, differences in the biodegradation efficiency between polycyclic hydrocarbons which belong to the same class of compounds but differ in number of methyl groups can be summarized as: parent molecule < methyl-isomers
6. This result was explained as a consequence of aging of n-alkanes in the investigated soil.


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During this experiment, an intensive degradation of aromatic compounds occurred in both soils the stimulated and the control. n-Alkanes remained present in low concentrations and with unchanged distributions until the end of the experiment in the stimulated process as well. It can be concluded that during biodegradation of oil pollutants, under the condition of reduced availability of certain classes of compounds (caused by their low amount), microorganisms opt for those which are more accessible i.e., those which are present in higher amounts, even if these compounds are less biodegradable. The susceptibility of the main classes of aromatic compounds in both processes remained the same and followed the sequence: phenanthrenes (most degradable) ≈ dibenzotiophenes > fluoranthenes/pyrenes > chrysenes and triaromatic steroids (least degradable). As expected, the stimulated process was much more efficient regarding the removal of the total petroleum hydrocarbons from the soil. A detailed analysis of the aromatic hydrocarbons which belong to the same class of compounds but differ in the number of methyl groups revealed another difference between non-stimulated and stimulated biodegradation experiments. For the nonstimulated process (natural biodegradation), differences in the biodegradation efficiency between polycyclic hydrocarbons which belong to the same class of compounds but differ in number of methyl groups can be summarized as: parent molecule > methylisomers > dimethyl-isomers > trimethyl-isomers. These changes can be characterized as typical and they are described in the literature. For the biostimulated process, differences in the biodegradation efficiency between polycyclic hydrocarbons which belong to the same class of compounds but differ in number of methyl groups followed the opposite sequence: parent molecule < methyl-isomers < dimethyl-isomers < trimethyl-isomers. It was assumed that stimulation of the biodegradation promoted the process of decomposition of methylated homologues of the investigated aromatic compounds by favoring the bacterial strains in the consortium that preferably degrade higher methylated homologues. It can be concluded that bioremediation is a promising technology for removal of oil pollutants from polluted soils. With proper choice of the microbial consortium and the biodegradation conditions, the removal efficiency of the pollutant can be increased and the biodegradation sequence of present compounds can be governed.

ACKNOWLEDGMENTS We thank the Ministry of Education, Science and Technological Development of the Republic of Serbia (Projects 176006 and III 43004) for supporting this research.



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Chapter 10

THE BIOREMEDIATION OF WASTES FROM THE SEAFOOD INDUSTRY Saima1,* and Mohammed Kuddus2 1 2

Department of Biotechnology, Integral University, Lucknow, India Department of Biochemistry, University of Hail, Hail, Saudi Arabia

ABSTRACT This chapter focuses on the role of chitinases in the bioremediation of wastes from the seafood industry. Bioremediation is a cost effective and eco-friendly technology that is powered by a number of microbial enzymes. In this process, pollutants are degraded into the useful products through the biological process. Fishery is one of the most important economic activity on our planet. According to the Food and Agriculture Organization of the United Nations, total fisheries production worldwide was 167 million tons in 2014 that consists of the species rich in the chitinous materials. India alone produces 60,000 to 80,000 tonnes of chitinous wastes annually that causes environmental hazards. Chitin is one of the most abundant biopolymers widely distributed in the marine and terrestrial environments. Increased amount of seafood wastes, resulting from the industrial processing of seafood (shells of crab, shrimp, prawn, krill and lobster), possess a serious problem, both for the environment and for the processing plants. Chitin can be decomposed by the various approaches such as conventional, chemical and biochemical methods; but these methods are harmful for the environment due to release of harmful gases in the environment that cause global warming. Alternatively, large number of microbes and microbial enzymes have been reported to be involve in the biodegradation. Degradation of chitin into its oligomers or monomers by using chitinase producing microbes received much attention due to broad applications of the products in the fields of agriculture, medicine, food and biotechnology.

Keywords: seafood waste, bioremediation, chitin, chitinase, fishery *



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INTRODUCTION Industrial and agricultural activities are continuously releasing highly toxic and nondegradable compounds into the environment. Due to these compounds, number of aquatic and terrestrial ecosystems gets polluted or contaminated. These toxic compounds may be accumulating in the environment, due to unavailability of the specific strategy for the treatment and management of pollutants and waste disposal, and create hazards for the health and life of human beings, animals and microorganisms. Moreover, these pollutants contaminate soil and water, and become environmental concern in many countries. Now a day, fishing and pisci culture are arises a key trade to millions of people throughout the world. Japan is the largest seafood consuming country followed by European Union in the world. Salmon, shrimp, tilapia, catfish and crab (major consumption in China and India) are among the most consumed species. Seafood processing generates potentially vast quantity of organic wastes and by-products from unpalatable fish parts and endoskeleton of the crustaceans. The waste disposed to the aquatic environments, generated from sea food industries, can change the physicochemical characteristics of the water and also affect biotic components of the ecosystem, resulting in the loss of biodiversity. India is the second leading producer of aquatic organisms in the world. It is an important exporter of the shrimp to USA, Europe and Japan. Sea food industries not only generate water pollution but also create an air and noise pollution during the food processing and storage. Chitin is the one of main contaminant, generated from sea food industries, which is second most bountiful renewable polymer in the nature after cellulose; and possibly the most abundant in marine environment [1]. Waste generated from seafood processing operations of marine food products having huge amount of organic pollutants in soluble, colloidal and particulate form (chitin 20 to 58% of its dry weight) [2]. Kurita [3] reported that in the marine environment total chitin has been estimated around 1560 million tons. Due to its easy deterioration, chitin may lead to environmental hazards [4]. The standard and traditional methods for chitin extraction are chemical processes followed by demineralization with strong acid, de-proteination with strong base and also involve bleaching step to decolorized chitin [5]. Chitin deacetylation process involves strong chemical conditions [6]. It is being reported that chemical chitin purification is perilous, energy consuming and threat to the environment due to the involvement of high mineral acids and bases [7]. However, chitin extraction from the waste of sea food industries and use as a substrate for chitinase production reduces cost of chitinase production and can also solve problem of sea food industries. Many reports are available in which bioremediation of sea food industries waste was done by the application of microbial chitinases that degrade chitin. Bioremediation are being used to detoxify and degrade environment pollutants by the application of microorganisms since long times [8]. Microbes uses various mechanism for removing contaminants from the environment such


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as adsorption over cell surfaces, exopolysaccharides complexations, intracellular accretion, biosynthesis of metallothionins and other proteins that entrap metals and convert them into volatile compounds, and producing enzymes to degrade pollutants into useful products [9]. At present, research is being focused on the development of genetically modified microbes and their application for the detoxification and degradation of environmental pollutants.

BIOREMEDIATION Detoxification and degradation of environmental pollutants by microorganisms is known as bioremediation. Microorganisms secrete enzymes that degrade contaminants from the environment. Bioremediation can also be defined as "treatment that uses naturally occurring organisms to break down hazardous substances into less toxic or nontoxic substances". Typically, microbes like bacteria, archaea and fungi are the prime organisms used in bioremediation. The bioremediation makes environment free from pollutants and can be categorized into two major categories namely in-situ and ex-situ bioremediation. It is further divided into three types as described below.

Intrinsic Bioremediation This is also called as natural attenuation which is most effective in soil and water bioremediation. The process of intrinsic bioremediation is mostly used in underground places like underground petroleum tanks. In such places, only microorganisms can remove the toxins and clean the tanks. Intrinsic bioremediation is highly recognized method for the bioremediation of ground water (saturated systems) contaminated with hydrocarbons.

Biostimulation In this process, microbial populations are stimulated by modification of the environment such as addition of electron acceptors, electron donors, or nutrients. The modifications encourage indigenous organisms to boost biomass in order to increase decontamination.



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Bioaugmentation Bioaugmentation is used for municipal waste water remediation in which microorganisms are used to extract the contaminants. This is generally insistent aspect as microorganisms are added into the contaminated environment.

ROLE OF BIOTECHNOLOGY IN BIOREMEDIATION Bioremediation has been defined as “a biological response to environmental abuse” [10]. In the limited natural resources, rapid urbanization, industrialization and increasing population are creating highly critical situation for the environment. In such cases, biotechnology play a vital role with highly developed environmental management systems to remove the contaminated materials from the environment. Environmental biotechnology utilizes the genetic engineering to stimulate the metabolic activities of microorganisms that digest the contaminants with more efficiently and help in restoration of the environment. The possibilities of involving genetic engineering into the bioremediation process was started from late 1980’s. Microorganisms are genetically engineered in order to make them more effective in biodegradation, biotransformation, biosorption and bioaccumulation. Recombinant DNA techniques assist the ability of microorganism to metabolize xenobiotic compounds by recognition of the degradative genes and modify them into a suitable host [11]. Bioremediation is effective not only in pollutants degradation but also in air, soil, water and industrial waste decontamination.

SEA FOOD INDUSTRIES The seafood industry involves processes like culturing, processing, preserving, storing, transporting, marketing or selling fish or fish products. Shellfish are harvested from shoreline and fresh water bodies like lake or river by commercial fishermen. Processing sector converts harvested fish or shellfish into consumables that are sold in retail stores and restaurants. The processing of edible portion of the fish is energy consuming process as heat is used to eliminate most of the microorganisms that control the spoilage of products. This pasteurized products could be canned and stored with out refrigeration. Other process used for sterilization of the sea food products are high pressure or irradiations.



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Sea Food Processing Fish processing involve various steps such as grading, washing, stunning, beheading, gutting, scaling, skinning, cutting/filleting, packaging and freezing. It also involves marking/grading of the fishes according to their species and size which is very significant part of processing (i.e., smoking, freezing, heat treatment, salting, etc.) and marketing. Mechanical graders have improved sorting accuracy and automated sorters are used in small freshwater fish processing plants as the raw products are already sorted out on the delivery. Automatic graders are highly efficient, rather than the manual grading. The sorting speed of different graders are depending on the type of device and the size of fish to be sorted. Generally, the graders also worked as a conveyor.

Chitin in Sea Food Industrial Waste Chitin is the second most abundant polymer on the Earth, after cellulose. Chitin is the structural polymer of various organisms such as fishes, fungal cell wall, arthropods and in the micro filarial sheath of nematodes; acting as a protective layer against the brutal conditions [12, 13]. Crustacean shells contain protein (20–40%), calcium carbonate (20– 50%) and chitin (15–40%). About 100 billion tonnes of chitin generated every year [14]. The powder form of dried shrimp shells has a value of $100–120 per tonne due to their applications as an animal feed supplement, as well as in chitin production [15]. The global industrial use of refined chitin in drug delivery, food and cosmetic industries is around 10,000 tonnes per year [16]. Biodegradation of chitin is not easy, it takes long time and causes serious problem of sea food industries. Wastes from sea food processing industries contains valuable products like chitin, proteins, and various pigments whose quantities depends on the processing methods, types of sea food, body parts etc. [5]. Chitin and chitiosan prepared from crab and shrimp shells are now available as commercial products throughout the world [17]. Total wastes generated by shell fish processing varies between 65-85% and also depends on the processed species. The estimated chitin from the clam oyster shells might be 22,000 tons annually. Additionally, waste management is a major problem in seafood industries especially in the crustacean sector, [18]. Meyers and Chen [19] as well as Shahidi and Synowiecki [20] have reported the importance of the pigments, proteins and caroteno-proteins obtained from processing of discarded shrimp and crabs.



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GENERAL STRUCTURE AND PROPERTIES OF CHITIN Chitin, a natural polymer of N-acetyl-D-glucosamine (NAG) linked with each other by β-1,4 glycosidic linkage [21]. It is colourless to off-white, hard, inelastic, nitrogenous polysaccharide. Molecular weight of chitin ranges from 1.03×106 to 2.5×106 Da. There are many reports on its multidimensional properties in biomedical and other industrial applications. Chitin is biodegradable, biocompatible and bio-absorbable, with antibacterial and wound-healing abilities with small immunogenicity [22]. It is highly ordered crystalline structure and hydrophobic in nature due to intermolecular hydrogen bonds as recommended by X-ray solvents diffraction studies [23]. Among the all, the most crucial feature of chitin is its ability to convert into other forms such as in fibers, hydrogels, beads, sponges, cotton, powder, films, flakes and membranes [24]. Derivation source of chitin affects its crystalline nature, purity, polymer chain arrangement, and dictates its properties [25]. It is similar to cellulose by both in chemical structure and biological functions except the fact that one of the hydroxyl groups of each glucoside residue is substituted by either acetylated or deacetylated amino group (Figure 1). There are three different crystalline polymorphic forms of chitin namely α, β and γ [22].

α-Chitin It is composed of alternating sequence of parallel and anti-parallel chains (Figure 2a). It is more stable, strongly packed together due to its crystalline structure and abundant isomorphism occurs mainly in crustaceans, insects and fungi naturally. The chains in anti-parallel fashion favours well-built strong hydrogen bonding. Alpha chitin is insoluble and does not bulge in ordinary solvent.

(a)

(b)

Figure 1. Chemical structure of (a) chitin (b) cellulose.



(a)

(b)

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(c)

Figure 2. Three polymorphic configurations of Chitin (a) α-chitin, (b) ß-chitin and (c) γ-Chitin.

β-Chitin It is composed of parallel chains and occurs in marine organisms only (Figure 2b). βchitin is soluble in formic acid, form series of crystalline hydrate structures and can be swollen in water by intercalating between the stacks of chains [26]. Throught the treatment with acid (cold 6N hydrochloric acid or by solution in formic acid) and by 45% fuming nitric acid, β-chitin can be permanently transformed into α-chitin [27-29].

γ-Chitin It is a combination of both α- and β–chitins (Figure 2c). Chitin is a white, amorphous, semi-transparent polymer which is generally insoluble in common solvents like water, acid, alkali, ethanol and other organic solvent but soluble in concentrated hydrochloric acid, sulphuric acid, acetic acid and 78-97% phosphoric acid. The fundamental units of this substance are linked mutually by condensation reactions to build up long chains. Hydrogen bonds link the chains together and help to make chitin rigid and strong. Chitin can be synthesized in two different ways like as in fungi the enzyme chitin synthase occurs in vesicles called chitosomes as inactive zymogens and requires activation by proteolytic actions while in arthropods this is present as membrane-bound enzyme. The solubility of chitin gets increased by partial deacetylation under mild conditions without degrading the polymer, thus escalating the polarity and electrostatic repulsion of the amino groups. It’s being reported that chitins with acetylation degree of around 0.45–0.55 displays good solubility in aqueous media [30, 26].


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ROLE OF CHITINASE AND ITS PRODUCING MICROBES IN BIOREMEDIATION OF SEA FOOD INDUSTRIAL WASTE Chitinases (EC 3.2.1.14) are glycosyl hydrolase, catalyzing the hydrolysis of chitin between the C1 and C4 of two consecutive NAG unit by either an endolytic or an exolytic mechanism [31, 32]; and produced N-acetyl-D-glucosamine (GlcNAc) and Nacetyl chito-oligosaccharides, which has been applied as a biologically functional materials in many biotechnological processes [33]. Chitinases are generally found in plants, animals, insects, bacteria, fungi, viruses and actinomycetes [34]. Chitinase is applied for degradation of sea industries waste rich in chitin by using chitinase producing microorganisms and exploited in various application such as biocontrol agent against various pathogenic fungi [35], play role in bioremediation by degrading the chitinous waste emerged from sea food industries, also used in production of single-cell protein, and in the segregation of protoplasts from fungi and yeast, etc. [36]. Recently, the application of chitinases has been received great attention as producing valuable chitin from sea food industries waste and becomes great alternative to the chemical approach. Chitinase enzyme from Aspergillus terreus CBNRKR KF529976 [37] and Bionectria CBNR BKRR [37] was used effectively for the biodegradation of chitin from chitinous waste (shells of crab, shrimp, snail and fish scales) in order to produce extremely active chitinase enzyme. Moreover, Pre-treatment of shell fish waste enhanced the production of chitinase by using shrimp and crab shell powder (SCSP) as the carbon source. It was noticed that chemically treated SCSP induced a significant increase of enzyme production, as compared with untreated SCSP [38]. Many other reports are also available for degradation of different shell fish waste by chitinase producing bacteria to enhanced the chitinase production and was found that Tichoderma harzianum [39] on shrimp shellfish and A. terreus [40] on fish-scales were most favourable for chitinase production. Actinomycetes and some other microorganisms utilize shrimp shells as substrate for chitinase production more efficiently than that of colloidal chitin. Pseudomonas sp. and bacillus sp. also utilize shrimp shells and wastes more effectively to produce chitinases. Aspergillus sp. produces more chitinases if grown on shrimp wastes in comparison to colloidal chitin medium. In addition, the functions of chitinases in various organisms are different. In yeast, cells are incapable to part normally if they are devoid of chitinase activity during the logarithmic growth phase; instead they tend to aggregate at the septum regions as clusters [41]. In plant, chitinase secretion is induced by infections of chitin-containing microbes (fungi, insects) or other injuries [42, 43, 34]. Chitinases activity is induced by fungal, bacterial and viral infections since it is considered as pathogenesis-related proteins [4445]. In fungi, chitinase action plays a physiological task in apical growth and fungal hyphae morphogenesis [46-47]. In bacteria, chitinases functions as in nutrition and parasitism. In insect and crustaceans, chitinases are necessary for partial degradation of



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old cuticle [36]. In fish, it is usually a part of digestive tract to degrade chitin containing pray [48], while in human, chitinases can be found in gastric juices and they are involved in catabolic activity [49]. In addition to the above applications, chitinolytic enzymes are used in different sectors including agricultural, biological, biotechnological and environmental fields [50]. It can be utilized for the production of chito-oligosaccharides which do function as antibacterial agents, for elicitors of lysozyme inducers, and immunoenhancers [51]. Chitinases are also used for the controlling of plant pathogens in the agriculture sector [52-53]. These enzymes can be either directly employed for biological control of microorganisms, or indirectly by genetic manipulation of purified enzyme [54-55]. In recent years, chitinase genes from bacterial species like Alteromonas [56], Bacillus circulans [57], Vibrio harveyi [58] and Vibrio vulnifucus [59] is being cloned and characterized. The chitinases as antifungal protein has various biotechnological applications as in food, in seed preservation and to develop resistant plants against phytopathogenic fungi [60]. Chitinases belongs to a set of complex hydrolytic enzyme that catalyzes the depolymerization of chitin and it can be classified into two major categories based on their mode of action (Figure 3).

Figure 3. Cleavage site on chitin by exo- and endo-chitinase [61].

Class 1 Endochitinases (EC 3.2.1.14) or chitinase is poly β-(1,4)-2-acetamide-2-D glucoside glycanohydrolase which randomly slice chitin chain at internal sites, and producing multimers of GlcNAc, like chitotetraose, chitotriose, and diacetylchitobiose.

Class 2 Exochitinases can be categorized into two subcategories, chitobiosidases (EC 3.2.1.29), which catalyzes the progressive release of diacetylchitobiose early at the non-


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reducing end of chitin microfibril, and other β-(1,4)-N-acetyl glucosaminidases (EC 3.2.1.30), which slice the oligomeric products of endochitinases and chitobiosidases, producing monomers of GlcNAc [62].

ENVIRONMENTAL POLLUTION GENERATED BY SEA FOOD INDUSTRIES Sea food industries generally generate a number of wastes that may be either solid or liquid. Effluents and waste from sea food industries contains high amount of organic waste, proteins, oils and suspended solids. They also consist of elevated level of phosphates and nitrates. More than half of the all fish processing by-products, particularly those processed at sea, are not used. In 2000, the amount of pollock by-product was estimated to be 707,707 tons, consisting of 32% viscera, 26% heads, 33% frames, and 9% skin [63]. Waste may be used in the production of another product or discharged as a waste which is not only dangerous for environment but also for the fishery itself. The various types of environmental pollutions caused by sea food industries are described as follows.

Air Pollution from Sea Food Industries The odours emitted from the storage sites of processing waste, cooking, and drying are generally the major form of air pollution in fish processing which polluted the atmospheric air and emitted methane ammonia and merceptans to the atmosphere [64]. Seafood processing also generates air pollution by adding volatile gases or poisonous gases to the environment during freezing, cooking, sterilization, freeze drying and cleaning.

Water Pollution from Sea Food Industries An enormous quantity of the solid bio-waste like carapace (head shells) and abdominal (tail) shells are generated by the fish industries because shrimps are usually sold as headless or peeled or both. Organic wastes generated from sea food industries changes BOD and TSS of the water. Canning of the fish is considered as the second source of solid waste (30-65% of the fish) along with crustaceans and molluscs processing that also cause generation of considerable amounts of solid residues (20– 50%) [65]. In the UK, approximately 313,000 tons of seafood processing residues are produced annually, thus merely 43% of the catches end up as products for human



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consumption [66]. The exoskeleton of the fishes such as shrimp and crab processing waste contains 14-27% and 13-15% of chitin respectively, which is discharged into the water bodies as a waste material. Therefore, recovery of the chitin from the sea food processing industries would be beneficial for many biotechnological applications [67]. Chitin also possesses various applications in the field like drug delivery, dietary fibres, treatment of waste water, wound healing etc. Almost 75% of total weight of the shell fish is considered as waste which comprised of almost 20-58% of chitin itself [68].

Noise Pollution from Sea Food Industries Noise pollution is generally not a big issue associated with the seafood industries. During the food processing such as cutting, pre-cooking, filling and sterilization, only minute noise may be generated. The noise measurements at any point of processing should not exceed the standard of Maximum Sound Level that is 140 dB [69], but they have additional value than Equivalent Continuous Sound Level of 24 hours.

RECYCLING OF SEAFOOD INDUSTRY WASTE AND ITS USES Use as a Substrate for Chitinase Production/Bioremediation Chitinase is an enzyme which hydrolyses chitin into its oligomers. Chitinase has various applications in the field of biotechnology but the production cost is very high. Due to high amplified demand many studies has been carried out for the production of chitinase in which industrial shell waste is used as a substrate. In many studies, increased production was observed when shell fish is used as a substrate. The utilization of shell fish waste for chitinase production not only reduces the environmental contaminations but also decreases production cost of chitinase, along with its disposal problem. A large amount of chitinous waste are generated from sea food processing industries as it is estimated to be 1,00,000 tons annually in India [70]. Different substrates including fungal cell wall, crab shell, prawn shell and shrimp shell are used for the production of chitinase. Various studies concluded the use of seafood waste as substrate for chitinase production including crab shells [71], shrimp waste [5] and prawn shells [72]. Mejia- saules and coworkers [4] also concluded that crude shrimp waste may be used for chitinase production [4]. Some other reports on the production of chitinases by using shell waste as a substrate are summarized in Table 1.



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Table 1. Chitinase production by using shell fish waste as a product in bioremediation studies Microorganisms Bacillus licheniformis, Chromobacterium violaceum, Rhizobium radiobacter, Streptomyces griseous Bionectria CBNR BKRR Aspergillus terreus Bacillus licheniformis Kurthia gibsoni MB 126 Bacillus subtilis Aspergillus sp. S1-1 Beauveria bassiana BTMF S10

Substrate Fish scale waste, Prawn shell waste

References [73]

Fish scale, Crab shell, Snail shell, Shrimp shell Snail shell, Shrimp shell, Crab shell, Fish scales Shrimp shell waste Prawn shell powder Shrimp powder Shrimp shellfish waste Prawn waste

[74] [75] [76] [77] [78] [79] [80]

Use as Biofertilizers India is the agriculture based country and most of the people (approximately 60%) are farmer. Now a day, agriculture facing many problems due to lack of modernized infrastructure for promoting the agriculture sector. Current agricultural practices are not sustainable either economically or environmentally and India's agricultural yields for certain commodities are small. Soil naturally contains nitrogen fixing bacteria for the proper growth of the plants. However; continuous utilization of chemical fertilizers may obliterate these nitrogen-fixing bacteria. Chemical fertilizers are inorganic material which is added to the soil in order to sustain plant growth. Many synthetic fertilizers are acidic, consist of sulphuric acid and hydrochloric acid, which increases soil acidity and also harms the beneficial microorganisms of the soil. Like, urea causes ammonia release, which may lead to acid rain, contamination of groundwater and ozone depletion due to release of nitrous oxide by de-nitrification process. Groundwater contamination is hazardous and may cause lethal diseases such as gastric cancer, thyroid malfunctions, birth malformations, and hypertension along with testicular cancer. Fishing and aquaculture in India has a long history. India is being the second largest aquaculture producer in the world, and largest exporter of shrimp to USA. Waste disposal and by-product management of food processing industry is a major problem in the areas of environmental protection and sustainability [81]. Approximately 25-50% of the raw materials, mostly ground fishes, are consumed for primary products. Remaining 50-75% of the raw materials are considered as a waste and used for the production of low-valued products, or disposed off. Waste disposal cost from the shellfish processors was estimated to be around £2.7 million per year. Discarded wastes from the food processing industry are rich in nitrogen which is vital nutrient for the plant growth. Waste generated from sea food industries is mainly chitin and calcium and disposed to aquatic environment and



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exert adverse effects on the biota at the cellular level. Chitin is degraded from chitinase producing microorganism present in the sea food industries waste and involve in the nitrogen balance as shows in Figure 4. Now a day, sea food industrial wastes are used as fertilizers; for example, a crushed crab shell is used as nematicide in horticulture. Crustaceans waste is also used as a fertiliser by non-EU countries. The shell waste is cleaned, dried, and grounded to powder and then exported to some countries like Korea and Japan, where it is used in golf greens as a natural fertilizer.

Figure 4. Role of chitinase in maintaining carbon and nitrogen balance in the nature.

Quicklime Formation Quicklime is used in various industrial application and in waste water treatment. Use of quicklime in waste water treatment is cost effective because it can also prevent eutrophication (algae build up) on water surface by precipitating dissolved chemicals. Calcium hydroxides have high pH (above 12) which destroys cell membranes of harmful pathogens. Due to high pH, flies and other insects cannot contaminate treated biological wastes. The high pH leads to precipitation of metals present in the waste reducing their solubility. Adding lime make it easier for handling of waste by increases precipitation of solid content of the waste. Due to such application of quicklime, we recover quick lime



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from the sea food waste industries. It contributes in waste water treatment and also reduces the hardness of the water and can apply in various applications.

Other Applications Chitinase enzymes can convert solid shell waste’s chitin into valuable single cell proteins. Fungal resource for the manufacture of SCP is Saccharomyces cerevisiae, Candida tropicalis, Hansenula polymorpha, and Myrothecium verrucaria. Reports for the production of chitinases are from S. cerevisiae where more than 60% SCP was produced with less nucleic acid content (1% to 3%) [36]. NAG, monomer of chitin is also used in food products as sweeteners, as growth factors, chemicals, and as pharmaceutical intermediates [82]. Chito-oligosaccharides are produced by enzymatic degradation of chitin extracted from sea food industrial waste. Chito-oligosaccharides play vital role in plant defence system and has ample use in medicine. For example, chitohexose and chitoheptose have an antitumor effect. Moreover, some researches utilize shellfish to obtained chito-oligosaccharides. Chito-oligomers obtained from shellfish waste degradation have wide range applications in biochemical, food, and various chemical industries also [83].

CONCLUSION The published literatures concluded that at present sea food industries becomes a main income source for most of the people in India. Waste materials generated from industries itself harms fishery and environment. It produces water, air and noise pollution. We can use conventional method to reduce pollutants but it is also dangerous due to use of high amount of chemicals involve in these methods. So, chitinase enzyme along with its producing organisms are the best alternative to reduce pollutants; and the products generated from the degradation of chitin may be used in various field of biotechnology.

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[67] Wiseman, A. (1996). Handbook of Enzyme Biotechnology. 2nd Edition. USA: McGraw-Hill Book Company. [68] Patil, R.S., Ghormade, V. and Deshpande, M.V., (2000). Chitinolytic enzymes: an exploration. Enzyme Microb. Technol, 26(7): 473-483. [69] Notification of the Ministry of Interior, issued under the Announcement of the Revolutionary Party No. 103 dated November 12, B.E.2519(1976). [70] Gopakumar, K. (2002). Textbook of fish processing technology, Indian council of agricultural research, 467-483. [71] Jung, W.J., Souleimanov, A., Park, R.D. and Smith, D.L. (2007). Enzymatic production of N-acetyl chitooligosaccharides by crude enzyme derived from Paenibacillus illioisensis KJA-424. Carb. Poly., 67: 256–259. [72] Shirai, K., Palella, D., Castro, Y., Guerrero-Legarreta, I., Saucedo- Castaneda, G., Huerta-Ochoa, S. and Hall, G.M. (1998). Characterisation of chitins from lactic acid fermentation of prawn wastes. In: Advance in chitin science, Chen, R.H. and Chen, H.C. (ed.) vol. 3. Elsevier, Taiwan, pp. 103–110. [73] Zerubabel Michael, R.A. and Sri Rami Reddy, D. (2016). Production of Chitinase from Chromobacterium violaceum Using Agro Industrial Residues under Solid State Fermentation. International Journal of Innovative Research in Science, Engineering and Technology, 5 (3). [74] Krishnaveni, B., and Ragunathan, R. (2014). Chitinase production from Seafood wastes by Plant pathogen Bionectria CBNR BKRR SPS and its application in Bioremediation studies. International Research Journal of Medical Sciences, 2(11): 15-19. [75] Krishnaveni, B. and Ragunathan, R. (2014a). Chitinase production from marine wastes by Aspergillus terreus and its application in degradation studies. 3 (1): 7682. [76] Abirami, S., Ronald, J., Kannan, M. and Ranjit Singh, A.J.A. (2012). Bioconversion of Shrimp Shell Waste by Bacillus Licheniformis for the Production of Antifungal Chitinase Enzyme. Res. J. Pharma. Biol. Chem. Sci., 3(4): 789-796. [77] Paul, M.K., Mini, K.D., Ally, C.A., Radhakrishnan, E.K. and Jyothis, M. (2012). Utilization of prawn shell powder for the production of chitinase by Kurthia gibsonii MB126. Int. J. Pharm. Bio. Sci., 3(3):163-172. [78] Wang, S., Lin, T., Yen, Y., Liao, H. and Chen, Y. (2006). Bioconversion of shelfish chitin wastes for the production of Bacillus subtilis W-118 chitinase. Carbohyd. Res., 341: 2507-2515. [79] Rattanakit, N., Plikomol, A., Yano, S., Wakayama, M., and Tachiki, T. (2002). Utilization of shrimp shellfish waste as a substrate for solid-state cultivation of Aspergillus sp. S1-13: Evaluation of a culture based on chitinase formation which is necessary for chitin-assimilation. Journal of Bioscience and Bioengineering, 93(6): 550-556.


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[80] Suresh, P.V. and Chandrasekaran, M. (1998). Utilization of prawn waste for chitinase production by marine fungus Beauveria bassiana by solid state fermentation. 14(5): 655-660. [81] Russ, W. and Pittroff, R.M. (2004). Utilizing waste products from the food production and processing industries. Crit Rev Food Sci Nutr., 44(2):57–62. [82] Felse, P.A. and Panda, T. (2000). Production of microbial chitinases-A revisit. Bioprocess and Biosystems Engineering, 23 (2): 127-134. [83] Sakai, K., Yokota, A., Kurokawa, H., Wakayama, M. and Moriguchi, M. (1998). Purification and characterization of three thermostable endochitinases of a noble Bacillus strain, MH-1, isolated from chitin-containing compost. Appl. Environ. Microbiol., 64: 3397-3402.




Chapter 11

STRATEGIES FOR PLASTIC WASTE MANAGEMENT Shikha Raghuwanshi* and Reeta Goel Department of Microbiology, College of Basic Sciences and Humanities, G. B. Pant University of Agriculture and Technology, Uttarakhand, India

ABSTRACT Waste is increasing at an alarming rate due to the inevitable use of the plastics in various industries like aerospace, agriculture, medical, electronics etc. The inert and recalcitrant nature of the plastics poses a great threat to the living creatures and the environment. Therefore, plastic waste management is a huge concern to the governments around the globe and hence targeted research to find out eco-friendly methods is needed to address this issue. It has been reported that an array of micro-organisms can degrade many of the plastics. Various enzymes and biosurfactants facilitate these microbes to utilize plastic as a sole carbon source during bioremediation. The biodegradation in bioremediation is confirmed by various analysis techniques like Fourier transform infrared (FT-IR) spectroscopy, scanning electron microscopy (SEM), thermal gravimetric analysis and thin layer or gas chromatography. Different in vitro and in situ bioremediation approach has been carried out because of its eco-friendly and economic nature. The chapter highlight recent findings related to plastics and systematically review various bioremediation techniques used in the plastic waste management. The methods for the determination of bioremediation through analysis of degraded sample have been discussed in detail.

Keywords: plastics, plastic waste management, bioremediation, plastics-degrading microbes, Fourier transform infrared spectroscopy, scanning electron microscopy, enzymes, and surfactant *



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Shikha Raghuwanshi and Reeta Goel

1. INTRODUCTION In the modern era, petroleum-based plastics are considered to be one of the most important and commonly used commodities of our daily lives. A range of valuable qualities such as lightness, cost-effectiveness, transparency, durability, and nonbiodegradability has made plastic a versatile synthetic material. Hence, plastics has successfully replaced other materials in wide variety of applications in domestic, medical, industrial, automotive and agricultural field. An array of traditional plastics has enhanced the grade and ease of life [Polypropylene (PP), Polystyrene (PS), Polyvinyl chloride (PVC), Polyethylene terephthalate (PET), Low-density polyethylene (LDPE) and Highdensity polyethylene (HDPE)] (Gewert et al., 2015). The global production of plastic materials was estimated to be 300 million tons in 2015. Every year, 34 million tons of plastic waste is generated worldwide (Mekonnen et al., 2013; Emadian et al., 2016). As a result of widespread demand and production, the consumption of plastic material is expected to grow in future and eventually, an increament in the quantity of plastic waste in the environment (Anjum et al., 2016). Due to the carbon dioxide (CO2) emission and long-term persistence in the environment, plastic waste has become a universal nuisance to the well-being of the environment, human, and other living beings (Lavers and Bond, 2017). Solution for the plastic waste management includes incineration, landfilling and recycling. These traditional methods are very challenging in terms of time, cost, and toxic by-products. Thus, in order to manage the plastic waste and to maintain sustainable environment, bioremediation found its way as an economic and eco-friendly technique (Joutey et al., 2013; Krueger et al., 2015). A variety of micro-organisms (including bacteria and fungi) has been identified to degrade the plastics i.e., Bacillus, Pseuodomonas, Enterobacter, Brevibacillus, Microbacterium, Ralstonia, Streptomyces, Arthrobacter, Aceinobacter, Aspergillus, Fusarium, Penicillium, Verticillium, Mucor etc. In current years, an array of enzyme viz. esterses, peroxidases, lipases, proteases, urease, and cutinases has been reported to degrade or modify the synthetic plastics (Wei and Zimmermann, 2017) Therefore, bioremediation with bacteria, fungi and enzymes, provide a promising tool for plastic waste management. Bioremediation can be in vitro (laboratory), and in situ (natural) on the basis of site of action. It can also be aerobic or anaerobic process, on the basis of oxygen demand of the microbes. Literature survey revealed the various techniques used to monitor the extend of degradation in the form of changes such as chemical, structural, thermal, nuclear, morphological and surfacial in the degraded plastics during bioremediation. These techniques include Fourier transform infrared spectroscopy (FTIR), scanning electron microscopy, thermal gravimetric analysis, nuclear magnetic resonance, chromatography etc. (Santo et al., 2013; Raghuwanshi et al., 2015; Anwar et al., 2016).


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The aim of present review is to summarize the traditional and recent methods for plastic waste management. Based on published literature, the comprehensive summary of microbes and techniques involved in bioremediation of plastics has been discussed.

2. PLASTICS Plastics are man-made polymers which contain carbon, hydrogen, silicon, oxygen, chloride and nitrogen (organic and inorganic components). These are non-metallic and thermoplastic in nature (have ability to mould into desired shape). Most commonly used plastics are polyethylene, poly (ethylene terephthalate), polypropylene, polystyrene, polyvinyl chloride, polyurethane and many more. Oil, natural gas and coal are basic raw materials for the synthesis of synthetic plastics, and therefore, called fossil-based or petrochemical plastics. The structural formulae of plastics which are studied in this chapter are given in Figure 1. The details of plastics and their applications are listed in Table 1. General properties of plastics are as follows: a. Complex 3D structure b. Recalcitrant nature c. Absence of free functional groups d. High degree of branching e. Tends to be waterproof f. Easy to shape g. Possess low density h. Good electrical insulator i. Resistant to corrosion and other chemical factors j. Good barrier for oxygen and other gases. According to the PlasticsEurope, 2016 report, China shares about 27.8% of world’s plastic production followed by Europe and NAFTA (18.5%). In the last decade, the world’s plastic production has been increased by approximately 100 million tons. The various sector shares in global demand of plastics are given in Figure 2. Packaging sector uses LDPE, HDPE, PP, PET and PS plastics. From the application point of view, packaging covers a large segment of about 39.9% and has highest recycling rate of 39.5% (more than 80% of total recycled amount).


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Figure 1. Structural formulae of some common plastics reviewed in the study. a) Polyethylene terephthalate (PET), b) Polyethylene (PE), c) Polyvinyl chloride (PVC), d) Polypropylene (PP), and e) Polystyrene (PS).

Figure 2. Pie distribution of plastic demand in different sectors in 2015 (PlaticsEurope, 2016) (others include consumer and household goods, furniture, health, sport, health and safety).



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Table 1. An overview of plastic properties and its applications Type of Plastic Symbol

Chemical formula (C10H8O4)n

Properties

Applications

Tough and clear, good strength and stiffness, chemical and heat resistant, good barrier properties for oxygen and carbon dioxide

Water and carbonated soft drink bottles, food containers, pillow, sleeping bag filling, clothing fibres, appliance industry, and transport

High-density polyethylene

(C2H4)n

Milk and shampoo bottles, pipes, injections, carrier bags, storage bins, and buckets .

Polyvinyl chloride

(C2H3Cl)n

Good processability, excellent balance of rigidity and impact strength, excellent chemical resistance, crystalline, melting point (130-1350C), and excellent water vapour barrier properties Versatility, energy saving, and fire resistance

Low-density Polyethylene

(C2H4)n

Polypropylene

(C3H6)n

Polystyrene

(C8H8)n

Glassy surface, clear to opaque, rigid, hard, high clarity, affected by fats and solvents

Other miscellaneous plastics like polycarbonate, polyurethane etc.

C16H18O5)n

Engineering sector

Polyethylene terephthalate

(C27H36N2O10)n

Low density, semi crystalline, low melting range and softening point, good chemical resistance, excellent dielectric properties, low moisture barrier, and poor stretch resistance Low density, excellent chemical and environmental stress resistance, high melting point, dielectric properties, low cost, creep resistance

Automobile seat covers, shower curtains, raincoats, bottles, shoe soles, garden hoses, electricity pipes and fittings, electrical wire insulation, floor coverings, synthetic leather products, and furniture Containers, wash bottles, tubing, laboratory equipments, carrier bags, bin liners, agriculture mulch, packaging films, wire and cable insulation

Plastic moldings, stationary folders, packaging materials, plastic tubs, non-absorbable sutures, diapers making bottles, medical containers, pipes, sheets, straws, films, furniture, house wares, luggage, toys, hair dryer, and fan Disposable cups, packaging materials, laboratory wares, certain electronic uses, and wall tiles

Plastic foams, cushions, coatings, rubber goods, synthetic leathers, adhesives, paints, fibers, tyres, gaskets, bumpers, in refrigerator insulation, sponges, and life jackets



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3. TRADITIONAL METHODS FOR PLASTIC WASTE MANAGEMENT The industrial production of fossil-based plastics started in the 1940s. Since then, the rate of manufacturing, consumption and as a result, waste generation has been increased considerably. Due to improper waste management, pollution has been created; therefore, plastic waste management attained focus of research communities and the government since last few decades. In 2011, global declaration has been signed by 65 associations in 34 countries. In view of above, 260 projects are underway, planned or completed (PlasticsEurope, 2016). The traditional methods used for the management of plastic waste with their inherent limitations are shown in Figure 3. These methods include incineration, recycling, landfilling (Zhang et al., 2004; Webb et al., 2013).

Figure 3. Common and traditional remediation practices for plastic waste management and their constraints.

3.1. Landfilling In landfilling, waste material is dumped inside the land where it undergoes photodegradation (UV degradation) and oxidation. The UV degradation is based on the use of UV light to degrade the waste. The oxidation process involves heat to breakdown the plastic.



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The major drawback of land filling is that it acts as carrier for a number of secondary environmental pollutants which include volatile organics, such as benzene, toluene, xylenes, ethyl benzenes and trimethyl benzenes. These are released as gases and also in the form of leachate (Urase et al., 2008). Beside that the land used in landfilling cannot be reused for several decades. Moreover, countries such as Switzerland, Austria, Netherlands, Germany, Sweden, Denmark, Belgium, Norway, and Luxembourg banned the landfill (Anderson et al., 2016). In these countries, more energy recovery and recycling methods are used to manage waste.

3.2. Incineration One of the routinely used method for plastic waste management is incineration which involves the combustion of organic materials in the plastic waste. This is also referred to as thermal treatment. It lowers the requirement of land. Also the, energy recovered is used for electricity generation. This way of managing plastic waste is of great concern because of released ash, gases by burning of waste (Hopewell et al., 2009). The burning of polyvinylchloride (PVC) produce persistent organic pollutants (POPs) such as polychlorinated diphenyls, furans and dioxins (Shah et al., 2008). Carbon and oxygen free radicles, carbon dioxide, and heavy metals are released during the process which is harmful for the environment directly or indirectly.

3.3. Recycling It is of two types namely mechanical (secondary recycling) and chemical (tertiary recycling). Single-monomer plastics i.e., PE, PP and PS can be recycled mechanically. Chemical recycling is an advanced technology in which smaller molecules (basically liquid or gas) are produced from plastic materials, which are used as a substrate for the synthesis of petro-based chemicals and plastics (Al-Salem et al., 2009). By recycling process, we cannot get rid of the plastic waste and its harmful effects as only 15% plastic waste is recyclable. Basically, it is a down cycle process in which one plastic form recycled to another form. Therefore, recycling of plastic is not a sustainable solution. Although, this process overcome the major environmental shortcomings of landfilling and incineration. But, it is a relatively expensive and inefficient process as the quality of product and procedure get affected adversely by the presence of impurities.


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4. BIOREMEDIATION From past 3-4 decades, plastic waste management is in concern not because of ecological point of view but also, due to stability and integrity of plastics. As traditional methods are not very effective so to overcome this issue, bioremediation finds a solution as viable strategy. In short, bioremediation is exploitation of biological agents in various processes for the removal of unwanted, hazardous pollutants present in the environment (Das and Adholeya, 2012).

Figure 4. The mechanism and agents involved in bioremediation of plastics.

The main four steps involved in bioremediation are as follows (Nanda et al., 2010) (Figure 4). 1) First step is biodeterioration, here, the microbial growth on or inside the surface of the polymers leads to the superficial modification of different properties (i.e., mechanical, chemical, and physical) of the polymer. 2) Second step is biofragmentation, which is characterized by the conversion of complex polymeric chain to oligomers, dimmers, and monomers by the action of catalytic agents secreted by microbes. 3) The produced fragments i.e., oligomers, dimers and monomers, act as the essential sources of carbon, energy and the nutrient for the microbes, that is assimilation and 4) Finally, the mineralisation take place and fragments are converted to CO2, water and biomass (Shah et al., 2008).



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4.1. Agents Involved in Bioremediation A number of micro-organisms including both bacteria and fungus, act as promising tool for bioremediation to clean up the environment (Singh et al., 2014). These biological agents uses plastic as sole carbon source for their growth. The microbes reported in bioremediation of plastics are listed in Table 2.

4.1.1. Bacteria Bacteria are prokaryotic micro-organisms and are omnipresent in nature. They can acclimatize any extreme environment (high or low temperature, salinity, pressure and pH) and perform various physiological functions. For the isolation of potential strains, extreme environmental samples are most suitable because those bacteria are best adapted to extreme conditions (Sahoo and Dhal, 2009; Dash et al., 2013). 4.1.2. Fungus Fungus is eukaryotic micro-organisms. Fungus can sustain in the environment with low nutrient availability as well as low pH and moisture. Bioremediation using fungus is called as mycoremediation. In the literature, mushroom was also reported to play a crucial role in mycoremediation (Purnomo et al. 2013; Kulshreshtha et al. 2014).

4.2. Degradation Pathways Degradation is basically the change of chemical or physical properties. The different degradation pathways followed by microbial communities in the environment to degrade plastics are: Physicochemical, biological and chemical (Da Luz et al., 2014) (Figure 5). The physicochemical pathway involves, photo-degradation, thermal degradation and mechanical degradation, whereas, chemical and biological pathway involves oxidation and hydrolysis. Photo-degradation is the process of decomposition of the polymeric material by the action of light especially UV (400–290 nm) and visible (700-400 nm) which is initially slow and further propagation is fast. These high energy radiations cause chemical transformations. Here, heat is not required. Considerably, this process is very costly. Wherein, thermal-degradation, heat is the primary requirement to induce molecular scission and this is comparatively fast process. But, this is not an environmentally accepted phenomenon. Under normal conditions, photo and thermal degradation are similar (Fotopoulou and Karapanagioti, 2017). Mechanical degradation is due to the physical parameters like temperature, pressure and moisture and many other stresses,


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which mechanically break the polymers under shear or mechanical forces (Kale et al., 2015). Furthermore, oxidation and hydrolysis refers to the susceptibility of polymers to atmospheric oxygen, and water present in the vicinity. Biological and chemical degradation both follow oxidation and hydrolysis and both are based on chemical dissolution of the polymer.

Figure 5. Different degradation pathways in the environment.

Figure 6. Classification of enzymes with their functions.


Table 2. The list of reported bacteria and fungus involved in plastic bioremediation Plastic as substrate

LDPE

HDPE

PP

PS

PET

Bacterial candidate

Reference

Fungal candidate

Reference

Brevibacillus borstelensis Acinetobacter baumannii Rhodocccus strain Chelatococcus sp. E1 and Pseudomonas aeruginosa Bacillus cereus, Bacterium Te68R, Bacillus cereus, Proteobacterium sp. and Arthrobacter luteolus Microbacterium sp. and Pseudomonas putida Lysinibacillus fusiformis Arthrobacter sp. and Pseudomonas sp.

Hadad et al. (2005) Pramila and Ramesh (2015) Koutny et al. (2009) Jeon and Kim (2013)

Lasiodiplodia theobromae Penicillium sp Aspergillus niger and A. japonicas Aspergillus and Fusarium sp

Sheik et al. (2015) Yamada-Onodera et al. (2001) Raaman et al. (2012) Das and Kumar (2014)

Sowmya et al. (2014) Soni et al. (2009)

A. clavatus strain JASK1 Geomyces pannorum

Gajendiran et al. (2016) Barratt et al. (2003)

Kapri et al. (2010a)

Aspergillus sp. and Paecilomyces lilacinu

Sheik et al. (2015)

Aspergillus japonicas and A. terreus

Immanuel et al. (2014)

A. terreus Penicillium oxalicum NS4 and Penicillium chrysogenum

Balasubramanian et al. (2014) Ojha et al. (2017)

Cacciari et al. (1993) Arkatar et al. (2009) Yang et al. (2015) Tang et al. (2017) Atiq et al. (2010)

Aspergillus niger Lasiodiplodia theobromae Gleoeophyllum trabeum Curvularia sp.

Cacciari et al. (1993) Sheikh et al. (2015) Krueger et al. (2015) Motta et al. (2009)

Yoshida et al. (2016) Goel et al. (2014)

Penicillium funiculosum Aspergillus sp, Penicillium sp and Fusarium sp

Nowak et al. (2011) Umamaheswwari and Murali (2013)

Rhodococcus rubber Enterobacter asburiae YT1 and Bacillus sp. YP1 Bacillus cereus, Bacillus pumilus and Arthrobacter luteolus Pseudomonas and Vibrio sp. Pseudomonas and Bacillus Exiguobacterium sp. strain YT2 Tenebrio molitor and Zophobas morio Microbacterium sp., Paenibacillus urinalis, Bacillus sp. and Pseudomonas aeruginosa Ideonella sakaiensis Bacterium, Microbacterium and Pseudomonas putida

Mukherjee et al. (2017) Balasubramanian et al. (2010) Gilan et al. (2004) Yang et al. (2014) Satlewal et al. (2008)


Table 2. (Continued) Plastic as substrate

PVC

PUR

PCL

Bacterial candidate

Reference

Fungal candidate

Reference

Pseudomonas otitidis, Acanthopleuribacter pedis Bacillus cereus and Acanthopleurobacter pedis strains Pseudomonas aeruginosa and Aureobasidium pullalans, Rhodotorula aurantiaca and Kluyveromyces sp. Curvularia sp, Trogia buccinalis and Phanerochaete chrysosporium Corynebacterium and Pseudomonas aeruginosa Pseudomonas protegens strain Pf-5 Alternaria sp.

Anwar et al. (2013); Anwar et al. (2016)

Phanerochaete chrysosporium

Ali et al. (2014)

Webb et al. (2000); Kawai (2010)

Entinus tigrinus, A. niger and A. sydowii

Sachin and Mishra (2013)

Caruso (2015)

Alcaligenes faecalis and Clostridium botulinum Arthrobacter and Enterobacter species

Ghosh et al. (2013)

Fusarium solani and Aureobasidium pullulans Geomyces pannorum Plectosphaerella, Nectria, Neonectria, Phoma, and Alternaria; Pestalotiopsis microspora Fusarium sp.

Sandra et al. (2010) Kay et al. (1991) Hung et al. (2016) Matsumiya et al. (2010)

Goel et al. (2008)


Bhardwaj et al. (2012) Cosgrove et al. (2007); Russell et al. (2011) Kim and Rhee (2003); Arefian et al. (2013)


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5. PRODUCTS OF MICROBES FOR BIODEGRADATION 5.1. Enzymes Enzymes are polymeric macromolecules composed of amino acids connected by amide bond and mainly function as biocatalyst. Enymes process specific functional groups. The molecular mass of enzymes ranges from kilodaltons to megadaltons. The main functional site that is catalytic sites, determines the specificity of enzymes to the substrate which are present deeply in hydrophobic regions.

5.2. Surfactant Biosurfactants are amphiphilic; surface-active biomolecules produced by an array of microbes. The biosurfactants are broadly classified into two major classes on the basis of molecular weight, namely low and high molecular weight biosurfactants. Low molecular weight biosurfactants includes glycolipids, lipopeptides, phospholipids (rhamnolipids, surfactin, polymixin, trehalolipids, viscosin,). High molecular weight biosurfactants are further of two types on the basis location of density of hydrophobic regions. Firstly, amphiphilic polymer in which hydrophobic region are aggregated at one end of the molecule, e.g., lipopolysaccharides, lipoglycans lipoteichoicacids, etc. Secondly, polyphilic polymers, here, hydrophobic groups distributed over the entire molecule, e.g., Emulsan, biodispersan, hydrophobic polysaccharides (Vijayakumar and sarvanan, 2015).

6. EFFECTS OF MICROBIAL ACTIVITY ON PLASTICS The colonization of bacteria on the surface of polymers leads to the several changes in the polymeric properties. According to literature survey, the extent of biodegradation has been monitored based on several properties i.e., hydrophobicity, crystallinity, surface morphology, mechanical properties, functional group on the surface of plastic, molecular weight distribution, and consumption of the polymer (Restrepo-Floreza et al., 2014). Various techniques are used to examine the changes induced by microbial activity at chemical, physical, surfacial and mechanical level (Figure 7) (Alshehrei, 2017). The Enzyme Commission (EC) for classification for enzymes has been established with co-ordinaion of International Union of Biochemistry and Molecular Biology (IUBMB) and International Union for Pure and Applied Chemistry (IUPAC) which classified enzymes into six classes. Figure 6 shows these classes along with the major function or reaction catalysed by them (Liese et al., 2006).



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The characteristics like reduced process time, economic, nontoxic, good quality products and eco-friendly nature imparts enzymes a greater value (Li et al., 2012; Choi et al. 2015). Due to aforementioned features, enzymes have vast range of applications in various industries such as beverage, dairy, baking, animal feed, pulp and paper, polymer, detergent, leather, organic synthesis, cosmetics, waste management (Gurung et al., 2013). The various enzymes such as depolymerase, laccase, peroxidase, urease, protease, esterase, lipase, cutinase, hydrolase, and peroxidase are reported for the degradation of plastics and are listed in Table 3.1 (Singh et al., 2016). Table 3.1. Microbial enzymes play role in plastics biodegradation Plastic as sole substrate* PE

PUR

PVC

PET

Enzyme

Microbe

Reference

Magnese peroxidase, Lignin degrading enzyme Esterases, Ureases, proteases Polyurethanase Fungal lignin peroxidase Laccase Cutinase Hydrolase Carboxylesterase

Phanerochaete chrysosporium

Liyoshi et al. (1998); Ehara et al. (2000)

Comamonas acidovorans

Nakajima-Kambe et al. (1999); Howard (2002); Cregut et al. (2013) Howard et al. (2012b) Khan et al. (2017) Sumathi et al. (2016)

Lipase and cutinase

PCL

Depolymerase Cutinase

Acinetobacter gerneri Phanerochaete chrysosporium Cochliobolus sp. Thermobifida alba Thermobifida fusca Thermobifida fusca Candida Antarctica and Humicola insolens Streptomyces thermoviolaceus Fusarium solani Azotobacter beijerinckii

Ribitsch et al. (2015) Barth et al. (2015a) Billig et al. (2010) Carniel et al. (2016) Chua et al. (2013) Singh et al. (2016)

Hydroquinone Nakamiya et al. (1997) peroxidase * PE = Polyethylene, PUR = Polyurethane, PVC = Polyvinyl chloride, PET = Polyethylene terephthalte, PCL = Polycaprolactones, PS = Polystyrene. PS

Table 3.2. An overview of reported microbial biosurfactants S.No. 1. 2.

Type of biosurfactant Cyclic lipopeptide Sophorolipid

3. 4. 5. 6. 7.

Rhamnolipid Anionic lipopeptide Surfactin Anion surfactant Glycolipids

Microbe Achromobacter sp. HZ01 Candida bombicola Trichosporon ashii P. aeruginosa PG1 Candida glabrata Bacillus licheniformis Lysinibacillus fusiformis Candida ishiwadae

Reference Deng et al. (2016) Reddy et al. (2016) Chandran and Das (2010) Patowary et al. (2017) Lima et al. (2017) Vijayakumar and Sarvanan (2015) Mukherjee et al. (2017) Thanomsub et al. (2004)



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Table 4. List of changes occurred during biodegradation of polymer Changes observed

Technique

Property measured

Functional groups on the surfacial region of the polymer Hydrophilicity

FT-IR

Keto-carbonyl and double bond

Contact angle

Molecular weight distribution Surface morphology

GPC

Crystallinity

FTIR, DSC and XRD Instron Gravimetric Carbon dioxide emission Uv-visible spectrophotometer

Mechanical properties Consumption of polymer Biomass accumulation

SEM and AFM

Effect on property Decrease or increase

References

Contact angle with water Distributed molecular weight Topography

Small

Roy et al. (2008)

Increase or slight change Superficial damage

Fontanella et al. (2010)

Percent crystallinity

Increase

Tensile strength Weight loss

Change Increase

Tribedi and Sil (2013) Pramilla and Rameh (2011b)

Optical density and lambda- max shift

Increase or change

Raghuwanshi et al. (2015); Soni et al. (2009)

Gilan et al. (2004); Nowak et al. (2011)

Raghuwanshi et al. (2016); Raghuwanshi et al. (2017) Sudhakar et al. (2007)

Figure 7. Techniques used for analyzing the changes at different levels after bioremediation.



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These may remain on microbial cell surface or secreted out by the microorganisms. They lower the interfacial tension at the surface of the liquid or at the interface of two immiscible liquids and enhance the biodegradation of plastics by increasing the solubility and bioavailability of hydrophobic or insoluble organic compound (Singh and Sedhuraman, 2015). Furthermore, film formation, swarming motility, and antimicrobial activity are also being assisted by surfactants (Van Hamme et al. 2006). Table 3.2 lists reported microbial biosurfactants. Furthermore, biosurfactants play role in food formulation, medical (Reddy et al., 2016). During biodegradation, microbes use plastic as sole carbon and energy source. Till date, there is no such study which has reported the plastic carbon incorporation into the macromolecule structure of micro-organisms. Table 4 summarise the major changes observed after microbial invasion and techniques used to determined the changes.

6.1. Functional Groups on the Surfacial Region To examine the type of functional group present on the surface of the polymer FT-IR spectroscopy is commonly used. The changes in the functional groups (wave number/ cm-1) such as carbonyls (1715), esters (1740), vinyls (1650) and double bonds (880) are studied regularly. According to literature survey, the increase or decrease in these functional groups in plastics such as polyethylene and polypropylene are monitored after microbial attack (Volke-Sepulveda et al., 2002; Gilan et al., 2004; Manzur et al., 2004; Artham et al., 2009; Mohan et al., 2016). The surface of polymer which have more oxidised groups are more prone the the biodegradation, since, oxidized gropus fascilitate microbial invasion by increasing hydrophilicity.

6.2. Hydrophilicity The presence of oxidised groups on the surface of the plastic material is directly proportional to the extent of microbial colonisation. This results in an increase of hydrophilicity of the surface as well the biodegradation. Furthe, hydrophobicity is generally determined by the contact angle of the surface with a probe liquid such as water, the more hydrophobic the surface, the higher the contact angle with water (Roy et al., 2008). YoungeDupré equation is an advance approach to study hydrophilicity of surfaces, which allows the estimation of the energy of adhesion to the solid as well as its acid, basic and Van derWaals components (Artham et al., 2009).



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6.3. Crystallinity Chemically, plastic molecules are semi-crystalline in nature, i.e., made up of crystalline as well as amorphous regions. It has been proved from experiments that amorphous parts are dissolved prior to the crystalline parts, which is the main cause of initial increase in crystallinity after microbial activity. Further, it has been hypothesized that small crystals are consumed leaving large crystals in composition (Raghvan and Torma, 1992; Sen and Raut, 2015).

6.4. Surface Morphology The microbial attack has led the changes in the surface structure and texture. The topographical changes include formation of microcolonies on the outer surface and penetration of fungal hyphae (Pramila and Ramesh, 2011a; Awasthi et al., 2017). According to the Das and Kumar (2014) fungal strains e.g., Aspergillus and Fusarium sp. were found to colonise and erode the LDPE surface. The surface of the polymer i.e polyethylene and polypropylene was reported as physically weak, porous and pitted (Otake et al., 1995; Raghuwanshi et al., 2016). Under FE-SEM, the early formation of pits, grooves, cracks, roughening of LDPE and HDPE surface were reported (Ojha et al., 2016).

6.5. Mechanical Properties Modification of mechanical properties is caused by oxidation-induced changes in crystallinity and the average molecular weight.The Universal mechanical testing system (UMTS) determines the mechanical properties of degraded plastic samples. In addition, storage and loss modulus are calculated by rheological measurements. However, breaking load is a common mechanical propery which represents the substrate deterioration.

6.6. Molecular Weight Distribution One of the major drawbacks of plastics is their recalcitrant nature which is assisted by high molecular weight. It was found that there is an increase in molecular weight distribution as a result of microbial attack, since low-molecular weight chains are consumed (Hadad et al., 2005). According to Ohtake et al. (1998) and Fontanella et al. (2010), UV radiation (abiotic factors) or reduction process play key role in affecting the molecular weight rather than microbial activity. To determine the molecular weight


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distributions, two different methods are used. Primarly, size exclusion chromatography at high temperature and secondly, rheological measurements that indirectly correlates to the molecular weight distribution (Bonhomme et al., 2003; Restrepo-Florez et al., 2014).

6.7. Consumption of the Polymer There are two techniques which estimate the assmilation (weight loss) of the poymer namely gravimetric measurements or CO2 evolution.Here, plastics have been used as sole carbon source by micro-organisms and further, this carbon is converted to CO2 during aerobic respiration. Estimation of CO2 is an indirect method of calculating the amount of plastics used. According to Seneviratne et al. (2006), the consumption of plastics and the rate of degradation are determined together. This method can also verify the potential of low molecular weight polyethylene-degrading microbial strains (Yoon et al., 2012). Furthermore, abiotic factors such UV light has enhanced the extent and rate of degradation. Recently, it was found that the ultraviolet (UV)-treated polypropylene started degrading at the faster pace and showed 1.95% gravimetric weight loss compared to thermal treatment (Aravinthan et al., 2016).

6.8. Biomass Accumulation and Lambda-Max (λ-Max) Change The preliminary analysis of biodegradation was checked by the increase in the biomass, which means that the plastic is being used as carbon source by the microbial community and hence increased its optical density (OD600). Moreover, the shifts in λ-max also represent the biodegradation (Anwar et al., 2013). Conclusively, the principal factors that influence the plastic’s bioremediation in the environment are the physical state and chemical structure of the polymer, the length and complexity of polymer chain, mobility, additives added to the polymer and degree of crystallinity (Artakar et al., 2009; Volova et al., 2010; Wang et al., 2012). Generally, polymers with a shorter chain, less crystallinity, hydrophillicity, more oxidized groups and less complex strucure are more prone to biodegradation by microorganisms. Moreover, the environment, in which the polymers are placed, also plays crucial role for their biodegradation. The pH, temperature, moisture salinity, sunlight, stress and the oxygen content are among the most significant environmental factors that must be considered in biodegradation of polymers, since, these abiotic factors influenced the microbial activity (Massardier-Nageotte et al., 2006; Tribedi and Sil, 2013).



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7. DIFFERENT STRATEGIES OF BIOREMEDIATION The microenvironments such as marine water, soil and compost have been extensively used to study plastic biodegradation (Sen and Raut, 2015). As per the literature survey, different strategies have been used to determine the potential mechanisms. These strategies include using single or mixed strains, aerobic or anaerobic microbe, in vitro or in situ condition. The degradation studies of plastics using single microbial strain or mixed compatible strains (consortium) have been reported (Artham et al., 2009; Fontanella et al., 2010; Mumtaz et al., 2010; Lobelle and Cunliffe, 2011; Nowak et al., 2011; Rajandas et al., 2012; Yoon et al., 2012). Using single strain as microbial system to degrade plastic is useful in studying the metabolic reactions/pathways or the effect of different environmental conditions. Whereas, using mixed strains is helpful as the intermediate products formed are degraded with the co-operation of other microbial strains and lead to perform a co-operative process. Moreover, different type of bioremediation occurs depending upon the environmental conditions: oxygen content or type of environmental system. Plastic materials can be degraded aerobically in the presence of oxygen or anaerobically in the absence of oxygen. The process needs to be aerobic or anaerobic, it depends on the oxygen demand of the micro-organisms used.Under aerobic degradation, oxygen is used as terminal electron acceptor and plastic materials are degraded into smaller molecules such CO2 and water. Moreover, under anaerobic degradation, nitrate, sulfate, iron, manganese, and carbon dioxide is used as terminal electron acceptor and plastic materials are degraded into smaller molecules such methane and water (Nayak and Tiwari, 2011). A number of studies have been conducted on the degradation of plastics with different condition such as laboratory condition (in vitro), natural condition (soil, compost, water) (in situ). Here, under natural conditions, soil and compost are rich habitat for plastic degrading microbe; so, most of the studies are preferably executed in these two systems compared to aquatic system (Adamcova et al., 2013; Ardisson et al., 2014; Emadian et al., 2016). Different type of results has been obtained during these studies, which are summarised in Table 5.

8. MEASURES TO ENHANCE THE BIODEGRADABIILITY OF PLASTICS Scientific community over the globe have made efforts to increase the biodegradability of petro-plastics by means of modification at different levels such as


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physical, mechanical or chemical of the polymer. This step has given a new way to fight against this environmental nuisance. Many reports about the modifications of microbe are also surveyed in the literature (Bhardwaj et al., 2012). The following are the parameters reported till date:

8.1. Modifications in the Polymer A variety of blends of plastics with biodegradable synthetic polymers (polylactic acid and polycaprolactum) have been made to increase the biodegradaion of plastics. Moreover, some of the blends like pyrolytic waste plastic oil and its diesel blend served as fuel (Khan et al., 2016). Nanoparticles like superparamagnetic iron oxide nanoparticles (SPION) with size ranging 10.6- 37.8 nm and Nanobarium titanate (NBT) have been reported to accelerate the biodegradation of LDPE (Kapri et al., 2010a, Kapri et al., 2010b). Recently, in one of the study, titania has enhanced the photodegradation (UV) PE films in humid environment (Mehmood et al., 2016). Moreover, the pretreatment of the plastics with heat, UV, radiation and chemicals has also assisted the ease to biodegradability (Awasthi et al., 2017). Further, the addition of additives, antioxidants, UV stabiliser is also important key in enhancing the degradation. Oxo-biodegradable plastics are special class of plastics which are designed to degrade, and contains additives (Sen and Raut, 2015). Moreover, addition of pro-oxidants or starch has been also reported (Zheng et al., 2005; Koutny et al., 2006a).

8.2. Modificaions in the Microbial Agent Micro-organisms which can efficiently degrade plastics should be cultured in higher quantity and must be patented (Goel et al., 2011). Also, the microbes must be induced for the production of biosurfactants or surface active agent which helps the microbes in attachment and degradation of the plastic materials. From past decades, genetic modification is a new technology which has modified the micro-organism according to the interest or desire of a scientist to fascilitate degradation of plastic in eco-friendly manner (Yoon et al., 2012; Joutey et al., 2013). Considerably, E. coli is most popular candidate host used in genetic engineering studies (Hook et al., 2016). The media composition can also be changed to increase the utilisation of polymer by microbial strains.


Table 5. Plastic-degrading studies conducted in different environments Name of plastic Plastic cups

Type of Environment In vitro

Microbe

Condition/period of biodegradability Standard incubation conditions (370C, pH = 7.0, 120 rpm) for 55 days Standard incubation conditions (370C, pH = 7.0, 120 rpm) for 40 days Standard incubation conditions (370C, pH = 7.0, 150 rpm) for 5 days

Criteria of measurement Weight loss

Percent Biodegradability/result 27.4

References

Polythene bags

In vitro

Bacillus, Staphylococcus and Pseudomonas

Weight loss

42.5 20 7.5

Singh et al. (2016)

LDPE (fullerene-60)

In vitro

Microbacterium, Pseudomonas and Bacterium

FT-IR TGA

Sah et al. (2010)

Standard incubation conditions (370C, pH = 7.0, 150 rpm) for 5 days

SEM

Formation of ν C-O frequencies Higher number of decomposition steps and also decrease in the heat of reactions Disruption of surface texture

PVC

In vitro (minimal broth)

PU

In vitro

Microbacterium, Pseudomonas,Bacterium, Acanthopleuro-bacter pedis, Bacillus cereus Enzyme-based (Papain and urease)

Temperature = 370C and pH = 7.0 for 6 months Temperature = 280C (90 days)

GPC ATR-FTIR

Change in molecular weight Change in frequencies

Phua et al. (1987)

LDPE and HDPE

In vitro (Minimal media)

Penicillium oxalicum and Penicillium chrysogenum

Weight loss FE-SEM FT-IR

Ojha et al. (2016)

Lysinibacillus fusiformis

30 days

Bacillus vallismortis Psuedomonas protegens Stenotrophomonas sp.

Temperature = 550C (120 days)

Weight loss FT-IR Zone of clearance method Weight loss

Higher in case of HDPE (>50%) Early formation of pits, grooves, cracks, roughening of surface Bond scission, new bond formation 7.006% of weight-loss was achieved Amount of carbonyls was increased Changes were found at every level

PE

In vitro

LDPE, HDPE

In vitro

Micrococcus luteus and Masoniella sp


Sivasanskari and Vinoth (2014)

Sah et al. (2011)

Mukherjee et al. (2017) Skariyachan et al. (2017)

Name of plastic

Type of Environment

Microbe

Condition/period of biodegradability

and Paenibacillus

Criteria of measurement FTIR SEM NMR Weight loss NMR FTIR, TGA, SEM Weight loss HPLC, NMR, FTIR, TGA, SEM Weight loss ATR-FTIR SEM

Percent Biodegradability/result

References

23% weight loss Changes a all vels were found

Mohan et al. (2016)

12.4% Changes at all levels

Sekhar et al. (2016) Kowalczyk et al. (2016)

Brominated High Impact Polystyrene High Impact Polystyrene

In vitro

Bacillus sp.

30 days

In vitro

Enterobacter sp.

30 days

HDPE

In vitro

Achromobacter xylosoxidans

150 days

Starch plastic

In situ (compost) In situ (soil) In situ (soil)

Municipal yard waste compost site Acanthopleuribacter Pseudomonas Bacillus sp Microbacterium, Pseudomonas andBacterium Lysinibacillus Xylanilyticus and Aspergillus niger Microbacteriu, Pseudomonas, Bacteriu, P. putida and P. aeruginosa Controlled compost composition

Temperature = 65 to 95°C (1 year) Natural 3 months Natural (6 months)

GPC

9% Chemical and structural changes were observed Change in molecular weight distribution

SEM FTIR SEM DSC

Change in surface morphology Change in wave numbers Change in surface morphology Bulk characteristics changed

Johnson et al. (1993) Anwar et al. (2013) Goel et al. (2014)

Natural (126 days)

FT-IR SEM XRD FTIR SEM DSC

Changes at structural, morphological and surfacial level

Esmaeili et al. (2013)

Structural, surfacial changes were found Bulk characteristics changed

Negi et al. (2011)

SEM

Surfacial changes were monitored

Adamconva and Vaverkova (2014)

PVC films PET

LDPE

In situ (soil)

LDPE

In situ (soil)

HDPE

In situ (compost)

Natural (3 months)

15 weeks



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CONCLUSION This chapter discussed the basics of plastics, their types, applications, and properties including the methods of plastic wastemanagement. The major drawback of traditional method includes emission of harmful gases, cost-ineffectiveness, and production and release of secondary pollutants in the environment, which poses threat on the living creatures. Bioremediation, an ecofrirndly waste management system, using an array of micro-organisms including bacteria and fungus along with enzyme and surfactants seems to be the best choice to deal with this environmental problem.Various effects of microbial activity on the polymeric properties and the techniques used to monitor these changes (such as Instron, FT-IR, GPC, NMR, SEM, CO2 emission and gravimetric methods, Lambda max, chromatography techniques) have been discussed in detail. Further, different strategies used by researchers to standardise the bioremediation methods under various conditions for better degradation are reviewed. From these studies it can be conclude that the modification in polymer and micro-organism enhances the rate of degradation. Therefore, bioremediation is emerging as a new technology to fight against the nuisance of plastic waste.

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Chapter 12

THE ROLE OF CYANOBACTERIA IN THE BIOREMEDIATION FOR RESTORING AQUATIC ECOSYSTEMS Vinod Rishi1,*, Ravindra Singh1 and A. K. Awasthi2 1

Department of Biological Sciences, Faculty of Science and Environment, Mahatma Gandhi Chitrakoot GramodayaVishwavidyalay, Chitrakoot, India 2 Department of Botany, Brahmanand Degree College, Kanpur, India

ABSTRACT Indiscriminate utilization of natural resources for infinite requirements of human population in terms of food, fuel, and pharmaceuticals causes a serious and widespread problem of pollution. The major cause of this condition was the industrial revolution that happened in the previous century and now this revolution increased in more advanced phase. Due to industrial development and exponential growth in human population, the entire ecosystem (Air, water and Soil) of the Earth has been contaminated by various harmful pollutants. The aquatic ecosystems are the most affected by such kinds of pollutants because these pollutants are directly or indirectly disposed into the different water bodies. The major pollutants are organic and inorganic fertilizers, pesticides, herbicides, dyes, petrochemicals (crude oil), domestic and industrial effluents.

*

Email: [email protected].


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Vinod Rishi, Ravindra Singh and A. K. Awasthi The treatment of such hazardous pollutants before disposal into the water bodies is of great concern. The conventional methods for the treatment of such harmful contaminants from the aquatic environment are highly expansive and time taking. The biological way for the remediation of harmful contaminants or pollutants is the best approach due to easy to use, cost-effective, eco-friendly and time-saving. Various micro-organisms such as bacteria, fungi, algae, cyanobacteria are natural bioremediation agents. Among them, cyanobacteria offer better option over the other microbes due to their photosynthetic and nitrogen-fixing ability. Cyanobacteria have the capability of detoxifying or removing or destroying a wide variety of pollutants and also produce a wide array of stress tolerant metabolites inside the cell or in the surrounding environment.

Keywords: cyanobacteria, pollution, bioremediation, agrochemicals, petro-chemicals, phenolics, wastewater

INTRODUCTION The Earth is approximately 4.5 billion years old and the life originated about 4 billion years ago in the water. Thus, water is a valuable natural resource and is the essential for all life forms. It constitutes 70%–90% of the total body weight of all living organisms. Rivers, lakes, ponds, pools are the natural repositories of freshwater in the world. Historically, the great civilizations have been developed along the bank of rivers and even today most of the development has taken place in those cities that are situated on or near the rivers. The rivers, lakes, ponds, pools provide water for the industrial, agricultural and domestic use. Unfortunately, these water bodies are being polluted by organic and inorganic contaminants due to inappropriate disposal of domestic sewage, industrial wastes, agricultural runoff and excessive anthropogenic activities. These organic and inorganic contaminants in the aquatic environments pose serious threats to the health and safety of human beings, domestic animals and wildlife [1, 2]. These chemical contaminants present in the aquatic environment may be immobilized and accumulated in the sediments or may be subjected to the transformation and activation processes [3]. As a consequence, enhancement in the levels of such harmful contaminants in the aquatic environments with time and space provide the ideal condition for the growth of various micro-organisms including cyanobacteria. It is of great concern to treat such hazardous pollutants before disposal in the water bodies. The traditional or conventional approaches for the removal of these harmful pollutants from aquatic environment are highly expansive and time taking task. The biological approach for remediation of harmful contaminants or pollutants form water bodies is a move on track due to ease of use, cost-effective, eco-friendly and time saving. Various micro-organisms


The Role of Cyanobacteria in the Bioremediation …

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such as bacteria, fungi, algae, cyanobacteria are natural bioremediation agents. Among them, cyanobacteria offer better alternative due to their photosynthetic and nitrogen fixing ability. Cyanobacteria are most ancient photosynthetic prokaryotes that originated on Earth about 3-4 billion years ago in Pre-Cambrian era. It assumed that they are the first photosynthetic organism [4]. In fact, the origin of photosynthetic organelle in eukaryotes is thought to have possibly arisen by the process of endosymbiosis between a phagotrophic host and a cyanobacterium [5]. Cyanobacteria are morphologically diverse group of gram-negative micro-organisms and exhibit different forms including unicellular, colonial, filamentous (i.e., heterocystous and non-heterocystous), planktonic or benthic [6]. They are cosmopolitan and are found in almost all kinds of habitats where moisture is available. They can also thrive in a wide range of ecological habitats, ranging from marine, freshwater, moist terrestrial environments to extremely adverse environmental conditions [7]. The ubiquity of cyanobacteria in these habitats, besides freshwater and marine environments, and evidence of their presence in numerous stromatolites of the Archaean and Mesozoic era provide the credibility to the theory that they are the first photosynthetic organisms to appear on the Earth [8]. These photosynthetic prokaryotes have contributed to the evolution of oxygen in aquatic and terrestrial environment over the Proterozoic era [9]. The utilization of cyanobacteria in the bioremediation of aquatic environment is a recent phenomenon proposed by Caldwell [10] and preliminary experiments were performed by Oswald, et al. [11]. Since 1980, the application of cyanobacteria in the treatment of waste water has increased and since then a lot of researches have appeared [12, 13, 14]. The cyanobacteria have great potential to uptake or remove external nutrients such as ammonium, nitrate, orthophosphate and heavy metals form polluted environments [15]. The cyanobacteria have two innate properties that make them useful bio-tool for remediation of polluted water. The first is their photosynthetic ability and second is their ability to change in the mode of nutrition according to environmental conditions. These properties provide them an advantage to thrive in those areas and conditions which are non-compatible for other organisms. Thus, cyanobacteria are potential bio-tool for the remediation of urban, agricultural, industrial effluents etc. in terms of solving the problems of eutrophication and metal toxicity in aquatic ecosystems. For a long time, the economic importance of cyanobacteria was mostly restricted to their use as bio-fertilizer in agriculture due to their nitrogen fixing ability. Recently, cyanobacteria have been recognized as most efficient tools for bioremediation of aquatic ecosystem because they can significantly remediate aquatic ecosystem contaminated with organic and inorganic wastes [4, 16-19], heavy metals [4, 20, 21], domestic and industrial effluents [22, 23], crude oil spills [4, 24, 25], naphthalene [26, 27], phenanthrene [28], phenol and catechol [29, 30], dissolved


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inorganic nutrients [31-35], agricultural wastes and pesticides [16, 17, 36], melanoidin [37] and also act as bio control agent [38-42]. The potential of cyanobacteria in the bioremediation of various pollutants from aquatic ecosystem have been reviewed by several researchers [43-49]. The present study aimed to provide an over view the potential role of cyanobacteria in bioremediation and restoration of aquatic ecosystems.

ROLE OF CYANOBACTERIA IN THE REMOVAL OF NUTRIENTS FROM WASTEWATER Nutrients, such as nitrate, nitrite, ammonia, total inorganic and organic phosphates are essential for the growth of cyanobacteria [50-52]. However, these organisms were efficiently utilized the inorganic nitrogen in both free and immobilized conditions. The cyanobacteria have high efficacy in the removal of both organic (BOD and COD), inorganic (heavy metals) as well as physical contaminants (suspended and dissolved solids) from the domestic and industrial effluents relatively in minimum time duration compare to other conventional methods [53]. Anabaena variabilis, A. oryzae and Tolypothrix ceytonica show great efficiency in the removal of fat, oil and grease contamination from wastewater [23]. Some cyanobacteria species like Anabaena variabilis, A. oryzae and Tolypothrix ceytonica also decrease the BOD, COD, TSS and TDS form polluted water [23]. Several studies have been made to investigate the nutrient removal capacity of cyanobacteria from contaminated media and systems [31-35]. Cyanobacteria were also reported as resourceful tools for the assimilation of organic matter from the contaminated systems [4, 18, 19]. Phormidium bohneri have the ability to remove phosphorus contaminants from wastewater [53]. Similarly, nitrates and phosphates ions can be removed from wastewater by P. laminosum [54, 55] and Spirulina platensis [56]. Tartte, et al. [57] demonstrated that the species of Anabeana and Nostoc were effective in the removal of nitrogenous spills from wastewater. Similarly, more than 90% reduction has been recorded in nitrate content by mixed algal population from wastewater [58]. Oscillotaria, Anabeana and Nostoc significantly reduced the Cl- from wastewater [59]. Chandra, et al. [60] concluded that more than 99% reduction in SO4-2 from the tannery effluent can be achieved with Nostoc sp. Some cyanobacteria with nutrient removal capability have been listed in Table 1.



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Table 1. Cyanobacteria involved in the removal of nutrients from wastewater Cyanobacteria Phormidium laminosum Phormidium laminosum Phormidium uncinatum Anabaena CH3 Anabaena doliolum Anacystis nidulans Chlorogloeopsis sp. strain ATCC27193 Coccochloris peniocystis Phormidium subfuscum, Phormidium bohneri, Phormidium tenue, Schizothrix calcicola, Oscillatoria sp. Oscillatoria, Synechococcus, Nodularia, Nostoc and Cyanothece Oscillatoria sp. Oscillatoria, Synechococcus, Nodularia, Nostoc and Cyanothece

Spirulina platensis Nostoc muscorum, Anabaena variabilis, Lyngbya majuscula and Oscillatoria salina

Pollutant/Compounds Nitrate and phosphate Phosphate, Nitrate and Nitrite Nitrate Nitrogenous compounds Nitrate and phosphate Bicarbonates Inorganic carbon

References [54] [162-163] [165] [166] [167] [168] [169]

Bicarbonates Nitrogen and Phosphorus

[170] [171]

Nitrate, Magnesium, Organic and Inorganic Phosphates Biological Oxygen Demand (BOD) and Chemical Oxygen Demand (COD) Biological Oxygen Demand (BOD), Chemical Oxygen Demand (COD), Dissolved Oxygen (DO), NH3, Nitrate, Magnesium, Organic and Inorganic Phosphates Chemical Oxygen Demand (COD) COD, BOD

[123] [172] [123]

[173] [174]

ROLE OF CYANOBACTERIA IN BIOREMEDIATION OF PETROLEUM HYDROCARBONS AND CRUDE OIL FROM POLLUTED AQUATIC ENVIRONMENT Polycyclic aromatic hydrocarbons (PAHs) are composed of two or more complex aromatic rings and they are major components of crude petroleum oils. PAHs are released into aquatic systems as crude oils or oil refinery products [61]. These hydrocarbons are found in huge amount in the water due to their low soluble nature. The US Environmental Protection Agency has identified 16 polycyclic aromatic hydrocarbons (PAHs) as major


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pollutants; 8 of them cause possible carcinogenic effects in human being [62]. Presence of PAHs in the aquatic environments causes toxic and carcinogenic effects in aquatic biota and also in humans via fish food [63]. The cyanobacteria can oxidize aromatic hydrocarbons under photoautotrophic growth conditions. The major cyanobacterial genera which may degrade such hydrocarbons under aerobic conditions are Aphanocapsa, Anabaena, Microcoleus, Nostoc, Oscillatoria and Phormidium [64]. Oscillatoria earli has the ability to assimilate the oil spills from oil refinery effluents [65]. Evidences supporting that the biodegradation ability of cyanobacteria for the elimination of crude oil residues is still very limited. Cultures of Microcoleus chthonoplastes and Phormidium corium were able to degrade n-alkanes [24]. Oscillatoria sp. and Agmenellum quadruplicatum oxidize naphthalene to 1-naphthol [66, 67]. Oscillatoria sp. strain JCM oxidizes biphenyl to 4-hydroxybiphenyl [27] and Agmenellum quadruplicatum metabolizes phenanthrene into trans-9, 10-dihydroxy-9, 10 dihydro-phenanthrene and 1-methoxy-phenanthrene [28]. Several other genera such as Oscillatoria, Phormidium, Lyngbya, Aphanocapsa, Anabaena, Microcystis, Cylindrospermum etc. may also have the capability to degrade crude oil spills and other complex organic compounds such as surfactants and herbicides from the aquatic systems [17, 68-70]. However, some recent studies showed that the association of organotrophic bacteria and cyanobacteria were more effective in degradation of PAHs [71-73]. The consortia of Oscillatoria-Gammaproteobacteria degrade the phenanthrene, dibenzothiophene, pristine and n-octadecane and the degradation rate of these compounds increased due to the presence of cyanobacteria [72]. Similarly, the consortia of Microcoleus chthonoplastes and organotrophic bacteria have the ability to degrade aliphatic heterocyclic organo-sulfur compounds along with alkylated monocyclic and polycyclic aromatic hydrocarbons [73]. Kumar, et al. [74] reported that Phormidium tenue has the ability to degrade some aromatic hydrocarbons like naphthalene and anthracene. Ibraheem [75] investigated the potential of five cyanobacterial strains, Phormidium sp., Nostoc sp., Anabaena sp. Aphanothece conferta, and Synechocystis aquatilis in degradation of crude oil spills from polluted water and concluded that these species efficiently digest n-octadecane and pristine. El-Sheekh and Hamouda [76] demonstrated that Nostoc punctiforme and Spirulina platensis have the ability to grow in different concentrations of crude oil under heterotrophic conditions and analysis revealed the complete removal of crude oil residues such as Decane, Pentacosane, Hexacosane, Octacosane and Nonacosane from the medium by these cyanobacteria. Some petrochemical degrading cyanobacteria have been listed in Table 2.



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Table 2. Degradation of petro-chemical wastes, crude oil and polycyclic aromatic hydrocarbons (PAHs) by cyanobacteria Cyanobacteria Aphanocapsa sp., Synechococcus elongates Oscillatoria sp., other cyanobacteria

Pollutants/Compounds Crude oil

References [175]

Naphthalene

[23, 84, 176181] [182]

Anabaena cylindrica, Phormidium Petroleum hydrocarbons foveolarum, Oscillatotia sp. Oscillatoria salina, Plectonema terebrans, Crude Oil Spills Aphanocapsa sp., Synechococcus sp. Nostoc punctiforme and Spirulina platensis Crude oil Residues (Decane, Pentacosane, Hexacosane, Octacosane and Nonacosane) Phormidium animale Crude oil Phormidium sp., Nostoc sp., Anabaena sp. n-octdecane, and pristine Aphanothece conferta, and Synechocystis aquatilis Oscillatoria agardhii, Anabaena spharica n-Alkanes and Polycyclic Aromatic Hydrocarbons Oscillatoria-Gammaproteobacteria Phenethrene, Dibenzothiophene, Pristine, n-octadecane Agmenellum quaduplicatum PR6 Phenanthrene Oscillatoria sp. strain JCM Naphthalene

[69-71] [76]

[183] [75]

[183] [72] [28] [27]

DEGRADATION OF PHENOLIC COMPOUNDS BY CYANOBACTERIA The phenolic compounds are frequently found in various industrial effluents. The concentrations of these compounds in effluents ranging from 50-2000 mg/liter but the admissible limit of these compounds are 3 mg/liter [30]. Thus, the phenolic compounds in wastewaters may cause the serious problems for the aquatic life along with human beings. Generally, bioremediation of phenol containing effluents treated by bacteria and fungi but some researchers emphasized the potential of cyanobacteria in the remediation of wastewater contaminated with phenolic compounds [77-80]. The cyanobacterium, Phormidium valderianum has been shown to remove these phenolic compounds at the rate of 35 mg/liter during the seven days treatment [30]. Similarly, the marine cyanobacterium Synechococcus PCC7002 successfully metabolized 100 mg/liter phenolic


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compounds in non-photosynthetic conditions [81]. Some cyanobacteria showed sensitivity against the phenolic compounds [82]. The cyanobacteria Nostoc linckia and Oscillatoria rubescens also have the ability to oxidized phenols into catechol [83]. The removal of other phenolics like lignin and tannins are also an interest due to their high structural complexity and abundance. The lignin is an amorphous heterocyclic aromatic polymer and is the major constituent of the cell wall of various plants, while the tannins are water soluble phenolic polymers and precipitate with proteins. The structure of tannins is very similar to lignin. Lignin and tannins are very stable compounds and accumulation of such compounds in the aquatic systems may cause toxic effects on aquatic life. Some reports are available on the biodegradation of lignin by cyanobacteria. The studies revealed that cyanobacteria such as Phormidium ambiguum and Chroococcus minutus have the ability to degrade 73% lignin from wastewater after five days treatment [22]. Similarly, the Oscillatoria sp. and Anabaena sp. showed 89% degradation of lignin [84]. The Anabaena azollae also have the capability to degrade lignin from wastewater systems. Some phenol degrading cyanobacteria have been listed in Table 3. Table 3. Degradation of phenolic compounds by cyanobacteria Cyanobacteria Spirulina platensis Anabaena azollae Phormidium valderianum Phormidium valderianum BDU30501 Oscillatoria sp., Agmenellum quaduplicatum Anabaena PD-1 Nostoc linckia and Oscillatoria rubescens

Pollutants /Compounds Phenolic Compounds Lignin Phenols Phenolic compounds Biphenyl

References [173] [185] [186] [30] [27, 60]

Polychlorinated biphenyls Phenolic and Polycyclic Aromatic Compounds

[187] [188]

ROLE OF CYANOBACTERIA IN DEGRADATION OF AGRO-CHEMICAL WASTES FROM WASTEWATER The excessive utilization of agrochemicals (such as fertilizers and pesticides, herbicides and insecticides) in modern agriculture practices has increased the yield of crops but at the same time, the different aquatic systems have contaminated with these harmful agrochemicals. Several microbes including cyanobacteria have the ability to assimilate or mortify or accumulate such kinds of contaminants from the aquatic



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ecosystems [85, 86]. The organophosphorus insecticides such as monocrotophos and quinalphos were degraded by various cyanobacteria like Synechococcus elongatus, Phormidium tenue and Nostoc linckia within 30 days when these pesticides ranged from 5 to 50 ppm [36, 87]. The cyanobacteria, Anabaena sp. and Aulosira fertilissima have the ability to accumulate DDT, fenitrothion and chlorpyrifos [88]. Thus, Anabaena sp. removed 1568 ppm DDT, 3467 ppm fenitrothion and 6779 ppm chlorpyrifos, while Aulosira fertilissima accumulated 1429 ppm DDT, 6651 ppm fenitrothion and 3971 ppm chlorpyrifos; both the cyanobacteria metabolized DDT to DDD and DDE [88]. Strains of Microcystis aeruginosa, Anabaena cylindrica, A. flos-aquae and A. spiroides have the capability to humiliate the toxic phenylurea herbicide [17]. Among these cyanobacterial strains, M. aeruginosa and Anabaena cylindrica were able to degraded 97% of fluometuron within a day. A gradual increase in herbicide concentration also enhanced the biodegradation capabilities of cyanobacteria, indicating that the biodegradation of fluometuron is species-dependent, and biodegradation capabilities increased with increasing exposure time [17, 88]. Methyl parathion, a toxic organophosphorus insecticide, was disintegrated by Anabaena sp. through a reductive process [89]. Some cyanobacterial isolates such as Oscillatoria sp., Synechococcus sp., Nodularia sp., Nostoc sp., Cyanothece sp., Synechococcus sp., M. aeruginosa and A. cylindrica have the ability to degrade lindane either individually or in combination. Lindane was completely degraded by these species within seven days; but some other intermediate metabolites formed during the lindane catabolism were not detected after the fourth day, thus, it can be assumed that they were able to complete mineralization of these contaminants [16]. Anabaena azotica 118 also have the ability to degrade the lindane at an initial concentration of 0.2 milligram/liter [90]. The Oscillatoria limnetica may also remediate the wastewater contaminated with organophosphorus herbicide glyphosate [91]. Anabaena catenula, Synechocystis aquatilis, Microcystis aeruginosa and Leptolyngbya boryana showed sensitivity against the glyphosate [92]. The cyanobacterial strains such as Synechocystis PCC6803, Anabaena variabilis ATT29413, Spirulina platensis, Arthrospira fusiformis, Nostoc punctiforme, Microcystis aeruginosa, Leptolyngbya boryana showed tolerance to the glyphosate and also capable to degrade it [93-97]. ElNahal, et al. [98] have investigated that cyanobacterial mats have the ability to degrade the acetochlor pesticide from soil or water systems. The organophosphorus pesticide ‘fenamiphos’ can be degraded by diazotrophic filamentous cyanobacteria [85]. Some agro-chemical degrading cyanobacteria have been listed in Table 4.


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Vinod Rishi, Ravindra Singh and A. K. Awasthi Table 4. Degradation of agro-chemical wastes by cyanobacteria

Cyanobacteria Oscillatoria-Gammaproteobacteria

Aulosira fertilissima Anabaena azotica, Anabaena sp. Nostoc linckia, N. muscorum, Oscillatoria animalis, Phormidium foveolarum Oscillatoria limnetica Anabaena PCC7120 Trichodesmium erythraeum Anabaena sp., Leptolyngbyaboryana, Microcystis aeruginosa and Nostocpunctiforme Anabaena variabilis Microcystis aeruginosa Spirulina sp. Anabaena variabilis and Nostoc calcicola Anabaena oryzae and Nostoc muscorum Anabaena fertilissima, Aulosira fertilissima, Westiellopsis prolific Anabaena variabilis, A. cylindrica Synechocystis PUOCCC64 Calothrix brevissima Phormidium sp., Oscillatoria sp. Anabaena sphaerica, Nostoc hatei, Westiellopsis prolifica Oscillatoria sp. Oscillatoria sp. Anabaena sp., A. flos-aquae, Aulosira fertilissima Anabaena inaequalis

Pollutants/Compounds Aliphatic heterocyclic organo-sulfur compounds; monocyclic and polycyclic hydrocarbons DDT Gamma-exachlorocyclohexane Methyl parathion

References [73]

Organophosphorus herbicide and glyphosate Methyl parathion Phosphonate Glyphosate

[91]

Glyphosate Glyphosate Organophosphonate Hexachlorocyclohexane Malathion 2,4-D (Dichlorophenoxyacetic acid)

[190] [97] [95, 96] [192] [193] [194]

2,4-DNP (Dinitrophenol) Anilofos Carbaryl Acetochlor Carbofuran

[195] [196] [197] [98] [198]

Carbendizm Cypermethrin Endosulfan

[199] [200] [201-204]

Isoproturon

[205]

[88] [90, 189] [36, 89]

[89] [190] [97]

ROLE OF CYANOBACTERIA IN DECOLOURIZATION OF TEXTILE DYES The synthetic dyes are irrepressible components of the textile dyeing industries. However, some toxicity tests indicated that most textile dyes are not toxic [100] but, these are persistent in nature resulting the matter of concern. These dyes may cause



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mutagenic and carcinogenic effects in the aquatic organisms [101-104]. In various textile processing units, more than 15% dyes do not attach to the fibers. Thus, these dyes mixed with effluent and these colored effluents may cause severe water pollution problem due to the presence of these toxic colored contents [105]. Some studies showed that the cyanobacteria have the ability to remediate the effluent contaminated with these textile dyes [105-112]. The species of Lyngbya and Oscillatoria have the ability to tarnish the textile effluents containing dyes such as remazol and venyl sulfone and also reduced the level of some inorganic compounds like nitrites, phosphates, ammonia, calcium and magnesium [113-117]. Phormidium sp. was able to degrade several dyes of different chemical classes. The cyanobacterium Phormidium valderianum discolored up to 90% of Acid Red, Acid Red 119 and Direct Black 155 dyes [107]. It has been demonstrated that Phormidium ceylanicum showed 80% degradation of Acid Red 97 and FF Sky Blue dyes after the treatment of 26 days [111]. Silva-Stenico, et al. [118] coined out that the strains of Phormidium, Synechococcus and Leptolyngbya showed more than 50% degradation of the indigo dye along with six other structurally different dyes. The Anabaena sp. was also able to decolorize the Blue Drin dye with more than 80% removal capacity [119]. Recently, Dellamatrice, et al. [120] concluded that the degradation of textile dyes by cyanobacteria was depending on species to species. They also emphasized that Phormidum autumnale UTEX 1580 efficiently decolorized the indigo dye while Anabaena flos-aquae UTCC64 was more effective against the sulfur black dye as well as in sludge decoloration and detoxification. Some textile dyes degrading cyanobacteria have been listed in Table 5.

ROLE OF CYANOBACTERIA IN THE TREATMENT OF DOMESTIC AND INDUSTRIAL EFFLUENTS The cyanobacteria are the versatile biological tool for wastewater treatment because they have extraordinary vitality in sewage water, faster growth rate, tolerating a wide range of temperature, pH and high load of pollutants [121]. The studies have revealed that the several species of Oscillatoria, Nostoc, Synechococcus, Nodularia and Cyanothece have the ability to assimilate, digest or degrade various kinds of pollutants found in domestic sewage and industrial effluents [52, 122-125]. Similarly, some cyanobacteria such as Spirulina, Nostoc and Oscillatoria also have the ability to assimilate certain kinds of pollutants from sewage water [126]. Some studies emphasized that the cyanobacteria were also capable to remediate several kinds of other industrial effluents such as distillery effluents [127], dairy effluents [128], sugar mill effluents [129], starch mill effluents [130] and tannery effluents [131]. Some cyanobacteria have also been listed in Table 5 responsible for the remediation of some industrial effluents.


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Vinod Rishi, Ravindra Singh and A. K. Awasthi Table 5. Cyanobacteria responsible for bioremediation of various industrial effluents and dyes

Cyanobacteria Oscillatoria sp. Phormidium Phormidium sp., Oscillatoria sp., Anabena azollae Oscillatoria, Phormidium, Spirulina, Synechococcus Oscillatoria sp., Lyngbya sp., Synechocystis sp. Microcystis, Anabaena Oscillatoria boryana BDV92181 Oscillatoria, Nostoc, Synechococcus, Nodularia, Cyanothece Anabaena sp. Anabaena sp. HB1017 Oscillatoria-Gammaproteobacteria

Oscillatoria boryana BDU92181 Nostoc muscorum, Anabaena variabilis, Lyngbya majuscula and Oscillatoria salina

Pollutants/Compounds Azo – anilina dyes Black oil Coir

References [106] [206] [207]

Distillery Slops

[208]

Declourization of Distillery Effluents [209] Pulp and paper industry wastewater Pure melanoidins Pharma and textile effluents

[180] [208] [123]

Decolourization of lindane Alkylbenzene Sulfonate Aliphatic heterocyclic organo-sulfur compounds; monocyclic and polycyclic hydrocarbons Melenoidin Textile mill effluent

[210] [68] [73]

[37] [174]

REMOVAL OF HEAVY METALS FROM WASTEWATER BY CYANOBACTERIA The heavy metal term is used for a group of metals or metal-like elements that are directly or indirectly cause serious problems of water pollution. Metals with densities higher than 4 gm/cm3 are considered as heavy metals. There are 53 elements defined as metals among the 90 naturally occurring elements. About 17 heavy metals are essential for the growth and development of living organisms. Among them Fe and Mo are used by organisms as micronutrients, and Zn, Ni, Cu, V, Co, W, Cr are required as trace element but their higher concentrations are toxic for living organisms. Other heavy metals such as As, Hg, Ag, Sb, Cd, Pb and U have no known function in the metabolism of organisms but these metals cause toxic effects in plants and other living organisms along with human beings [132-134]. These heavy metals are accumulated and transferred to higher level via food chain. The various industrial effluents are major sources of these heavy metals and are directly or indirectly disposed into the water bodies. Heavy metals



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generally enter in the human body via various mechanisms like food chains, inhalation, and ingestion. The heavy metals have been used for long time by humans for making metal alloys, pigments for paints, cement, paper, rubber, and other materials. The toxicity of heavy metals in humans is well known. However, once the heavy metals enter the human body through any method, it stimulates the immune system and may cause nausea, anorexia, vomiting, gastrointestinal abnormalities and dermatitis [135, 136]. The toxic heavy metals may also affect the central nervous system [137], change blood composition [138], injuries of lungs [139], kidneys [140], liver [141], and other important organs [142]. The long-term exposures of human population to heavy metals may also cause physical, muscular, and neurological disorders. The removal of heavy metals from the contaminated water can be made by various processes like adsorption, sedimentation, electrochemical degradation, ionic exchange, coagulation and filtration. The bioremediation of heavy metals from wastewater is also mediated by several plants and microorganisms along with cyanobacteria. Recent and past researches emphasized the potential applications of cyanobacteria in the removal of various heavy metals from contaminated aquatic systems [143-146]. Utilization of cyanobacteria in the removal of heavy metals has numerous advantages over the other conventional method due to their photosynthetic, nitrogen fixation, hydrogen metabolism and low nutritional requirements. The capsulated cyanobacteria such as Cyanospora capsulata and Nostoc PCC7936 showed remarkable efficiency in the removal of copper in lab environment [147]. The diazotrophic cyanobacterium Nostoc calcicola also have the ability to uptake the copper from wastewater [148]. Similarly, Cyanothece and Nostoc sp. showed biosorption of copper from contaminated water [149]. El-Enany and Issa [150] made experiments to observe the Zn and Cd uptaking ability of Nostoc linckia and N. rivularis. The detoxification of lead contaminants from wastewater can be achieved by Gloeocapsa and this species showed the removal of lead 232.56 mg/liter from wastewater [151]. The investigations made by Roy, et al. [152] emphasized that the growth of Synechocystis was enhanced up to 300 mg/liter concentration but its growth reduced in the concentration of 400 mg/liter. The Oscillatoria sp. have the capacity to remove the Pb, Cd, Cu, Zn, Co, Cr, Fe and Mn ions from wastewater [153-155]. The Spirulina platensis successfully remove the Cu, Pb, Zn, Ni, Cd and Cr ions from contaminated water [156]. The Spirulina platensis also have remarkable potential in the removal of mercury and lead [157]. The Synechococcus sp. also involves in the removal of Cu, Pb, Cd, Ni and Cr form wastewater [158]. Nostoc calcicola HH12 and Chroococcus sp HH11 also showed the accumulation of the chromium [159]. The cyanobacterium Phormidium sp. can accumulate heavy metals like Cd, Zn, Pb, Ni and Cu [143]. The biosorption of cadmium was also observed in Spirulina sp. [16]. Oscillatoria quadripunctulata can remove 37-50% copper, 20-33% cobalt, 35-100% lead and 32-100% zinc from the sewage and petrochemical industrial effluents [161]. Some species of Anabaena, Oscillatoria, Phormidium, Synechococcus, Synechocystis and Westiellopsis are highly


308


tolerant to the heavy metals [144]. Some cyanobacteria have been listed in Table 6 with heavy metal removal ability from freshwater. Table 6. Heavy metal removal by cyanobacteria from wastewater Cyuanobacteria Anabena doliolum Spirulina Oscillatoria sp. Spirulina platensis Spirulina maxima and Synechocystis sp. Nostoc calcicola Spirulina platensis Oscillatoria angustissima Microcystis sp. Synechococcus sp. Cyanospora capsulata, Nostoc PCC7936 Cyanothece and Nostoc sp. Nostoc linckia, N. rivularis, Tolyputhrix tenuis Nostoc rivularis Gloeocapsa sp. Synechocystis sp. Oscillatoria sp.

Heavy Metals Copper(Cu) and Iron(Fe) Lead (Pb) Cadmium (Cd) Copper (Cu++) Copper(Cu) and Zinc(Zn) Copper (Cu) Cu, Pb, Zn, Ni, Cd, Cr Cu, Zn Ni, Cd Cu, Pb, Cd Cu Cu Zn Cd Pb Pb Pb, Cd, Cu, Zn, Co, Fe, Mn Spirulina platensis Pb, Hg Synechococcus sp. Cu, Pb, Cd, Ni, Cr Nostoc calcicola HH12, Chroococcus sp. HH11 Cr Nostoc muscorum, Anabaena subcylindrica Cu, Co, Mn Spirulina platensis, Aphanothece flocculosa Hg Spirulina platensis Cu, Co, Zn Oscillatoria sp. Cd Spirulina sp. Cr, Cd, Cu Gloeothece magna Cd, Mn Aulosira fertilissima Ni, Cr Nostoc muscorum, Spirulina platensis, Anabaena oryzae Hg2+, Cd2+, Cu2+ and Pb2+ Limnothrix planctonica, Synechococcus leopoldiensis Hg++ and Phormidium limnetica Plectonema boryanum Au Nostoc muscorum, Anabaena variabilis, Lyngbya Mn, Ni, Zn majuscula and Oscillatoria salina Phormidium sp., P. bohneri, P. ambiguum, P. corium Cd, Zn, Pb, Ni, Cu, Cr, Hg,


References [211] [212] [213] [214, 215] [216] [148, 216] [156] [218] [219] [220] [147] [149] [150] [150] [151] [152] [153-155, 221] [157] [158] [159] [222] [223] [224] [225] [160] [226] [227] [228] [20] [229] [174] [143-146]


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CONCLUSION This chapter emphasizes the potential role of cyanobacteria in the bioremediation of various harmful pollutants found in water bodies. Several reports indicate that cyanobacteria have the ability to remove nutrients like nitrogen and phosphorus from wastewater and also decrease the BOD, COD, TDS, and TSS. Beside these the cyanobacteria can also degrade or uptake some other pollutants such as petro-chemicals, agrochemicals, textile dyes, heavy metals and other industrial effluents (Figure 1). Thus, cyanobacteria are potential bio-tool for bioremediation of aquatic ecosystems.

Figure 1. Potential applications of cyanobacteria in the bioremediation of various harmful contaminants from wastewater.


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Chapter 13

THE BIODEGRADATION OF OIL IN SOIL-GROUNDWATER UNDER THE INFLUENCE OF FE(III) Yu. N. Vodyanitskii1, and A. T. Savichev2,3 1

Soil Science Faculty, Lomonosov Moscow State University, Vorob’evy gory, Moscow, Russia 2 Geological Institute, Russian Academy of Science, Moscow, Russia 3 Dokuchaev Soil Science Institute, Moscow, Russia

ABSTRACT Anaerobic оxidation of oil hydrocarbons depends on the type of electron acceptors and decreases in the following sequence: denitrification > Mn4+ reduction > Fe3+ reduction > sulfate reduction > methanogenesis. Usually, not all of these redox reactions develop in contaminated excessively moistened soils and sediments. Fe(III) reduction and methanogenesis are the most common: the latter is manifested near the contamination source, while the former develops in less contaminated areas. Fe reduction hinders the methanogenesis. In places enriched in sulfur and contaminated with oil, Fe-reduction and sulfate reduction are combined, the latter activates Fe-reduction due to the formation of iron sulphides. Concurrently with oil degradation in excessively moistened soils and sediments, the composition of iron compounds changes. In addition, Fe (III) is involved in the anaerobic oxidation of methane, which is formed during the biodegradation of petroleum hydrocarbons. A negative consequence of Fe (III) reduction in oil contaminated groundwater is the activation of arsenic, which in the initial state was fixed with Fe (III) hydroxides.





328

Yu. N. Vodyanitskii and A. T. Savichev

Keywords: Oil hydrocarbons, Fe-reduction, methanogenesis, sulfate reduction, iron hydroxides, mobilization of arsenic

INTRODUCTION Contamination with the oil and its components is one of the most dangerous type of pollution for the environment (Avetov and Shishkonakova, 2010; Illarionov, 2004; Solntseva, 1998; Shoba et al., 2001). In the world, about 400 thousand tons of oil and petrol enters into the soil and water accidentally (Schwarzenbach et al., 2006). More than one thousand tons of oil and more than 9 thousand tons of production water contaminated with oil was spilled at the oil-extracting enterprise in Khanty-Mansi autonomous region, Russia, in 2007 (Kuramshina et al., 2011). Above 49 thousand ha are already oilcontaminated in the Middle Ob region (Vasil’konov, 2009). Many soils have lost their fertility and water reservoirs have lost their fish trade value because of oil spills. Elevated concentration of oil products inhibits or terminates completely the vegetation growth. The total projection coverage of vegetation does not exceed 10% at oil content >15% in the oil-contaminated soils of Western Siberia (Vasil’konov, 2009). Despite the soils’ capacity for self-purification from oil, this process goes slowly without human interference, particularly in the north, where hydromorphic conditions prevail (Pikovskii et al., 2003). The lowest rate of oil biodegradation is registered in permafrost–tundra and taiga regions with widespread peatbogs, in which the reducing conditions are usually established in summer (Pikovskii et al., 2003). In Russia, the microbial oil destruction is far more poorly studied for hydromorphic soils than for automorphic soils (Solntseva, 1998). The existing rehabilitation techniques are aimed at accelerating oil oxidation in the spill area. The measures to decrease the contamination imply loosening soil in order to intensify the aerobic microbes that oxidize hydrocarbons using oxygen from the air. Many efforts are made to inoculate contaminated soils with carbon-oxidizing anaerobic bacteria; however, the results are often unsatisfactory (Zvyagintsev, 1987; Illarionov, 2004). To eliminate the subsurface contamination, the same idea of replacing the anaerobic medium by the aerobic medium is used, and for this purpose, the air is pumped to the contaminated water-saturated layer to provide the long-term soil ventilation (Illarionov, 2004). The positive purification effect is achieved at a high remediation cost. In recent years, attention has been paid to the study and intensification of anaerobic destruction of hydrocarbons in hydromorphic soils and water-saturated sediments. In water-saturated oil-contaminated soils, the microbial metabolism reduces the oxygen content leading to self-inhibition of aerobic oxidation of hydrocarbons (Lovley, 1991, 1997). Therefore, the role of other acceptors of electrons becomes more important in water-saturated layers, with Fe(III) being one of them. Due to the abundance of iron-containing minerals, their contribution to the


The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) 329 anaerobic degradation of oil products is very significant in mineral soils. The situation is different in organic soils, they are often poor in iron. This is manifested not only in their adverse agrochemical properties but also in decelerating redox reactions due to a deficit of Fe(III) as the acceptor of electrons. The chapter considers the following problems: redox reactions as the basis of oil-products degradation in anaerobic environment, the role of Fe in redox-reactions in the anaerobic environment, and the factors controlling the biological reduction of Fe(III), as well as the research on the degradation of particular oil components and the results of long-lasting, activation of arsenic as a result of Fe (III) reduction in the zone of oil pollution of groundwater.

REDOX REACTIONS IN OIL-CONTAMINATED SOIL-GROUNDWATER The natural soil-groundwater usually contains little organic matter and is characterized by a high value of the redox potential. But when water is contaminated with oil, regenerative processes begin to develop in soil-groundwater. Redox reactions in soilgroundwater are characterized by zoning (Baedecker et al., 1993; Bjerg et al., 1995; Christensen et al., 1994, 2000).

Donor–Acceptor Balance in the System Redox reactions require equal contents of donors and acceptors of electrons. Redox reactions are impossible with the deficit in any of them. Organic pollutants are subdivided into two groups depending on their redox properties, i.e., oxidized and reduced. Oil products are classified as reducers, and their degradation is expressed in oxidation, i.e., in electrons passing to acceptors. Oxygen acts as the electron acceptor in the aerobic medium. Under anaerobic conditions, with deficient oxygen, the nitrates and sulfates (occurring in solution), as well as solid-phase compounds (Mn- and Fe(hydr)oxides), accept electrons (Lovley, 1991, 1997). Fe minerals are more significant due to a higher Clarke of Fe than that of Mn. Oil products disturb natural equilibrium by producing excessive donors of electrons. Ox ≠ Red imbalance decelerates the rate of oil degradation, for example, upon iron deficit in sandy soils/deposits or in high moor peatbogs. Introducing Fe(III) as an electron acceptor triggers the redox-reactions, leading to the degradation of oil components. The redox processes of organic matter degradation go only with the decreasing free energy of reaction ΔG < 0. This condition controls the participation of certain iron compounds in the purification of soil or groundwater from organic pollutants. Redox conditions vary throughout the year in the waterlogged soils (Kaurichev and Orlov, 1982). In steppe, forest, and tundra zones they usually change with seasons in the following order. In spring with rising temperature to a level favorable


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for bacterial activity the moisture increases upon snow melting. The biological reduction gets more intensive in soil, and Fe(II) content rises. The reduction of oxidized oil components develops in oil-contaminated sites at this time. In summer, upon soil desiccation and enrichment with oxygen, Fe(II) is oxidized to Fe(III), and the reduced oil components are oxidized. A minor reductive period is often observed in the fall due to rains. Abiotic oxidation processes predominate in winter. Upon year-round observations, soil scientists notice the periodical change in redox processes according to the varying color of medium- and fine-textured hydromorphic soils. For instance, in the Volga– Akhtuba floodplain, gray soils acquire brown color in every winter because of seasonal oxidation of iron (Kozlovskii and Kornblyum, 1972). Seasonal variation in redox conditions provides the degradation of both reduced and oxidized organic pollutants.

The Role of Iron-Reducers Bacteria in the Anaerobic Oxidation of Petroleum Hydrocarbons The anaerobic decomposition of hydrocarbons is of special importance in hydromorphic soils. In the northern taiga and tundra, the aerobic degradation of hydrocarbons is less important, whereas the anaerobic degradation is more important. The oxidative degradation of liquid fuel components is provided by iron-reducing bacteria, specifically, the metal reducers, since the same microorganisms reduce manganese oxides as well. These microbes use Fe(III) and Mn(IV) as electron acceptors. In the anaerobic zone (Cozzarelli et al., 2001), the volatile aromatic hydrocarbons are widespread, such as benzene С6Н6 and its homologues, i.e., toluene С7Н8, ethyl benzene С6Н5–С2Н5 and xylene С6Н4(СН3)2 (these four hydrocarbons are designated by the BTEX abbreviation). For example, toluene is rather quickly oxidized to СО2 in the zone, where Fe(III) reduction with participation of Geobacter metallireducents predominates (Anderson and Lovley, 1999; Anderson et al., 1998: Lovley et al., 1989: Lovley and Lonergan, 1990). With participation of Fe(III) and these bacteria, other hazardous monoaromatic compounds are also oxidized in water-saturated sediments, such as phenol С6Н5ОН (benzene derivate), p-cresol (its homologue), and other substances. Benzene and toluene are readily dissolved in water, leading to contamination of soil- and groundwater with these chemicals. This is the reason for the interest to the problem of hydrocarbon oxidation. Microbiologists revealed the role of Geobacteraceae genus in the oxidation of hydrocarbons (Rooney-Varga et al., 1999). Oxidation of toluene is studied in detail. Two microorganism are involve in this process, i.e., Geobacter metallicreducens and Geobacter grbiciae (Coates et al., 2001; Cornell and Schwertmann, 2003). Benzene is more stable than toluene; it is almost not oxidized even upon long-term boiling with the strong oxidizer KMnO4 solution (Grinberg, 2002). The possibility of benzene oxidation


The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) 331 in the zone of Fe(III) reduction was studied for the oil-contaminated water-saturated light-textured soil (Anderson et al., 1998). The soil sampled from different places was inoculated for 80 days under strictly anaerobic conditions. In the bulk of samples, 14C benzene isotope was not oxidized to 14СО2. This agrees with the earlier obtained data on the benzene stability in non-ameliorated sediments. However, a longer period of microbial population adaptation (which is capable of degrading benzene) may be required. One sample, however, showed oxidation of half 14C in benzene to 14СО2 in 80 days, i.e., anaerobic degradation in situ. The molecular analysis of 16spRNA attested to Geothrix fermentes predomination among prokaryotes everywhere, except the place where half benzene was oxidized and Geobacteraceae genus prevailed, since it is this genus that provides benzene oxidation in the zone of Fe(III) reduction (Anderson et al., 1998). In the oil-contaminated ground flow in Bemeji, Minnesota, the Geobacter group dominates among the Fe (III)- reducers (Amos et al., 2012).

Succession of Electron Acceptors by Oxidability In accordance with chemical thermodynamics, the zone of various reductive processes appear in water-saturated contaminated deposits. The formation of these zones is controlled by various redox reactions (Table 1). In the aerobic environment, where oxygen acts as the electron acceptor, the organic pollutants are oxidized most quickly. In water-saturated layers with deficient oxygen, other compounds, i.e., both water-soluble substances and solid-phase compounds, act as electron acceptors. The Gibb’s potential in electron accepting reactions grows in the following sequence: Denitrification with electron acceptor NO3 (ΔG = - 550kJ/M) < Mn4+ reduction with electron acceptor MnO2 (ΔG = -417 ÷ -383 kJ/M) < Fe3+ reduction with electron acceptor Fe(OH)3 (ΔG = -5 ÷ 96 kJ/M) < sulfate reduction with electron acceptor SO42- (ΔG = 150 kJ/M) < methanogenesis with electron acceptor HCO3 (ΔG = 184 kJ/M) (Bethke et al., 2011). Interval of ΔG values for reactions with participation of solid-phase electron acceptors is explained by their varying thermodynamic properties; this is true for both diverse manganese oxides and ferrihydrite particles with different crystallization degrees. The minus sign for Gibb’s potential designates the spontaneous reaction, whereas the plus sign points to the external source of energy necessary for the reaction (the energy of organic substance oxidation, in this case). The nitrate reduction is most widespread in agricultural landscapes. The sulfate reduction develops in southern regions, where soils are rich in sulfur. Methanogenesis prevails in the areas of high oil contamination. Fe(III) reduction develops in mineral soils at the sites of moderate oil pollution. Zonality is believed to be caused by the competition among different microbial groups (Lovley, 1991). For example, oxygen inhibits the activity of various anaerobic bacteria under the aerobic conditions. Nitrates inhibit Fe(III)



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reduction under over moistening conditions. In its turn, Fe(III) inhibits sulfate reduction and methane formation. In the course of further studies, however, the idea about the competition between bacterial groups for the electron acceptors ad the basis for the formation of zonality was significantly revised (though not completely rejected). The incompatible groups of microorganisms are found in the same zone, e.g., the sample may manifest both a high content of Fe-reducers and methanogens, or sulfate-reducers (Bekins et al., 2001). The fact is that the interacting microorganisms not only compete for the common substratum (oil) but also cooperate in its use (Zavarzin and Kolotilova, 2001). Table 1. Most prominent redox process in landfill leachate plumes (Bjerg et al., 2003) Reaction Methanogenic Sulfate reduction Iron reduction Manganese reduction Denitrification Aerobic respiration CO2 reduction Nitrification Methane oxidation

Process 2CH2O → CH4 + CO2 2CH2O + SO42- + H+ → 2CO2 + HS- + 2H2O CH2O + 4Fe(OH)3 + 8H+ → CO2 + 4Fe2+ + 11H2O CH2O + 2MnO2 + 4H+ → CO2 + 2Mn+ + 3H2O 5CH2O + 4NO3- + 4H+ → CO2 + 2N2 + 7H2O CH2O + O2 → CO2 + H2O HCO3- + H+ + 4H2 → CH4 +3H2O NH4+ + 2O2 → NO3- + 2H+ + H2O CH4 + 2O2 → HCO3- + H+ + H2O

ROLE OF FE(III) IN REDOX REACTIONS Preliminary Remarks Many articles have been devoted to the study of Fe (III)-reduction (Lovely, 1991; Baedeker et al., 1993). Oxides and hydroxides of Fe (III) are widespread in the sediments under aerobic conditions. When creating recovery conditions after oil pollution, these minerals serve as an important source of electron acceptors. Oxides and hydroxides of Fe (III) have a significant redox buffering (Heron et al., 1994a). The process of their recovery in an anaerobic environment is accompanied by the iron-reducers bacteria. Hydrocarbons of oil are the source of energy necessary for Fe (III). Let us analyze the role of Fe(III) in the degradation of oil and its components. Let us consider the degradation of oil components in laboratory. The laboratory analyses permit us to specify the role of separate groups of microorganisms and iron compounds in the degradation of various oil components and to identify the newly formed oil-destruction products. However, laboratory modeling poses problems in the transferring its results to


The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) 333 contaminated natural bodies. The main problem lies in the different time scales of interaction between the organic pollutants and soil or sediment. In the laboratory, upon short-term observation, much higher rate of oil-component degradation may be obtained than upon long-term monitoring in the field. The reason is that organic pollutant degradation develops exponentially, attenuating gradually with time. The time factor consideration is of utmost importance for the prediction of pollutant degradation and transformation and, in particular, for the calculation of remediation parameters aimed at accelerating the environment rehabilitation. The subjective drawbacks of modeling may be also noted. For instance, synthetic iron hydroxides are often used in models, although their natural analogues are more reducible (Tobler et al., 2007). The tendency to use synthetic materials in the laboratory is explained by the difficulty in accurate extraction of particular minerals from soil. The second remark is also caused by subjective reasons. Ferrihydrite is most often used in experiments as an X-ray amorphous hydroxide, whereas feroxyhyte is never used, although the latter is more abundant in soils that ferrihydrite (Vodyanitskii, 2010a). The reason for this narrowness is also clear: the western microbiologists follow the Schwertmann’s paradigm, who obviously underestimated the feroxyhyte abundance in soils.

Possible Deficit of Fe(III) (Hydr)Oxides as Electron Acceptors Iron, being the most widespread element with the variable valence in the Earth’s crust, plays an important role in redox reactions. Fe(III) in (hydr)oxides and ironcontaining silicates is reduced to Fe(II) at the expense of organic substance (including oil) energy and with participation of iron-reducing bacteria. The efficiency of Fe(III) as an electron acceptor depends on many factors, above all, on the sufficient amount of iron in soil. Sandy soils are usually depleted in iron, with its considerable amount included in coarse crystallized particles with low reactivity. Such strong synthetic complexing agents as EDTA (ethylene diamine tetraacetic acid) are successfully applied for activating Fe(III) in oil-contaminated sandy soils and deposits (Lovley et al., 1994). Similarly, when studying contaminated precipitates with rough texture in the laboratory, the introduction of Fe (III) in an accessible form activates the process of reducing organic pollutants (Albrechtsen et al., 1995). Soils of heavier texture contain more iron compounds. Ironcontaining minerals in soils show various grain size and crystallization degree, ranging from poorly crystallized X-ray amorphous ferrihydrite to coarse and well crystallized particles of goethite, hematite, and magnetite (Vodyanitskii, 2010b; Cornell and Schwertmann, 2003). The rate and scale of dissimilative biological reduction of Fe(III), as well as that of decomposition of oil and oil-transformation products, depend on the size and crystallization of these particles.


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Effect of the Grain Size of Iron (Hydr)Oxides on Their Bioreducibility The first investigations of Fe(III) (hydr)oxides performed in the closed reactor testified to much more quick biological reduction of X-ray amorphous ferrihydrite than of the better crystallized hematite and goethite. The conclusion was made that abundance of hematite and goethite in soils is due to their high thermodynamic stability (Cornell and Schwertmann, 2003). However, Roden and Zachara modeling revealed another pattern (Roden and Zachara, 1996). The rate of biological reduction depends directly on the specific surface of minerals. The influence of a specific surface is so high that the rate of hematite reduction standardized by its specific surface was higher than the rate of ferrihydrite reduction. In the experiment with anthraquinone (as an electron shuttle), hematite was reduced much more considerably than goethite or X-ray iron hydroxide as calculated for the low specific surface (Zachara et al., 1998). The velocity of enzyme transfer of electrons depends insignificantly on the thermodynamic properties of iron (hydr)oxides, and it depends substantially on their surface properties. The rate of bacterial reduction of iron (hydr)oxides in soils is evidently controlled by the density of reactive sites on the surface, which is higher for hematite than for goethite or ferrihydrite upon the similar specific surface. The fundamental work (Bethke et al., 2011) elucidated only a partial applicability of chemical thermodynamics for the description of biological redox processes related to participation of iron.

Intensified Reduction of Fe(III) Compounds Due to Natural and Artificial Chelates Natural and artificial chelates influence significantly the reduction rate of iron hydroxides with the participation of prokaryotes (Roden and Urrutia, 2002; Tobler et al., 2007). The action of both natural and artificial chelates is analyzed in the laboratory. Malate, i.e., the salt of dicarboxylic acid (malic acid), was used in one of the experiments. The stability constant of malate to Fe(II) is equal to logK 3.5. An increase in malate concentration from 0.3 to 30 mM due to Fe(II) complexing resulted in a twofold increase in goethite reduction (Urrutia et al., 1999). As proceeds from calculation, complexing Fe(II) with malate involves up to 90% of the whole soluble bivalent iron. Synthetic chelates (NTA, nitrilotriacetate acid; and EDTA, ethylene diamine tetraacetic acid) are able to accelerate reduction of iron (hydr)oxides by complexing Fe(II), as it was earlier established for malate action (Urrutia et al., 1999). In soil- and groundwater, organic ligands accelerate noticeably the biological reduction of crystalline iron (hydr)oxides by binding Fe(II). Phosphates play an opposite role in the reduction of iron hydroxides. An increasing concentration of phosphate from 0.044 to 4.4 mM in malate-containing culture medium reduced the goethite reduction by 40% (Urrutia et al., 1999). Iron phosphate


The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) 335 Fe3(PO4)2 (vivianite) is formed. Vivianite is deposited on the surface of iron oxide particles and iron reducing cells. A high concentration of phosphate inhibited the triggering activity of anthraquinone-containing electron shuttle (AQDS) upon biological reduction of hematite (Zachara et al., 1998). Complexing or deposition of Fe(II) may evidently affect significantly the biological reduction by changing the type of Fe(II) association with (hydr)oxides surface and cells of iron-reducers. To study these effects better, the role of complexing agents Fe(II) in reduction of iron (hydr)oxides and the role of lateral removal of Fe(II) in iron (hydr)oxide reduction in the open systems were investigated.

The Role of Water Regime in Fe(III) Bioreduction The degradation rate of oil hydrocarbons depends in many respects on oil capacity to migrate in a water-saturated ground layer. Oil migrates more quickly in the highpermeable layer composed of gravel-and-sand mixture than in (loamy) clayey aquiclude, where hydrocarbons penetrate only by diffusion (Web and Anderson, 1996). Since gravel and sand layers are poorly ferruginated, the role of iron-reducing bacteria in hydrocarbon degradation is limited there. The water regime also controls another aspect of hydrocarbon degradation, i.e., the necessity to remove their decomposition products from the reaction place. As is known, the accumulating products of biochemical reactions stop the microbial population development (Zavarzin and Kolotilova, 2001). The effect of hydrocarbon degradation depends on the velocity of reaction product removal. No surprise that much attention is paid to the analysis of water regime in contaminated soils and deposits. Under stagnant conditions, newly formed Fe(II) enters the mineral lattice and is fixed on the surface of Fe(III) minerals, preventing further iron reduction. Reducibility of partially bioreduced goethite is restored after washing out its particles with 1M Na-acetate solution with pH 5 (Roden and Zachara, 1996). Three-fourths of adsorbed Fe(II) was removed from the surface of goethite particles. Next, bioreduction of purified goethite was continued. Washing was a significant success, and the Fe(II) content rose 4-fold in three days. Inhibition of the reduction due to the adsorbed Fe(II) was registered for the particles of other iron hydroxides (Roden and Urrutia, 2002). However, the adverse role of adsorbed Fe(II) is not limited by the inactivation of iron hydroxide particles. Fe(II) also neutralizes the action of iron-reducing bacteria. These bacteria bind Fe2+ upon neutral pH due to the complexing effect of negatively charged carboxyl and phosphate groups and due to the extracellular polysaccharides of high capacity (Urrutia et al., 1999). Hence, the surface of iron-reducing cells similar to the surface of iron (hydr)oxides represents geochemically active particles able to bind firmly the biogenic Fe(II). According to calculations, Shewanella algae and Geobacter metallireducens cells manifest the adsorption capacity to Fe(II) (~0.1 mM/g) comparable


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to that of synthetic goethite particles (0.25 mM/g) (Urrutia et al., 1999). Consequently, saturation of iron-reducing cells with bivalent iron decreases the activity of outer membrane reductase, thus inhibiting the reduction of iron (hydr)oxide particles. The recurring inoculation (after the replacement of a one-month-old Shewanella algae culture with fresh cells) stimulated significantly the reduction of synthetic particles of goethite and hematite to raise the Fe(II) production by 2–10 times (Urrutia et al., 1999).

DEGRADATION OF OIL WITH THE PARTICIPATION OF IRON REDUCTION AND METHANOGENESIS Combination of Fe-Reduction and Methanogenesis (Laboratory Experiments) The combination takes place very often. Let us discuss the laboratory experiments performed by Bethke et al., (2011). The experiment on acetate oxidation by goethite αFeOOH in the presence of alluvium containing the natural microbial consortium lasted for 300 days. One-third of the liquid was replaced by fresh water each week. Fe2+ concentration grew in the solution for about a month, then it stabilized due to the slowing goethite reduction; it began to decrease in month four. The finest goethite particles were evidently first to be reduced, while the coarser particles dissolved more slowly. After four months, methane started emitting (its concentration rose 20-fold as compared to the beginning of the experiment), being accompanied by quick oxidation of acetate. At this late period, the rate of methanogenesis was scores of times higher than that of Fereduction (Bethke et al., 2011). There are also data on the combination of two redox processes (Fe reduction and methanogenesis) in the field, i.e., in the oil-contaminated site in Bemidji area (Minnesota, USA) (Bekins et al., 2001; Cozzarelli et al., 2001). The maximal contamination is preserved near the oil spill center, where methanogenesis prevails. Aromatic hydrocarbons (toluene, ethyl benzene, xylene) penetrate deeper; according to the data, in one borehole, hydrocarbons percolated to a depth of 30 cm. The zone of high methane concentration also descended to this depth. However, the area of Fe-reduction spread nearly one meter deeper during this period, to the zone with the lower content of aromatic hydrocarbons (Cozzarelli et al., 2001). Upon long-term monitoring, the oil components were subdivided into heavy nonvolatile and light volatile ones (Bekins et al., 2001). The solid phase of water-saturated contaminated deposits is represented by Mn(IV) and Fe(III) compounds. Aerobic, Mn/Fe reducing and methanogenic bacteria were distinguished among the bacterial mass. Mn(IV) and Fe(III) reducers do not differ, since they are usually represented by the same bacterial groups (Pinevich, 2005). The content of volatile oil components reached its maximum in six


The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) 337 years, afterwards their slow decrease started. It may be due to both the lateral washout and degradation of heavy oil components. The content of volatile hydrocarbons including benzene, toluene, ethyl benzene, and xylene manifests different kinetics. The maximal concentration was reached earlier, in five years; after that the content of volatile hydrocarbons began to decrease rather quickly, with the decrease rate having reached its maximum in 11 years after the accident. A sharp shortage in volatile hydrocarbons in the period 5–11 years is related to their oxidation due to Fe-reduction (Bekins et al., 2001). The reverse dependence between the concentration of volatile hydrocarbons and Fe(III) amount for the same plot was described in paper (Cozzarelli et al., 2001). However, the content of volatile hydrocarbons started rising after 11 years. This late increase in the content of volatile hydrocarbons is time-related to the decree-sing amount of solid-phase Fe(III) in soil and the lower rate of Fe-reduction. The mass of Mn/Fe reducing bacteria grew in two stages. During the first five years, the reduction of instable manganese oxides predominated; after the initial decrease in the Mn(IV) content, the bacterial mass increased substantially up to 12 years due to Fe(III) reduction. Thus, the growth in the biomass of Mn/Fereducers reached its peak after 12 years, and later it decreased due to the reduction of Fe(III) resources. This is a vivid example of inhibition of oil component destruction due to the deficit of electron acceptors. Only after 16 years did the biomass of methanogenic consortium become predominant among the biomass of microorganisms (Bekins et al., 2001). The growth in the methanogenic bacterial mass (even without limitations in carbon content in the hydrocarbons) depends on the other nutrients, above all, phosphorus and nitrogen (most desirable, in the form of nitrate). Therefore, the introduction of mineral fertilizers to contaminated water-saturated soil often turns to be efficient (Illarionov, 2004).

Combination of Fe-Reduction and Methanogenesis in a Contaminated Ground Flow The long-term monitoring of oil spill locations permits tracing of the kinetics of oil oxidation and revealing of factors either accelerating or inhibiting its destruction. With oil ingress, the soil and underlying layers are enriched in monoaromatic hydrocarbons such as benzene С6Н6, toluene С6Н5СН3, ethyl benzene С8Н10, and xylene С6Н4(СН3)2. Vast anaerobic zones are formed in water-saturated layers contaminated with aromatic hydrocarbons (Anderson and Lovley, 1997; Lovley, 1997). Oil components as the source of necessary energy favor the reduction of iron (hydr)oxides (Anderson and Lovley, 1999; Lovley et al., 1994). As a result, technogenic gleying develops in the oilcontaminated mineral soils (Heron et al., 1994b). Polluting hydrocarbons are intensely oxidized during technogenic gleying. Both natural and technogenic gleying is


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accompanied by the changing color of mineral soils from brown to dove-gray. The changing color points to the reduction of brown iron (hydr)oxides and the formation of Fe(II) compounds. It is important that in technogenically gleyed mineral soil, the content of hydrocarbons decreases due to the participation of Fe(III) as the electron acceptor (Satapanajaru et al., 2003). Thus, the development of the gley horizon is a positive morphological sign pointing to the destruction of reduced oil components in mineral soils. Degradation processes are studied in detail in the location of an oil pipeline rupture in 1979 near the town of Bemidji (Minnesota, USA). Tens of boreholes were drilled in the accident area, and the contaminated site became a test plot, where numerous researchers (i.e., hydrochemists, mineralogists, microbiologists, etc.) worked. Owing to its light texture (the soil-forming material is composed of medium and fine sand for 80– 90%), about 400000 l of oil penetrated to a depth of more than 4 m. On the basis of longterm monitoring over the crude oil transformation, the kinetics of aerobic and anaerobic decomposition of its components was established (Baedecker et al., 1993; Essaid et al., 1995). In the aeration zone, the biological oxidation of oil developed due to the aerobic processes; at a lower depth, it went at the expense of anaerobic processes with participation of methanogenesis, as well as Fe and Mn reduction. The authors note extremely high biogeochemical heterogeneity in the contaminated area, as the chemical composition of deposit and microbial population vary sharply at a distance of only a meter (Bekins et al., 1999; Cozzarelli et al., 1999). The actual heterogenic conditions could not be simulated; therefore, the field monitoring acquires particular importance. So, during the 12 years after the accident, 46% of total dissolved oil components have degraded. Only 40% of them were decomposed at the expense of aerobic processes. The other 60% covered anaerobic degradation. The oil component degraded due to Mn reduction (5%), Fe reduction (19%) and methanogenesis (36%) (Essaid et al., 1995). Thus, 1/5 of the entire oil or 1/3 of oil in the anaerobic zone degraded due to iron (hydr)oxides. At the late stage of oil degradation, in 14–18 years after the oil spill, the following changes were registered. During four years, the contaminated zone spread below the groundwater level; the area with the low content of dissolved oxygen (0.0–0.1 mg/l) has broadened significantly, whereas the area with its elevated content of О2 (0.1—0.5 mg/l) has reduced (Cozzarelli et al., 2001). Contamination spread both downward and laterally with the groundwater flow. In the new area with the enrichment with volatile hydrocarbons, their oxidation is intensified due to the reduction of Fe(III)minerals. In four years, the zone with the increased content of water-soluble iron was broadening at a rate of 3 m per year (Cozzarelli et al., 2001). After transformation of Fe(III) to Fe(II), the remaining hydrocarbons may be oxidized to methane in this zone. Thus, the noticeable role of Fe(III) reduction in oxidizing volatile and mobile oil components is evident. In the course of organic substance degradation, new minerals are formed; the biogenic mineral formation covers a separate section in natural microbiology (Zavarzin


The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) 339 and Kolotilova, 2001). In this section, a lot of laboratory studies have been performed. Let us restrict ourselves to the analysis of mineral formation in the field. Participation of iron compounds in redox reaction that led to oil degradation is pronounced in the significant modification of initial iron compounds. At the Bemidji test plot, the content of acid-soluble (0.5 M HCl) iron rose sharply (from 1.7 to 4.6%) in the clay fraction collected from the sediment in the zone of oil contamination (Shelobolina et al., 2004). This is explained by the intensified decomposition of stable iron compounds due to the long-term impact of numerous electrons coming with oil. The magnetic part of clay fraction in soil ( Fe3+ reduction with the electron acceptor Fe(OH)3 > sulfate reduction with electron acceptor SO42+ > methanogenesis with electron acceptor HCO3. As a rule, not all above-listed redox reactions develop in contaminated waterlogged soils and sediments. Denitrification becomes noticeable in agrogenic oil-contaminated soils. Sulfate reduction is manifested in the southern regions rich in natural sulfur or in the places of combined contamination with oil and sulfur. In the oil-spill sites, in addition to iron reduction, other redox processes develop, i.e., methanogenesis and sulfate reduction. In this case, both competition and cooperation between different microbial groups is possible. In non-sulfate media, competition with iron reduction prevents methanogenesis, and the latter starts after the reserve of bioavailable Fe(III) particles is exhausted. In the medium rich in sulfates, iron reducers cooperate with sulfate reducers. Sulfate reducers stimulate iron reduction by formation of Fe sulfides. In contaminated sediments, up to 1/3 of oil may be oxidized at the expense of iron (hydr)oxides in anaerobic zone. As a result of Fe reduction, the iron mineralogy changes in the water-saturated contaminated zone. In addition, Fe (III) participates in the anaerobic oxidation of methane, which is formed during the biodegradation of petroleum hydrocarbons. It is also possible a negative consequence of Fe (III) reduction in oilcontaminated groundwater. There is activation of arsenic, which in the initial state was fixed with Fe (III) hydroxides.


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The Biodegradation of Oil in Soil-Groundwater under the Influence of Fe(III) 349 Roden, E. E. and Urrutia, M. M. (2002). Influence of biogenic Fe(II) on bacterial crystalline Fe(III) oxide reduction. Geomicrobiol. J. 19, 209–251. Rooney-Varga, J. N., Anderson, R. T., Fraga, J. L., Ringelberg, D. and Lovley, D. R. (1999). Microbial communities associated with anaerobic benzene degradation in a petroleum-contaminated aquifer. Appl. Environ. Microbiol. 65, 3056–3063. Satapanajaru, T., Shea, P. J., Comfort, S. D. and Roh, Y. (2003). Green rust and iron oxide formation influences metolachlor dechloration during iron treatment. Environ. Sci. Technol. 37, 5219–5227. Schwarzenbach, R. P., Escher, B. I, Fenner, K, Hofstetter, T. B., Johnson, C. A., von Gunten, U., and Wehrli, B. (2006). The challenge of micropollutants in aquatic systems. Science. 313, 1072–1077. Shelobolina, E. S., Anderson, R. T., Vodyanitskii, Y. N., Sivtsov, A. V., Vuretich, R., and Lovley, D. R. (2004). Importance of clay size minerals for Fe(III) respiration in a petroleum-contaminated aquifer. Geobiology. 2, 67–76. Solntseva, N. P. (1998). Oil Mining and Geochemistry of Natural Landscapes. Moscow State University, Moscow. Shoba, S. A., Trofimov, S. Y., Avetov, N. A., et al., (2001). Ecological standardization of oil concentrations in taiga soils of Western Siberia. International Conference “New Technologies for Purification of Petroleum-Polluted Waters, Soil, Processing and Utilization of Petroleum Wastes”. Moscow. pp. 125–127. Tobler, N. B., Hofstetter, T. B., Straub, K. L., Fontana, D. and Schwarzenbach, R. P. (2007). Iron-mediated microbial oxidation and abiotic reduction of organic contaminants under anoxic conditions. Environ. Sci. Technol. 41, 7765–7772. Urrutia, M. M., Roden, R. E., and Zachara, J. M. (1999). Influence of aqueous and solidphase Fe-complexants on microbial reduction of crystalline Fe(III) oxides. Environ. Sci. Technol. 33, 4022–4028. Vasil’konov, E. S. (2009). Candidate’s Dissertation in Biology. Moscow. Vodyanitskii, Y. N. (2010a). Iron hydroxides in soils: a review of publications. Eur. Soil Sci. 43(11), 1244–1254. Vodyanitskii, Y. N. (2010b). Iron Compounds and Their Role in Soil Protection. Dokuchaev Soil Science Institute. Moscow. Vodyanitskii, Y. N., Avetov, N. A., Trofimov, S. Y., Savichev, A. T. and Shishkonakova, E. A. (2013). Influence of oil and stratal water contamination on the ash com-position of oligotrophic peat soils in the oil-production area (the Ob’ region). Eur. Soil Sci. 46(10), 1032–1041. Wasserman, G., Liu, X., LoIacono, N., Kline, J., et al., (2014). A cross-sectional study of well water arsenic and child IQ in Maine schoolchildren. Environ. Health. 13 (1), 2340.


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Chapter 14

BIOSENSORS IN BIOREMEDIATION Ghazala Yunus1,* and Mohammed Kuddus2 1

Department of Basic Sciences, University of Hail, Hail, KSA 2 Department of Biochemistry, College of Medicine University of Hail, Hail, KSA

ABSTRACT Biosensors are analytical devices that can analyze and detect the molecules of our interest, by using biological and electrical components. Bioremediation is a process which removes pollutants from the environment by using microorganisms and/or their enzymes. Biosensors are also used for the detection and monitoring of a nitrogen, phosphorous, dissolved oxygen and different pollutants along with various toxic compounds present in the environment. Bioremediation abilities of the microbial population can also be analyzed by using genomic tools, that help in planning of efficient remediation strategies. In this chapter, we discussed about various biosensors and their applications in the bioremediation process.

Keywords: biosensors, enzymes, bioremediation, pollutants, reporter gene

INTRODUCTION In the last few decades, many techniques based on innovative methods have been established to eliminate contaminants from the environment [1]. Several conventional methods have been found to be unsuccessful due to impermeability, different subsurface *

Corresponding Author address: Email: [email protected].


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Ghazala Yunus and Mohammed Kuddus

situations, and pollutants mixture. To overcome these limits of conventional methods, researchers have concentrated on the bioremediation techniques, which uses microbes to degrade various contaminants present in the environment. Bioremediation tactics are potential substitute due to their security, speediness, less price, and great efficacy in the removing of contaminants from the environment. The principle of bioremediation is the degradation of hazardous pollutants by the microbes; that yield energy to grow and reproduce microorganisms.. Bioremediation may also occurs naturally due to presence of essential materials, required for bacterial growth, at the polluted sites. For this purpose, the bioremediation techniques needs an engineered biological system to accelerate the biodegradation of organic compounds [2, 3]. Bioremediation is an attractive technique that utilize biological organisms (usually bacteria, fungi, actinomycetes, cyanobacteria and plants) to decrease or remove toxic pollutants form the polluted sites. The methods of pollutant elimination relies on the type of the pollutants, which may comprise: dyes, heavy metals, hydrocarbons, nuclear wastes, plastics etc. Bioremediation techniques are cost effective, environmental friendly and can be used over large areas. These techniques can be classified as ex-situ or in-situ according to the sites of application [4, 5]. In in-situ bioremediation, there is stimulation of microbial activity at the contaminated site itself. However, in ex-situ bioremediation, there is restoration of contaminated materials by land-farming and composting methods. Bioremediation approaches are a potential substitute for the removal of organic and/or inorganic pollutants from the environment. In bioremediation, microbes are able to produce energy by degrading hazardous pollutants; and the harvested energy are utilised by the microbes for their growth and reproduction [6]. Biosensors are the devices that utilize specific biological compounds to yield a detectable signal. There are many bioactive compounds of the cell such as enzymes along with the whole cell that may be used in the biosensors. In the analysis of clinical samples, biosensors based on luminescence system are used.. In this chapter, we will specially discuss about the biosensors that are designed by using enzyme expression systems and specific promoters, commonly known as molecular biosensors.

MOLECULAR BIOSENSORS In the last decade, one of the most remarkable field to use biosensors was environmental bioremediation. These biosensors are prepared by fusion of pollutantresponsive promoters with a reporter gene coding for a protein that can be easily quantified [7]. Biosensors are very important technique for carrying out environmental safety assessments on the contaminated sites and for observing the efficiency of a remediation policy, due to their capability to detect only bio-available fractions of the pollutants. Whole-cell microbial biosensors are one of the novel molecular tools in the environmental monitoring [8-10]. Microbes, due to their less price and range of suitable


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pH and temperatureadaptability, have been broadly used as a bio-sensing element in the manufacturing of biosensors [11]. Figure 1 described block diagram of a biosensor for the environmental monitoring. Biosensors have various advantages over traditional detection methods because it provides more accurate response on the bioavailibility and toxicity of a sampleBiosensors are low-priced and simple way of detecting pollutants [12]. Table 1 shows some application of biosensors in bioremediation.

Figure 1. Block diagram of whole cell biosensor.

Table 1. Application of biosensors in bioremediation Host chassis

Target

Promoter-Reporter

E. coli XL1-blue

Zinc and Copper Lead Zinc Zinc Copper

E. coli TOP10 E. coli XL1-blue E. coli XL1-blue E. coli XL1-blue

Reference

zraP-gfp, cusC-gfp

Detection limit (Mm) 0.05–1

zraP-gfp-ompC zraP-gfp-HydG zraP-gfp cusC-gfp-CusR

0.3–1 0.01–1 0.1–1 0.004–1

[30] [31] [32] [33]

[29]

Reporter for Biosensors The function of a reporter gene is to generate a biological signal. Reporter gene has specific features which is used in the biosensors. The reporter gene commonly codes an enzyme that catalyze a specific reaction. The sensitivity and detection limit of a biosensor is determined by different reporter genes. There are several reporter genes that frequently used in biosensors such as gfp, lacZ, lucFF, luxAB, and luxCDABE.

Bacterial Luciferase Reporter System The bioluminescence is a process of light emission by enzymatic response of luciferase enzyme activity [13]. The experimental studies shows that the emission of light is achieved through the lux operon. The lux operon (lux CDABE) has been cloned from Vibrio fischeri, Photorhadbus luminescens. V. fischeri and P. luminescens. The difference between the lux operon systems of these microbes are the thermal ability of their lux systems. The luciferase enzyme from V. fischeri is stable up to 30°C, above


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which the enzyme starts losing its activity while enzyme from Photorhadbus is stable up to 42°C [14]. The action of these systems could be directly examined by a photomultiplier tube for signal analysis. The Vibrio fischeri luxCDABE gene cassette is widely used because it is a very sensitive reporter and does not needs any exogenous substrate [40, 41]. Eukaryotic luciferases have more light production and apply a lower metabolic load on the host cells than bacterial luciferases; but they need exogenous luciferin which makes difficulties for the assays [42]. The bioluminescence reporters have quick response time and high sensitivity in comparison to green fluorescence proteins.

Green Fluorescent Protein Reporter System The green fluorescent protein (GFP) can also be used as a reporter system. The green fluorescent protein of the jellyfish Aequorea victoria absorbs light (ultraviolet or blue light) with maximum wavelength of 395 nm, and fluoresces with emission (bright green light) at maximum wavelength of 510 nm. This protein allows jellyfish to fluoresce due to transfer of energy from the Ca2+ activated photoprotein aequorin [15]. However, GFP fluorescence has a disadvantage for the biosensor applications as it can be detected long after reporter gene induction has been ceased and even after the cell death [41], but laserinduced fluorescence confocal spectroscopy techniques can be as sensitive as bioluminescent reporters [43]. Fluorescent reporter proteins have chance for applying multiple fluorescent reporters in the same cell that fluoresce at various wavelengths (eg., green fluorescent protein and red fluorescent protein), so there is a chance to use dual fluorescence reporting where one color is repressed by an analyte and another color is induced [44]. This approach increases the sensitivity of the assay but commonly not used because luminescence and fluorescence biosensors are not suitable for the use in soil particles or pigments that can attenuate light transmission.

Promoters as Biosensors Promoters are the 5′-flanking sequences in a genetic material or operon that have significant role in DNA transcription. Promoters are able to give response against target molecules. The target molecules can also interact itself with promoters through a receptor system. Therefore, in bioremediation designing of a biosensor and selection of promoter is based on the target molecules. Accordingly, it may be concluded that the promoter is the actually sensing component of a biosensor. These type of biosensors have three main components viz. plasmid, which helps in the maintenance of the biosensor in the particular host; second is the promoter, which is the actual sensing component of the biosensor; and the third one is reporter system, which could be lux operon, GFP or any other signal producing molecules. The selection of promoter for the construction of a biosensor depends on the target



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molecule that is to be monitored. The sensitivity and specificity are the main factors for choosing a promoter. Promoters often react with group of compounds rather than to a particular compound, and may also behave in a different way in different microbes [16]. There are various types of promoters available for biosensors to be used in bioremediation.

Stress Promoters There is some stress protein formation when the bacteria faces an adverse environmental conditions such as nutrients hunger, exposure to contaminations, heavy metals etc. These type of proteins are called Heat Shock Proteins (HSP), and help bacteria to survive in the adverse conditions [8]. uspA, grpE or dnaK are the promoters from E. coli and sub-cloned in the lux expression vector. When E. coli is used as the host, these promoters have been established for their non-specific reaction to several stresses. The response time for this biosensor is less than 5 minutes. Nutrients Monitoring The bioremediation technique needs a specific ratio of a carbon, nitrogen and phosphorus for the biological degradation. The biosensors uses promoter glnA for nitrogen and phoA for phosphorus that are designed for the nutrient monitoring; similarly Pnah promoter of lux system (in case of Pseudomonas) is used for the recognition of carbon level of a particular region. Applegate et al., developed a biosensor for the monitoring of benzene, toluene, ethyl benzene, and/or xylene in the polluted samples [17]. Metal Ions and Physical Parameter Monitoring The biosensors are designed for the monitoring of heavy metals such as Cu2+ and Cd2+. In the case of E. coli, the universal stress protein A is encoded by uspA gene [8]. The promoter uspA can be switched on non-specifically by the conditions that limits cell growth. This includes nutrient starvation and exposure to toxic chemicals [18]. These biosensors are nonspecific and used only under distinct situations. However, biosensor for monitoring mercury in the environmental sample can target only mercury present in either inorganic or organic form. These biosensors uses particular promoters derived from mer operon, characterized for decontamination of both the forms of Hg(II) with a sensitivity of 1 mM concentration [19, 20]. The biosensors can also be used for monitoring of physical parameters such as pH, dissolved oxygen and temperature that affect growth of microorganisms [21, 22].


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PREPARATION OF BIOSENSORS Biosensors are constructed in two steps. The first step is the designing of the biosensor and second is its application as a monitoring device. In the manufacturing of molecular biosensors, different genetic engineering process are used that needs prior knowledge of the gene and its expresion system. A series of genetic events in the cells can produce typical physiological response by switching on/off of a gene operon. Therefore, the promoter of this gene could be considered as a very significant factor for the designing of a biosensor. After developing a biosensor, next step is the signal production and its quantification. For the signal production, a reporter system is needed which has multiple cloning sites at the 5′ end to sub-clone the promoter. In the designing of multiple cloning sites, the precaution should be taken that it should not disturb the coding sequence for the reporter protein. In response to target molecule, a signal is generated and the level of signal depends on the interaction of target molecule and the promoter [15, 23, 24]. The last step is quantification of signal that is very critical step. The generated signal could be a protein or sometimes it acts as a functional protein to find out the biochemical reaction. For the purpose of signal quantification, luciferase expression system, fiber optic device and GFP fluorescence can be used and may be connected to a data processing unit for signal monitoring.

APPLICATION OF BIOSENSORS Biosensors for Environmental Monitoring Whole-cell biosensors offer very good tools for the monitoring of environmental contaminations and toxicity. They have the ability of in-situ and on-line monitoring but facing problem to maintain constant sensing and variability for long periods [54]. To cope these problems, many other techniques have been applied including freeze drying, vacuum drying, continuous cultivation, and immobilization [45-47]. Encapsulating whole-cell biosensors with polymers are very beneficial for the detection of environmental contaminants [49-51]. In encapsulated whole cell biosensors, cells are protected from the contaminated substances in their environment and have increased plasmid stability [52]. In comparison to conventional detection techniques for monitoring environmental impurities, whole-cell biosensors provide many advantages such as simple to use, low cost, more specific and high sensitivity [53]. Since the biosensors detects only bioavailable fraction of the compounds, they provide more accurate response on the toxicity of a sample. Bioavailability is a very significant factor in the bioremediation. If substances are bioavailable, they will be biodegradable.



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Biosensors for Monitoring the Toxicity of BTEX Compounds Benzene, toluene, ethylbenzene and xylene are together stated as BTEX compounds. These compounds are present in the environment due to their extensive industrial and domestic applications [38]. BTEX compounds have carcinogenic and mutagenic properties even at low concentration, hence become a focus of concern. Biosensors can monitor changes in acute toxicity and biodegradation of BTEX compounds in the soil [39]. Luminescent biosensors are also able to detect toxicity and bioavailability in water extracts from BTEX-impacted soils as degradation proceeded. The biosensors can also be used to assess the performance of different treatments applied for natural reduction of BTEX compounds from the polluted soils.

Biosensors in On-Line Monitoring The biosensors can be used for on-line monitoring also. This can be achieved by using lux as a reporter system. The microbes can be used as immobilized cells or in suspended growth culture, either in the batch or continuous mode, which is crucial for sensitivity of the reaction [25]. The immobilized cells produce light signal through the lux reporter system [26]. Ikariyama and coworkers reported a specific promoter designed to monitor a benzene and its derivatives [27]. To measure the noxiousness of wastewater, Gu et al., [28] developed a two-stage mini bioreactor. First reactor uses the promoter xylS under the control of xylR, which is a regulatory protein, and the second reactor behaves as a reaction pitcher for light signal quantification.

In Situ-Online Environmental Monitoring The biosensors can also be used for in situ-online monitoring of environmental pollutants. In this process, these biosensors are using reporter enzymes whose activity can be monitored by the electrochemical transducers [33]. The electrochemical transducers are very sensitive, reproducible with small size analyzer and throwaway electrodes. These tools are able to measure various samples simultaneously. In addition, as these measurements are not optical, it can be used in the crude or turbid solutions such as for monitoring of contaminants in water systems or even in soils [36]. Electrochemical systems are very useful for monitoring gene expression in microbes, yeasts and even in mammalian tissue cultures also [34, 35]. For on-line and on-site monitoring, a small size and portable device is needed that may be obtained by reporter gene systems at nanoscale level. Popovtzer et al., [37] established an integrated electrochemical nanochip for the monitoring of contaminants in the water. This nanochip is prepared on small electrodes and the data are collected on a portable computer.


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TCRS-Based Heavy Metal Biosensors in Bioremediation Two-component regulatory systems (TCRS) based biosensors have two components viz. a histidine kinase (HK) and a response regulator (RR). Histidine kinase senses stimuli from the external environment and autophosphorylates conserved histidine residues in the kinase itself. Response regulator is regulated by histidine kinase, which phosphorylates aspartate residues on the RR [6]. The phosphorylated RR bind with promoters and generates output, thus activates gene expression [57]. These type of biosensors have broad applications in the recognition and degradation of environmental contaminants and toxic compounds. TCRS can determine a wide range of environmental signals including light, oxygen, pH, temperature, heavy metals and organic contaminants [58]. Various types of TCRS-based biosensors are reported in the literature for the environmental monitoring but for heavy metals detection there are very few biosensors due to toxic heavy metal cations. Many research groups developed TCRS based heavy metal ion biosensors that are useful in bioremediation of environmental contaminants [29-33].

FUTURE PROSPECTS Molecular biosensors have high sensitivity and specificity that provides necessary features in the bioremediation. The future prospects of biosensors should be concentrated on biological components and recombinant plasmids for the preparation of pollutants specific remediation techniques. For the construction of real time biosensors, more research are needed for the selection of suitable host. The proteins have an important role in biosensors that is also a growing area of research. Protein has been covalently bind with chromophore and they are very useful in various analysis [55]. Fluorophore molecules, that play important role in signal transmission and enhancement, could also support development of this technology [56]. The basic research in this area includes designing of new reporter systems. Thus, instead of applying the traditional reporter systems such as lux or GFP, a functional proteins such as an enzyme can be used as a reporter system. The enzymes can be associated with a fluorophore molecule that produces light signal and can be used as signal amplification system. The important limitation, for the future development of biosensors in bioremediation, is the accessibility of solid promoters that respond only to relevant stimuli. To avoid this problem, more knowledge on gene regulatory networks in microbes is required. An another option may be synthesis of ‘super promoters’ based on consensus sequences obtained from comparative studies of various promoters in known regulatory networks [39]. Single-cell detection by flow cytometry offers more accurate data that extend possibility to use biosensors directly in in situ bioremediation. The single-cell detection is now restricted by the necessity of costly equipment, but the future development of ‘low budget’ flow



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cytometers will surely overcome this problem of fluorescent biosensors. Also, the singlecell detection by biosensors have various advantages but this approach is restricted by the shortage of fluorescent reporter genes. Future research is required in the field of TCRS based biosensor to monitor environmental contaminants and heavy metal ions. In the future, new advances in biosensors combined with new equipments will allow the improvement of fully automatic on-line biosensor systems. Such biosensors will be definitely useful in industrial food production, drinking water, sewage systems and in several other areas.

CONCLUSION In this chapter, different types of biosensors used in bioremediation are discussed. In the last few decades, the advancement in micro-electromechanical systems (MEMS) provides new class of biosensors. There are many problems in the manufacturing of biosensors for which some techniques are reported, but still more research are required in order to discover better alternatives such as (1) contamination of bio-elements and chemicals used in the biosensors are needed to prevent from leaking, (2) detection limit and sensitivity should be large enough, and (3) research efforts should be focused on the lesser price biosensors. At present, more rapid, reliable, precise, handy and lesser price biosensors are required.

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ABOUT THE EDITOR Dr. Mohammed Kuddus Associate Professor and Head Department of Biochemistry, University of Hail, Hail, Kingdom of Saudi Arabia Email: [email protected]

Dr. Mohammed Kuddus has completed his PhD in Enzyme Biotechnology from SHUATS, Allahabad, India. After completing doctoral degree, he served at Integral University, India; and at present working as an Associate Professor and Chairman of Biochemistry department at University of Hail, Saudi Arabia. Dr. Kuddus’s main research area includes enzymology and microbial biotechnology. He has published more than 50 research articles in reputed international journals along with 12 book chapters; and presented more than 35 abstracts at national/international conferences/symposia. He has been serving as an Editorial Board Member and Reviewer of various international reputed journals. He has been awarded Young Scientist Projects from Department of Science and Technology, India and International Foundation for Science, Sweden.



INDEX A acclimation, 74, 89, 145 accumulators, 80, 81, 104 actinomycetes, vii, x, 66, 113, 114, 115, 117, 119, 120, 121, 124, 126, 128, 138, 246, 352 activated carbon, 151, 171, 175 adsorption, 2, 58, 59, 62, 66, 82, 85, 89, 103, 114, 142, 152, 153, 154, 155, 157, 158, 159, 160, 161, 162, 163, 165, 166, 167, 173, 174, 178, 183, 241, 307, 318, 335, 361 aeration, 141, 211, 215, 220, 233, 338 aggregation, 49, 56, 57, 64, 65 agrochemicals, ix, 19, 28, 296, 302, 309, 324 agro-waste, 20 air pollution, 75, 248, 320 algal biomass, 53, 72, 76, 77, 82, 89, 92, 94 antioxidative defense, 85 aquaculture, 76, 250 azo dye, x, 7, 33, 35, 39, 40, 41, 42, 45, 47, 146, 169, 172, 175, 176, 177, 178, 184, 317

B bacteria, x, 3, 4, 5, 7, 9, 11, 13, 18, 23, 28, 29, 30, 31, 35, 38, 39, 40, 42, 44, 45, 53, 57, 60, 61, 63, 67, 68, 69, 75, 115, 117, 119, 125, 128, 133, 138, 149, 153, 166, 176, 185, 186, 187, 188, 189, 191, 192, 193, 194, 195, 196, 197, 199, 201, 202, 203, 204, 205, 206, 207, 209, 210, 219, 233, 236, 237, 241, 246, 250, 260, 267, 269, 271, 281, 282, 284, 289, 291, 292, 296, 297, 300, 301, 312, 315, 324,

328, 330, 331, 332, 333, 335, 336, 337, 340, 342, 348, 352, 355, 360, 361 bacterial luciferase, 353, 354 benzene, toluene, xylene, ethylbenzene (BTEX), 212, 330, 343, 346, 357, 360, 362 bioaccumulation, 2, 3, 5, 11, 15, 41, 43, 82, 87, 97, 108, 112, 178, 192, 206, 242, 326 bioaugmentation, 2, 6, 195, 211, 213, 233, 242 bioconcentration, 72, 77, 78, 81, 82, 85, 87, 89, 93, 103, 316 bioconcentration factor, 81, 82, 85, 87, 89 bioconversion, 21, 79, 257 biofertilizers, 187, 188, 197, 250, 314 biofilm, 25, 49, 50, 51, 52, 53, 54, 56, 61, 64, 66, 131, 137, 139, 141, 196, 285, 287, 290, 291, 292, 315 biofuel, 20, 27, 43, 77 bioindication, 73, 74, 76, 77, 81, 101 biological oxygen demand (BOD), 35, 134, 248, 298, 299, 309 biological treatment, x, 35, 73, 147, 149, 169, 176, 182, 282 biomedical waste, 19 biopile, 214, 215, 216, 222, 224, 225, 229, 230, 231, 232 bioplastics, 28, 38, 39, 44, 284 bioreactor systems, x, 113, 114, 129, 130, 131, 135 bioremediation of polluted soils, 185 biosensors, viii, xi, 351, 352, 353, 354, 355, 356, 357, 358, 359, 360, 362, 363 biosorption, vii, x, 2, 3, 5, 11, 13, 14, 16, 17, 18, 36, 47, 51, 64, 68, 69, 78, 80, 82, 92, 101, 104, 106, 108, 112, 151, 152, 153, 154, 155, 156, 158, 159,


Index

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160, 161, 162, 163, 164, 165, 166, 167, 178, 179, 180, 181, 182, 183, 184, 192, 242, 307, 321, 325, 326 biostimulation, 2, 6, 211, 213, 233, 241 biosurfactants, 11, 18, 24, 145, 192, 259, 271, 272, 274, 278, 291, 292 biotechnology, xi, 14, 19, 38, 39, 40, 42, 43, 44, 45, 46, 62, 102, 103, 104, 107, 139, 140, 141, 142, 143, 144, 145, 146, 148, 149, 151, 163, 164, 165, 167, 185, 202, 203, 205, 206, 208, 209, 210, 237, 239, 242, 249, 252, 253, 254, 257, 285, 291, 318, 360,362, 365

C cadmium, vii, x, 15, 17, 66, 71, 73, 76, 78, 81, 83, 86, 87, 88, 89, 91, 92, 93, 94, 95, 96, 97, 99, 102, 103, 104, 105, 106, 107, 108, 109, 111, 112, 153, 155, 166, 167, 186, 195, 203, 207, 208, 210, 307, 308, 320, 321, 325, 326, 362 chelators, 93, 192, 201 chemical oxygen demand (COD), 35, 131, 134, 135, 136, 137, 142, 172, 175, 298, 299, 309, 317 chitin, x, 58, 154, 161, 166, 239, 240, 243, 244, 245, 246, 247, 248, 249, 250, 252, 253, 254, 255, 256, 257, 258 chitinase(s), x, 24, 239, 240, 246, 247, 249, 250, 251, 252, 253, 254, 255, 256, 257, 258 chitinolytic enzymes, 247, 254, 255, 257 chitinous wastes, 239 chito-oligosaccharides, 246, 247, 252 chlorophenols, 114, 117, 118, 129, 139, 140, 143, 145, 149, 361 chlorophyll synthesis, 84 chromium, vii, ix, 14, 17, 58, 63, 71, 73, 78, 79, 81, 83, 84, 85, 89, 91, 92, 93, 94, 95, 96, 98, 99, 100, 101, 102, 104, 105, 106, 108, 109, 110, 111, 112, 153, 162, 167, 180, 186, 193, 202, 204, 205, 207, 208, 209, 210, 307, 319, 320, 326, 344 coagulation, 34, 171, 172, 184, 307 co-contaminated water, 86 complexation, 2, 78, 79, 81, 85, 109, 154, 161, 192, 241 contaminant(s), ix, 1, 2, 3, 6, 8, 11, 12, 13, 32, 49, 56, 58, 72, 73, 80, 81, 92, 105, 110, 155, 161, 169, 175, 197, 201, 219, 240, 241, 242, 296, 298, 302, 307, 309, 346, 348, 349, 351, 356, 357, 358, 359, 360

contaminated site, ix, 1, 2, 32, 330, 336, 338, 339, 352 cooperative interactions, 93 cross-tolerance, 94, 96 cyanobacteria, viii, x, 44, 47, 53, 188, 295, 296, 297, 298, 299, 300, 301, 302, 304, 305, 306, 307, 308, 309, 310, 311, 312, 313, 314, 315, 316, 317, 318, 319, 320, 321, 322, 323, 324, 325, 326, 352 cytochrome P450, 29, 30, 119

D decolorization, 7, 34, 35, 36, 39, 44, 45, 47, 146, 169, 172, 177, 178, 179, 180, 181, 182, 183, 184, 315, 317, 318 depollution, 58, 59, 62 detoxification, vii, ix, 11, 13, 18, 19, 31, 41, 42, 45, 72, 75, 79, 83, 84, 96, 104, 105, 106, 147, 148, 195, 198, 241, 305, 307, 319 dibenzothiophene, 4, 15, 211, 218, 232, 300, 301 dioxygenases, 7, 8, 9, 11, 29, 30, 118, 121, 122, 124, 126, 139, 140, 141 dye removal technologies, vii, 169, 171 dyes, ix, 19, 32, 33, 34, 35, 36, 38, 39, 41, 44, 45, 46, 169, 170, 171, 172, 173, 174, 175, 176, 177, 178, 179, 180, 181, 182, 183, 184, 295, 304, 306, 317, 318, 352

E electrolysis, 175 endochitinases, 247, 248, 258 environmental monitoring, 352, 358 environmental pollutants, xi, 241, 265, 357, 362, 363 environmental-friendly, 73, 83, 97 enzymes, ix, 1, 5, 6, 7, 8, 9, 10, 13, 14, 15, 16, 23, 24, 26, 29, 30, 32, 35, 36, 40, 79, 84, 86, 87, 89, 94, 100, 102, 108, 110, 115, 118, 119, 121, 122, 126, 140, 144, 147, 176, 177, 180, 181, 182, 183, 188, 190, 239, 241, 247, 252, 255, 256, 259, 260, 268, 271, 272, 287, 324, 326, 351, 352, 357, 358 eutrophication, 74, 169, 251, 297 ex situ bioremediation, 211, 212, 213, 215, 233, 235 exochitinases, 247 exposure time, 34, 71, 73, 76, 96, 303


Index F fenton, 114, 171, 175, 286 Fe-reduction, 327, 328, 336, 337, 341, 347 fertilizer(s), 28, 87, 152, 250, 251, 295, 297, 302, 312, 337 fishery, 239, 248, 252 fixed bed (film) reactors, 130 flocculation, 34, 49, 52, 67, 114, 171, 172, 184 flotation, 34, 49, 51 fluidized bed bioreactor, 130 fluoranthene/pyrene, 211, 212, 218, 228, 231, 232, 233 fluorescent biosensors, 359, 363 food and agriculture organization, 239 Fourier transform infrared spectroscopy (FTIR), 25, 151, 152, 160, 161, 162, 259, 260, 273, 279, 280, 289, 291

G genotoxicity, 85 granulation, 52, 56, 57, 59, 66, 68, 217 green fluorescent protein (GFP), 354, 356, 358, 359

H heavy metal tolerance, 13, 71, 72, 81, 96 heavy metals, x, xi, 1, 3, 4, 5, 11, 12, 13, 14, 15, 16, 17, 18, 24, 33, 43, 58, 72, 73, 74, 77, 78, 79, 80, 81, 82, 83, 84, 86, 89, 90, 91, 92, 93, 94, 96, 97, 98, 99, 100, 101, 103, 104, 105, 106, 107, 108, 109, 110, 111, 151, 152, 154, 155, 156, 158, 159, 160, 161,163, 164, 166, 167, 185, 187, 190, 191, 192, 193, 194, 195, 196, 197, 198, 199, 200, 201, 202, 219, 265, 297, 298, 306, 307, 309, 318, 319, 320, 326, 343, 352, 355, 358

I immobilization, 11, 18, 56, 83, 185, 191, 193, 210, 326, 356, 362 immobilization of microbes, 56 immobilized algal cells, 83, 96 immobilized biomass, 130, 132, 137, 321 incineration, 13, 24, 260, 264, 265

369

induced chlorophyll fluorescence, 72, 73, 90, 91, 94, 95 industrial effluents, xi, 36, 176, 295, 297, 298, 301, 305, 306, 307, 309, 313, 318, 319 industrialization, xi, 50, 151, 152, 242 ion exchange, 2, 11, 153, 154, 155, 158, 160, 161, 167, 171, 173, 174, 178 iron hydroxides, 328, 333, 334, 335, 349

L laccases, 7, 9, 15, 17, 29, 36 land filling, 260, 264, 265 lipid peroxidation, 86, 89, 100, 110, 116

M membrane filtration, 34, 171, 172, 178 metal detoxification, 99, 102, 191, 325 metal transformation, 185 methanogenesis, 133, 327, 328, 331, 336, 337, 338, 344, 347 microalgae, 15, 53, 67, 71, 72, 73, 74, 75, 76, 77, 82, 83, 85, 86, 89, 92, 96, 103, 104, 180, 312, 313, 314, 315, 320, 321, 322, 323, 325 microbes, vii, ix, 1, 2, 3, 5, 6, 7, 10, 11, 12, 13, 15, 26, 37, 52, 55, 62, 151, 186, 195, 197, 202, 205, 239, 240, 241, 246, 253, 259, 260, 261, 266, 267, 271, 274, 278, 284, 286, 291, 296, 302, 311, 328, 330, 352, 353, 355, 357, 358 microbial biosensors, 352, 360 microbial cell(s), ix, 5, 23, 26, 37, 49, 50, 51, 53, 54, 55, 56, 58, 59, 62, 114, 115, 116, 127, 129, 179, 192, 274, 363 microbial enzymes, 2, 6, 7, 23, 24, 35, 37, 116, 239, 272, 283, 285, 291, 292 microbial fuel cells, 61, 63, 65, 66, 67, 68, 69 microfiltration, 172, 174 mitigation, vii, x, 111, 151, 152, 155, 163, 204 mitochondrial oxidative phosphorylation, 115, 116 mobilization of arsenic, 328, 343 monooxygenases, 7, 8, 11, 13, 30, 117, 118, 122, 123 mycoremediation, 17, 29, 267, 286 mycorrhizal fungi, x, 185, 196, 207

N N-acetyl-D-glucosamine (NAG), 244, 246, 252


Index

370

nanofiltration, 34, 64, 172, 173, 174 nickel, vii, ix, 71, 73, 78, 79, 81, 83, 85, 86, 89, 91, 92, 93, 94, 95, 96, 97, 98, 99, 100, 101, 102, 103, 104, 105, 107, 108, 109, 110, 111, 112, 153, 166, 186, 206, 207, 208, 326, 344 noise pollution, 240, 249, 252 non-photochemical quenching of chlorophyll fluorescence, 91

O oil hydrocarbons, xi, 327, 328, 335, 342 oil pollutant, x, 211, 212, 213, 215, 218, 219, 220, 221, 222, 223, 224, 225, 233, 234 oxidation of methane, 327, 339, 340, 344, 346, 347 oxidation of oil hydrocarbons, 344 oxidative damage, 85, 88, 98, 193 oxidoreductases, 7, 30 oxygenases, 7, 8, 13, 118 ozonation, 171, 175, 176, 178

P pathogens, 10, 186, 190, 204, 247, 251, 254 peroxidases, 7, 10, 15, 24, 29, 30, 41, 260 petro-chemicals, 296, 309 petroleum polluted soil, 218, 219, 236 phenanthrene, 211, 218, 226, 228, 229, 232, 233, 297, 300, 301 phenolic compounds, xi, 7, 9, 10, 16, 17, 113, 114, 115, 116, 118, 119, 121, 124, 133, 141, 183, 301, 302, 311, 322 phenolics, 98, 114, 119, 133, 190, 193, 296, 302 photodegradation, 22, 264, 278, 287 photosynthetic pigments, 77, 87, 97, 108 phycoremediation, 28, 30, 313, 314, 322 physiological markers, 71, 73, 89, 94, 101 phytochelatin, 79, 81 phytoextraction, 29, 72, 78, 98, 101, 102, 104, 105, 194, 195, 198, 206, 207, 208 phytohormones, 188, 189, 190, 192, 197, 202, 204, 207 phytopathogenic microorganisms, 186, 189 phytoremediation, x, 8, 28, 38, 40, 47, 77, 80, 92, 96, 97, 98, 99, 100, 101, 103, 104, 105, 107, 108, 110, 111, 185, 187, 191, 192, 193, 194, 195, 196, 197, 198, 201, 202, 205, 206, 207, 208, 210 phytosequestration, 78

pile, 213, 214, 216, 220, 221, 222, 223, 226, 227, 228, 232 pKa value, 115 plant growth promoting bacteria (PGPB), viii, x, 185, 186, 187, 188, 189, 190, 191, 192, 193, 194, 195, 196, 197, 198, 199, 200, 201, 202, 203, 204, 205, 208, 209 plant growth promotion mechanisms, 185, 187 plant growth regulators, 189 plant growth-promoting rhizobacteria (PGPR), 187, 188, 193, 203, 204, 205, 206, 209 plant hormones, 186, 189 plastic waste management, x, 259, 260, 261, 264, 265, 266 plastics, ix, x, 19, 20, 21, 22, 23, 24, 27, 28, 42, 44, 45, 46, 140, 259, 260, 261, 262, 263, 264, 265, 266, 267, 271, 272, 274, 275, 276, 277, 278, 281, 282, 283, 284, 285, 286, 287, 288, 290, 291, 292, 352 plastics-degrading microbes, 259 pollutants, viii, xi, 1, 3, 4, 5, 6, 7, 11, 13, 16, 24, 39, 49, 52, 53, 58, 59, 66, 71, 75, 77, 93, 96, 105, 107, 114, 115, 116, 117, 118, 119, 129, 130, 133, 134, 137, 138, 139, 152, 154, 163, 164, 167, 171, 174, 176, 178, 184, 187, 188, 191, 194, 196, 210, 211, 235, 236, 239, 240, 241, 242, 252, 265, 266, 281, 295, 296, 298, 300, 301, 302, 304, 305, 306, 309, 313, 323, 324, 329, 331, 333, 343, 346, 351, 352, 358 pollution, ix, x, 8, 12, 15, 17, 24, 28, 38, 39, 47, 73, 75, 77, 78, 81, 88, 90, 94, 95, 96, 98, 99, 102, 103, 104, 106, 108, 111, 147, 151, 152, 164, 165, 167, 173, 175, 178, 185, 194, 204, 205, 206, 208, 209, 210, 213, 226, 236, 237, 248, 264, 295, 296, 314, 315,317, 322, 328, 329, 331, 332, 340, 344, 345, 348, 360, 362 polycyclic aromatic hydrocarbons, 3, 4, 13, 137, 211, 212, 213, 226, 237, 299, 301 priority pollutants, 114

Q quantum yield of photosynthesis, 90 quicklime, 251


Index R recalcitrant, vii, ix, x, 19, 20, 23, 31, 36, 37, 133, 154, 179, 182, 233, 259, 261, 275, 292 recombinant DNA techniques, 242 recombinant plasmids, 358 reporter gene, 351, 352, 353, 354, 357, 359 reproductive rate, 74, 94, 97 reverse osmosis, 173 rhizobacteria, 187, 207 rhizoremediation, 29, 42, 208

S scanning electron microscopy, 64, 151, 160, 162, 259, 260, 288 Scenedesmus acuminatus, 87, 88, 91, 92, 94, 95, 96 seafood industry, x, 239, 242 seafood waste, x, 239, 249, 253, 255, 256, 257 sedimentation, 52, 53, 307 shrimp shells, 161, 243, 246 siderophore, 191, 195, 209 signal production, 356 sludge, 34, 52, 56, 63, 65, 66, 67, 68, 69, 86, 97, 115, 116, 118, 132, 133, 134, 136, 137, 139, 140, 142, 145, 146, 149, 154, 169, 171, 178, 184, 236, 305, 359 soil, viii, x, 5, 7, 9, 11, 12, 14, 15, 16, 18, 23, 24, 25, 28, 29, 30, 31, 34, 40, 42, 43, 44, 45, 46, 66, 98, 104, 105, 108, 110, 111, 112, 113, 114, 140, 145, 152, 164, 181, 187, 188, 189, 190, 191, 192, 193, 194, 196, 197, 198, 199, 201, 203, 204, 205, 206, 207, 208, 209, 210, 211, 212, 213, 214, 215, 216, 217, 219, 220, 222, 224, 227, 232, 233, 234, 235, 236, 237, 238, 240, 241, 242, 250, 277, 280, 282, 283, 284, 286, 288, 290, 292, 295, 303, 312, 316, 317, 321, 327, 328, 329, 330, 331, 333, 334, 337, 339, 342, 343, 347, 348, 349, 350, 354, 357, 359 soil microorganisms, 187, 196, 212, 339 solid-liquid separation, 52, 57 substituted phenols, x, 113, 114, 115, 117, 118, 119, 120, 126, 128, 130, 133, 138, 140 sulfate reduction, 14, 327, 328, 331, 332, 339, 340, 341, 342, 344, 347 surface free energy, ix, 49, 50, 68 surfactant, 259, 271, 272, 287

371 T

TCA cycle, 121 TDS, 298, 309 textile due, 169 textile dyes, x, 34, 169, 170, 171, 172, 177, 180, 182, 183, 206, 304, 309, 318 tolerance, 53, 74, 75, 76, 77, 80, 89, 90, 94, 96, 97, 99, 102, 103, 105, 106, 109, 110, 111, 115, 120, 129, 186, 189, 190, 195, 196, 198, 204, 205, 210, 303, 316, 317, 324 transfer electron microscopy, 151 TSS, 248, 298, 309

U ultra filtration, 172, 174 urbanization, ix, 20, 50, 151, 152, 242

V vitality index, 90

W waste management, 1, 243, 259, 264, 272, 281 wastewater, ix, 14, 15, 17, 36, 49, 50, 52, 53, 56, 58, 59, 60, 61, 62, 73, 75, 76, 77, 82, 83, 100, 102, 103, 109, 111, 112, 113, 114, 115, 117, 118, 119, 130, 131, 133, 134, 135, 136, 137, 138, 140, 142, 143, 144, 145, 149, 155, 164, 165, 167, 169, 171, 172, 173, 174, 175, 176, 178, 181, 182, 296, 298, 299, 301, 303, 305, 306, 307, 308, 309, 311, 312, 313, 314, 315, 319, 322, 326, 357 wastewater effluents, 169, 326 wastewater treatment, 49, 50, 52, 53, 56, 59, 61, 62, 73, 75, 76, 77, 82, 83, 102, 103, 104, 109, 114, 115, 140, 149, 164, 165, 175, 305, 322 water pollution, 72, 73, 75, 76, 78, 81, 84, 85, 91, 93, 101, 167, 240, 248, 305, 306 water-polluting agents, 76

X xenobiotic compounds, 20, 242 x-ray photoelectron spectroscopy, 151
