Emerging Technologies in Environmental Bioremediation [1 ed.] 0128198605, 9780128198605

Table of contents :
Cover
Emerging Technologies in Environmental Bioremediation
Copyright
Contents
List of contributors
Preface
1 Immobilization of anaerobic ammonium oxidation bacteria for nitrogen-rich wastewater treatment
1.1 Introduction
1.2 Anammox bacteria and their metabolic process
1.3 Cell immobilization: a strategy to improve microbial wastewater treatment
1.3.1 What is cell immobilization?
1.3.2 Different approaches for cell immobilization
1.3.2.1 Granulation
1.3.2.2 Biofilm formation
1.3.2.3 Gel entrapment
1.4 Why is gel immobilization advantageous?
1.5 Gel materials used for the immobilization of anammox
1.5.1 Polyvinyl alcohol and polyvinyl alcohol/sodium alginate
1.5.2 Waterborne polyurethane
1.5.3 Polyethylene glycol gel
1.6 Application of cell immobilization in anammox and partial nitrification
1.6.1 Application of immobilized anammox
1.7 Commercialization of immobilizing technology
1.8 Conclusion
Acknowledgments
References
2 Accelerated bioremediation of petroleum refinery sludge through biostimulation and bioaugmentation of native microbiome
2.1 Introduction
2.2 Petroleum refinery waste: composition and hazard
2.3 Microbiology of hydrocarbon-associated environments
2.4 Microbial bioremediation of waste sludge
2.4.1 Accelerated bioremediation
2.4.1.1 Biostimulation
2.4.1.2 Bioaugmentation
2.5 Factors affecting bioremediation
2.6 Future scope
References
Further reading
3 Degradation and detoxification of waste via bioremediation: a step toward sustainable environment
3.1 Introduction
3.2 Bioremediation and the role of bioavailability
3.2.1 Surfactants
3.2.2 Biodegradation
3.2.3 In situ and ex situ bioremediation
3.3 The degradation and/or detoxification of pollutants
3.3.1 Heavy metal pollutant
3.3.1.1 Sources
3.3.1.2 Bioremediation of heavy metals
3.3.1.3 Adsorption
3.3.1.4 Microorganisms for the detoxification of heavy metals
3.3.2 Dyes
3.3.2.1 Physical and chemical removal of dye effluent
3.3.2.2 Biological treatment
3.4 Role of genetic engineering in bioremediation
3.4.1 Bioremediation through microbial systems biology
3.4.1.1 Genomics
3.4.1.2 Metagenomics
3.4.1.3 Transcriptomics
3.4.1.4 Proteomics and metaproteomics
3.5 Limitations and future prospect
Acknowledgment
Competing interests
References
Further reading
4 Fungal laccases: versatile green catalyst for bioremediation of organopollutants
4.1 Introduction
4.2 Distribution and physiological functions of laccases
4.3 Production of laccases
4.3.1 Screening of laccase-producing fungi
4.3.2 Cultural and nutritional conditions for laccase production
4.3.3 Heterologous production of laccases
4.3.4 Biochemical properties of laccases
4.3.4.1 Kinetic properties of laccases
4.3.4.2 Effect of pH and temperature on activity of laccases
4.3.4.3 Effect of inhibitors on activity of laccases
4.3.5 Mode of action of laccases
4.3.6 Classification of laccases according to substrate specificity
4.3.7 Laccase mediator system
4.3.8 Immobilization of laccase
4.4 Application of laccases for bioremediation of environmental pollutants
4.4.1 Degradation of xenobiotic compounds
4.4.2 Decolorization of synthetic dyes
4.4.3 Treatment of industrial effluent
4.4.4 Potential applications in pulp and paper industry
4.4.5 Applications of laccases to develop ecofriendly processes
4.5 Limitations and future prospects
References
5 Emerging bioremediation technologies for the treatment of wastewater containing synthetic organic compounds
5.1 Introduction
5.2 Electrobioremediation
5.3 Bioelectrochemical systems/technology
5.4 Phytotechnology (phytoremediation)
5.4.1 Phytoreactors and constructed wetlands
5.4.2 Plant–microbe phytoremediation
5.4.3 Plant enzymes and metabolites
5.4.4 Hydroponic systems
5.4.5 Plant tissue culturing
5.5 Electron beam irradiation
5.6 Conclusion: unresolved challenges and future perspectives
Acknowledgments
References
6 Bacterial quorum sensing in environmental biotechnology: a new approach for the detection and remediation of emerging pol...
6.1 Introduction
6.2 Mechanisms of bacterial quorum sensing
6.2.1 Two-component system in Gram-positive bacteria
6.2.2 Acyl homoserine lactone in Gram-negative bacteria
6.3 Quorum sensing in environmental biotechnology
6.3.1 Heavy metal detection
6.3.2 Pathogen detection
6.3.3 Bioremediation
6.3.4 Biofilm formation
6.3.5 Hydrocarbon remediation
6.4 Limitations of microbial quorum sensing
6.5 Conclusion
References
7 Bioremediation: an effective technology toward a sustainable environment via the remediation of emerging environmental po...
7.1 Introduction
7.2 Emerging pollutants
7.2.1 Bisphenol A
7.2.1.1 Inorganic–organic clays
7.2.1.2 Photodegradation
7.2.2 Polycyclic aromatic hydrocarbons
7.2.2.1 Bacterial catabolism of polycyclic aromatic hydrocarbons
7.2.2.2 Halophilic/halotolerant bacteria and archaea for the degradation of polycyclic aromatic hydrocarbons
7.2.2.3 Fungal degradation
7.2.2.4 Chemical oxidation and photodegradation
7.2.3 Polychlorinated biphenyls
7.2.3.1 Biological degradation
7.2.3.2 Halogenated organic compounds technology for polychlorinated biphenyls degradation
7.2.4 Pharmaceutical wastes
7.2.4.1 Photodegradation
7.2.4.2 Sludge treatment
7.2.5 Hospital effluents as source of emerging pollutants
7.2.5.1 Biological treatment
7.2.6 Other emerging pollutants
7.3 Types of bioremediation
7.3.1 Microbial bioremediation
7.3.1.1 Bacterial bioremediation
7.3.1.2 Mycoremediation
7.3.2 Phycoremediation
7.3.3 Mixed cell culture system
7.3.4 Phytoremediation
7.3.4.1 Phytoextraction
7.3.4.2 Rhizofiltration
7.3.4.3 Phytostabilization
7.3.4.4 Phytodegradation
7.3.4.5 Phytovolatilization
7.3.4.6 Rhizoremediation
7.3.5 Enzymatic bioremediation
7.3.6 Zooremediation
7.3.7 Vermiremediation
7.4 Emerging techniques
7.4.1 Application of biosurfactants
7.4.2 Immobilization techniques
7.4.3 Adsorption and electrostatic binding
7.4.4 Entrapment in porous matrix and encapsulation
7.4.5 Electrokinetic remediation
7.4.6 Metagenomics
7.4.7 Protein engineering
7.4.8 Bioinformatics
7.4.9 Nanotechnology
7.4.10 Genetic engineering
7.4.11 Designer microbe and plant approach
7.4.12 Rhizosphere engineering
7.4.13 Manipulation of plant–microbe symbiosis
7.4.14 Cometabolic bioremediation
7.5 Conclusion
Acknowledgment
Competing interests
References
8 Application of metagenomics in remediation of contaminated sites and environmental restoration
8.1 Introduction
8.2 Mechanism of bioremediation
8.3 Approaches used to study microbial communities involved in in situ and ex situ bioremediation
8.3.1 Culture-based techniques
8.3.2 Culture-independent techniques
8.3.2.1 Polymerase chain reaction–based molecular techniques
8.3.2.1.1 Denaturing gradient gel electrophoresis/temperature gradient gel electrophoresis
8.3.2.1.2 Single-strand conformation polymorphism
8.3.2.1.3 Restriction fragment length polymorphism/amplified ribosomal DNA restriction analysis
8.3.2.1.4 Terminal restriction fragment length polymorphisms
8.3.2.1.5 Automated ribosomal intergenic spacer analysis
8.3.2.1.6 Random amplified polymorphic DNA
8.3.2.1.7 Stable-isotope probing
8.3.2.1.8 Quantitative polymerase chain reaction
8.3.2.2 Nonpolymerase chain reaction–based molecular techniques
8.3.2.2.1 Fluorescence in situ hybridization
8.3.2.2.2 Guanine plus cytosine content
8.3.2.2.3 Microarrays
8.4 Metagenomics: a culture-independent insight
8.4.1 Functional-based metagenomics
8.4.2 Sequence-based metagenomics
8.4.3 Metatranscriptomics
8.4.4 Metaproteomics
8.4.5 Metabolomics
8.4.6 Metagenomics sequencing strategies
8.5 Next-generation sequencing technologies to explore structure and function of microbial communities
8.6 Conclusion
References
Further reading
9 In situ bioremediation techniques for the removal of emerging contaminants and heavy metals using hybrid microbial electr...
9.1 Introduction
9.1.1 Bioremediation for pollution control and classification of bioremediation techniques
9.1.2 Microbial electrochemical technology
9.2 In situ bioremediation using microbial electrochemical technologies
9.2.1 Constructed wetlands-microbial fuel cells
9.2.2 Sediment-microbial fuel cells
9.2.3 Soil-microbial fuel cells
9.2.4 Plant-microbial fuel cells
9.3 Future scope of research
9.4 Summary
References
10 Gene-targeted metagenomics approach for the degradation of organic pollutants
10.1 Introduction
10.2 Gene-targeted metagenomics
10.3 Methods used for metagenomics studies
10.4 Bacterial community abundance
10.4.1 Biodegradation pathway involved in the degradation of organic compounds
10.4.1.1 Aerobic pathway
10.4.1.2 Anaerobic pathways
10.4.2 Functional metagenomics
10.5 Conclusion
10.6 Future perspective
References
Further reading
11 Current status of toxic wastewater control strategies
11.1 Introduction
11.2 Causes and effects of toxic wastewater pollution
11.3 Current interventions in toxic wastewater control
11.3.1 Treatment using aquatic systems
11.3.2 Treatment using microalgae
11.3.3 Treatment using vermifiltration
11.3.4 Other interventions in toxic wastewater control
11.4 Wastewater reuse
11.5 Conclusion
Acknowledgments
References
12 Latest innovations in bacterial degradation of textile azo dyes
12.1 Introduction
12.2 Bacteria in degradation of azo dyes
12.2.1 Bacteria as source
12.2.2 Mechanism of azo dye degradation by bacteria
12.2.3 Phases of treatment
12.2.4 Recent studies in bacterial mediated azo dye degradation
12.2.5 Analytical methods in azo dye degradation
12.2.6 Analysis of efficiency of bacterial dye degradation by toxicity tests
12.3 Computational inputs in enhancing biodegradation
12.3.1 Choice of strains: adapted versus nonadapted strains
12.3.2 In silico analysis as a valuable tool
12.4 Alternative front-runners: fungi, yeast, and algae-mediated azo dye degradation
12.5 Future perspective
References
13 Development in wastewater treatment plant design
13.1 Introduction
13.1.1 Conventional wastewater treatment technology
13.1.1.1 Primary treatment
13.1.1.1.1 Screens
13.1.1.1.2 Grit chambers
13.1.1.1.3 Primary settlers or sedimentation
13.1.1.2 Secondary treatment
13.1.1.2.1 Aerobic treatment system
13.1.1.2.2 Anaerobic treatment
13.1.1.3 Tertiary treatment
13.1.2 Recent advances achieved in wastewater treatment plant
13.1.2.1 Primary treatment
13.1.2.2 Secondary treatment
13.1.2.2.1 Aerobic secondary treatment
13.1.2.2.2 Anaerobic secondary treatment
13.2 Tertiary treatment
13.3 Conclusion
References
14 Engineering biomaterials for the bioremediation: advances in nanotechnological approaches for heavy metals removal from ...
14.1 Introduction
14.2 Bioremediation
14.3 Nanotechnology and bioremediation
14.3.1 Nanomaterials used for removing pollutants
14.3.1.1 Carbon-based nanomaterials
14.3.1.2 Hydroxyapatite nanomaterials
14.3.1.3 Metallic nanoparticles
14.3.1.4 Biogenic uraninite nanoparticles
14.3.1.5 Dendrimers
14.3.1.6 Polymeric nanocomposites
14.3.1.7 Nanozymes
14.3.1.8 Microorganisms
14.3.2 Bioremediation of soil
14.3.2.1 Phytoremediation
14.3.2.2 Microbial bioremediation
14.3.2.3 Nanomaterials based bioremediation
14.4 Conclusion
Acknowledgement
References
15 Algal–bacterial symbiosis and its application in wastewater treatment
15.1 Introduction
15.2 The symbiotic process
15.2.1 Exchange of information in the form of bioactive compounds for symbiosis
15.2.1.1 Quorum sensing
15.2.2 Exudates that can inhibit the microbes in the vicinity
15.2.3 Exudates that can stimulate the microbes in the vicinity
15.2.3.1 Cofactor auxotrophy
15.2.4 Factors affecting symbiotic systems
15.2.4.1 Dissolved oxygen
15.2.4.2 Carbon dioxide
15.2.4.3 Light
15.2.4.4 Hydraulic retention time
15.2.4.5 Initial algae:bacteria ratio
15.2.4.6 Substrate concentrations
15.2.4.7 pH and temperature
15.3 Applications in wastewater treatment
15.3.1 Types of reactors
15.3.2 Nutrients removal
15.3.3 Metal removal
15.3.4 Organic matter removal
15.3.5 Emerging contaminants removal
15.3.6 Removal of refractory compounds
15.4 Energy generation
15.4.1 Algal biohydrogen production
15.4.2 Algal lipid production
15.4.3 Microbial fuel cell reactor using algal–bacteria interaction
15.5 Conclusion and future directions
References
16 Role of plant growth–promoting rhizobacteria in mitigation of heavy metals toxicity to Oryza sativa L.
16.1 Introduction
16.2 Different genera of plant growth–promoting rhizobacteria
16.3 Plant growth–promoting rhizobacteria role in heavy metals dynamics in the soil
16.4 Plant growth–promoting rhizobacteria’s role in controlling pathogens in rice
16.5 Plant growth–promoting rhizobacteria in remediation of the environment
16.6 Conclusion and future prospect
References
Further reading
17 Study of transport models for arsenic removal using nanofiltration process: recent perspectives
17.1 Introduction
17.1.1 Sources
17.1.1.1 Anthropogenic sources
17.1.1.2 Natural sources
17.1.2 Health effects
17.1.3 Permissible limit
17.2 Chemistry of arsenic
17.3 Methods of arsenic removal from water/wastewater
17.3.1 Membrane technology
17.4 Nanofiltration of arsenic
17.4.1 Modeling of nanofiltration membranes for arsenic removal
17.5 Conclusion and future perspective
Nomenclature
Greek symbols
Abbreviations
References
Further Reading
18 Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms
18.1 Introduction
18.2 Remediation of Pb
18.2.1 Conventional methods for Pb remediation or recovery
18.2.2 Bioremediation of Pb
18.2.3 Mechanisms of bioremediation
18.3 Case study of aqueous Pb biorecovery by local Pb-resistant organisms
18.3.1 Characterization of bacterial consortia
18.3.2 Precipitate identification
18.3.3 Microbiological and kinetic study
18.3.3.1 Laboratory-scale system design and experimental approach
18.3.3.2 Laboratory-scale system results and discussion
18.3.4 Case studies of varied operating conditions
18.3.4.1 Effect of elevated heavy metal concentrations
18.3.4.2 Effect of Zn(II) or Cu(II) ions on Pb(II) bioprecipitation
18.3.4.3 The minimum inhibitory concentration of Pb(II) for the B-consortium
18.3.4.4 Effect of substrate composition
18.4 Conclusion and outlook
Acknowledgment
References
19 Microbial bioremediation of azo dye through microbiological approach
19.1 Introduction
19.2 Classification of dyes
19.3 Role of environmental parameters on microbial biodegradation and bioremediation of azo dye
19.4 Effects of environmental parameters on azo dye degradation
19.4.1 Temperature
19.4.2 Oxygen
19.4.3 Dye concentration
19.4.4 Electron donor
19.4.5 pH
19.4.6 Dye structure
19.4.7 Redox potential
19.4.8 Redox mediator
19.4.9 Decolorization by genetically modified organisms
19.5 Conclusion
References
Further reading
20 Novel process of ellagic acid synthesis from waste generated from mango pulp processing industries
20.1 Introduction
20.1.1 Waste from mango pulp processing industries
20.2 Composition of mango wastes
20.3 Types of tannins
20.4 Bioconversion of tannin to ellagic acid
20.5 Microbes involved in the production of ellagic acid
20.6 Applications of ellagic acid
20.7 Conclusion
References
Further reading
Index
Back Cover

Citation preview

Emerging Technologies in Environmental Bioremediation

Emerging Technologies in Environmental Bioremediation Edited by Maulin P. Shah Industrial Waste Water Research Lab, Division of Applied & Environmental Microbiology, Enviro Technology Limited, India

Susana Rodriguez-Couto Ikerbasque, Basque Foundation for Science, Maria Diaz de Haro 3, Bilbao, Spain

¨r S. Sevinc¸ ¸Sengo Middle East Technical University, Department of Environmental Engineering, Ankara, Turkey

Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, Netherlands The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States Copyright © 2020 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress ISBN: 978-0-12-819860-5 For Information on all Elsevier publications visit our website at https://www.elsevier.com/books-and-journals

Publisher: Susan Dennis Acquisition Editor: Kostas Marinakis Editorial Project Manager: Vincent Gabrielle Production Project Manager: Omer Mukthar Cover Designer: Typeset by MPS Limited, Chennai, India

Contents List of contributors ............................................................................................... xv Preface.................................................................................................................xix Chapter 1: Immobilization of anaerobic ammonium oxidation bacteria for nitrogen-rich wastewater treatment ........................................................ 1 Shou-Qing Ni, Hafiz Adeel Ahmad and Shakeel Ahmad 1.1 Introduction .......................................................................................................... 1 1.2 Anammox bacteria and their metabolic process ................................................. 3 1.3 Cell immobilization: a strategy to improve microbial wastewater treatment .... 5 1.3.1 What is cell immobilization? ...................................................................... 5 1.3.2 Different approaches for cell immobilization .............................................. 5 1.4 Why is gel immobilization advantageous?.......................................................... 8 1.5 Gel materials used for the immobilization of anammox..................................... 9 1.5.1 Polyvinyl alcohol and polyvinyl alcohol/sodium alginate ......................... 10 1.5.2 Waterborne polyurethane .......................................................................... 10 1.5.3 Polyethylene glycol gel............................................................................. 11 1.6 Application of cell immobilization in anammox and partial nitrification........ 12 1.6.1 Application of immobilized anammox ...................................................... 12 1.7 Commercialization of immobilizing technology ............................................... 13 1.8 Conclusion .......................................................................................................... 17 Acknowledgments....................................................................................................... 18 References................................................................................................................... 18

Chapter 2: Accelerated bioremediation of petroleum refinery sludge through biostimulation and bioaugmentation of native microbiome...................... 23 Jayeeta Sarkar, Ajoy Roy, Pinaki Sar and Sufia K. Kazy 2.1 Introduction ........................................................................................................ 23 2.2 Petroleum refinery waste: composition and hazard .......................................... 26 2.3 Microbiology of hydrocarbon-associated environments ................................... 28

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2.4 Microbial bioremediation of waste sludge ........................................................ 43 2.4.1 Accelerated bioremediation ...................................................................... 44 2.5 Factors affecting bioremediation ....................................................................... 49 2.6 Future scope ....................................................................................................... 51 References................................................................................................................... 51 Further reading............................................................................................................ 64

Chapter 3: Degradation and detoxification of waste via bioremediation: a step toward sustainable environment ........................................................... 67 Komal Agrawal and Pradeep Verma 3.1 Introduction ........................................................................................................ 67 3.2 Bioremediation and the role of bioavailability.................................................. 68 3.2.1 Surfactants ................................................................................................ 68 3.2.2 Biodegradation ......................................................................................... 69 3.2.3 In situ and ex situ bioremediation ............................................................. 70 3.3 The degradation and/or detoxification of pollutants ......................................... 70 3.3.1 Heavy metal pollutant............................................................................... 70 3.3.2 Dyes ......................................................................................................... 73 3.4 Role of genetic engineering in bioremediation ................................................. 76 3.4.1 Bioremediation through microbial systems biology .................................. 77 3.5 Limitations and future prospect ......................................................................... 78 Acknowledgment ........................................................................................................ 78 Competing interests .................................................................................................... 78 References................................................................................................................... 79 Further reading............................................................................................................ 83

Chapter 4: Fungal laccases: versatile green catalyst for bioremediation of organopollutants ................................................................................. 85 Ajit Patel, Vanita Patel, Radhika Patel, Ujjval Trivedi and Kamlesh Patel 4.1 Introduction ........................................................................................................ 85 4.2 Distribution and physiological functions of laccases ........................................ 87 4.3 Production of laccases........................................................................................ 88 4.3.1 Screening of laccase-producing fungi ....................................................... 88 4.3.2 Cultural and nutritional conditions for laccase production ........................ 89 4.3.3 Heterologous production of laccases ......................................................... 91 4.3.4 Biochemical properties of laccases ........................................................... 94 4.3.5 Mode of action of laccases ....................................................................... 97 4.3.6 Classification of laccases according to substrate specificity .................... 100

Contents vii 4.3.7 Laccase mediator system ........................................................................ 101 4.3.8 Immobilization of laccase ....................................................................... 104 4.4 Application of laccases for bioremediation of environmental pollutants ....... 106 4.4.1 Degradation of xenobiotic compounds .................................................... 107 4.4.2 Decolorization of synthetic dyes ............................................................. 110 4.4.3 Treatment of industrial effluent .............................................................. 113 4.4.4 Potential applications in pulp and paper industry .................................... 113 4.4.5 Applications of laccases to develop ecofriendly processes ...................... 114 4.5 Limitations and future prospects...................................................................... 115 References................................................................................................................. 116

Chapter 5: Emerging bioremediation technologies for the treatment of wastewater containing synthetic organic compounds .............................131 Kunal Jain, Jenny Johnson, Neelam Devpura, Rohit Rathour, Chirayu Desai, Onkar Tiwari and Datta Madamwar 5.1 5.2 5.3 5.4

Introduction ...................................................................................................... 131 Electrobioremediation ...................................................................................... 134 Bioelectrochemical systems/technology .......................................................... 135 Phytotechnology (phytoremediation)............................................................... 138 5.4.1 Phytoreactors and constructed wetlands .................................................. 138 5.4.2 Plant microbe phytoremediation ........................................................... 139 5.4.3 Plant enzymes and metabolites ............................................................... 140 5.4.4 Hydroponic systems................................................................................ 141 5.4.5 Plant tissue culturing .............................................................................. 141 5.5 Electron beam irradiation................................................................................. 142 5.6 Conclusion: unresolved challenges and future perspectives ........................... 144 Acknowledgments..................................................................................................... 146 References................................................................................................................. 146

Chapter 6: Bacterial quorum sensing in environmental biotechnology: a new approach for the detection and remediation of emerging pollutants .......151 Debapriya Sarkar, Kasturi Poddar, Nishchay Verma, Sayantani Biswas and Angana Sarkar 6.1 Introduction ...................................................................................................... 151 6.2 Mechanisms of bacterial quorum sensing ....................................................... 152 6.2.1 Two-component system in Gram-positive bacteria ................................. 153 6.2.2 Acyl homoserine lactone in Gram-negative bacteria ............................... 153 6.3 Quorum sensing in environmental biotechnology........................................... 154

viii

Contents 6.3.1 Heavy metal detection ............................................................................ 154 6.3.2 Pathogen detection .................................................................................. 158 6.3.3 Bioremediation ....................................................................................... 158 6.3.4 Biofilm formation ................................................................................... 159 6.3.5 Hydrocarbon remediation ....................................................................... 160 6.4 Limitations of microbial quorum sensing........................................................ 161 6.5 Conclusion........................................................................................................ 161 References................................................................................................................. 161

Chapter 7: Bioremediation: an effective technology toward a sustainable environment via the remediation of emerging environmental pollutants ...165 Komal Agrawal, Ankita Bhatt, Venkatesh Chaturvedi and Pradeep Verma 7.1 Introduction ...................................................................................................... 165 7.2 Emerging pollutants ......................................................................................... 166 7.2.1 Bisphenol A ............................................................................................ 166 7.2.2 Polycyclic aromatic hydrocarbons .......................................................... 168 7.2.3 Polychlorinated biphenyls ....................................................................... 169 7.2.4 Pharmaceutical wastes ............................................................................ 170 7.2.5 Hospital effluents as source of emerging pollutants ................................ 171 7.2.6 Other emerging pollutants ...................................................................... 172 7.3 Types of bioremediation .................................................................................. 173 7.3.1 Microbial bioremediation........................................................................ 173 7.3.2 Phycoremediation ................................................................................... 175 7.3.3 Mixed cell culture system ....................................................................... 176 7.3.4 Phytoremediation .................................................................................... 176 7.3.5 Enzymatic bioremediation ...................................................................... 179 7.3.6 Zooremediation....................................................................................... 179 7.3.7 Vermiremediation ................................................................................... 179 7.4 Emerging techniques ........................................................................................ 180 7.4.1 Application of biosurfactants ................................................................ 180 7.4.2 Immobilization techniques .................................................................... 181 7.4.3 Adsorption and electrostatic binding ..................................................... 181 7.4.4 Entrapment in porous matrix and encapsulation ................................... 181 7.4.5 Electrokinetic remediation .................................................................... 182 7.4.6 Metagenomics....................................................................................... 182 7.4.7 Protein engineering ............................................................................... 183 7.4.8 Bioinformatics ...................................................................................... 183 7.4.9 Nanotechnology .................................................................................... 183 7.4.10 Genetic engineering .............................................................................. 184 7.4.11 Designer microbe and plant approach ................................................... 184

Contents ix 7.4.12 Rhizosphere engineering ....................................................................... 185 7.4.13 Manipulation of plant microbe symbiosis ........................................... 185 7.4.14 Cometabolic bioremediation ................................................................. 186 7.5 Conclusion ........................................................................................................ 186 Acknowledgment ...................................................................................................... 186 Competing interests .................................................................................................. 186 References................................................................................................................. 187

Chapter 8: Application of metagenomics in remediation of contaminated sites and environmental restoration ...................................................................197 Vineet Kumar, Indu Shekhar Thakur, Ajay Kumar Singh and Maulin P. Shah 8.1 Introduction ...................................................................................................... 197 8.2 Mechanism of bioremediation ......................................................................... 200 8.3 Approaches used to study microbial communities involved in in situ and ex situ bioremediation............................................................................... 202 8.3.1 Culture-based techniques ........................................................................ 203 8.3.2 Culture-independent techniques .............................................................. 203 8.4 Metagenomics: a culture-independent insight ................................................. 215 8.4.1 Functional-based metagenomics ............................................................. 216 8.4.2 Sequence-based metagenomics ............................................................... 217 8.4.3 Metatranscriptomics................................................................................ 218 8.4.4 Metaproteomics ...................................................................................... 218 8.4.5 Metabolomics ......................................................................................... 219 8.4.6 Metagenomics sequencing strategies....................................................... 220 8.5 Next-generation sequencing technologies to explore structure and function of microbial communities.................................................................. 220 8.6 Conclusion ........................................................................................................ 224 References................................................................................................................. 225 Further reading.......................................................................................................... 232

Chapter 9: In situ bioremediation techniques for the removal of emerging contaminants and heavy metals using hybrid microbial electrochemical technologies ................................................................233 M.M. Ghangrekar, S.M. Sathe and I. Chakraborty 9.1 Introduction ...................................................................................................... 233 9.1.1 Bioremediation for pollution control and classification of bioremediation techniques ...................................................................... 234 9.1.2 Microbial electrochemical technology .................................................... 235

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9.2 In situ bioremediation using microbial electrochemical technologies............ 235 9.2.1 Constructed wetlands-microbial fuel cells............................................... 235 9.2.2 Sediment-microbial fuel cells ................................................................. 241 9.2.3 Soil-microbial fuel cells .......................................................................... 243 9.2.4 Plant-microbial fuel cells ........................................................................ 246 9.3 Future scope of research .................................................................................. 250 9.4 Summary............................................................................................................. 251 References................................................................................................................. 251

Chapter 10: Gene-targeted metagenomics approach for the degradation of organic pollutants ..........................................................................257 Raghawendra Kumar, Dinesh Kumar, Labdhi Pandya, Priti Raj Pandit, Zarna Patel, Shivarudrappa Bhairappanavar and Jayashankar Das 10.1 10.2 10.3 10.4

Introduction .................................................................................................... 257 Gene-targeted metagenomics ......................................................................... 258 Methods used for metagenomics studies ....................................................... 259 Bacterial community abundance.................................................................... 262 10.4.1 Biodegradation pathway involved in the degradation of organic compounds ........................................................................................... 262 10.4.2 Functional metagenomics ..................................................................... 265 10.5 Conclusion...................................................................................................... 268 10.6 Future perspective .......................................................................................... 269 References................................................................................................................. 269 Further reading.......................................................................................................... 273

Chapter 11: Current status of toxic wastewater control strategies .......................275 Sushma Chityala, Dharanidaran Jayachandran, Ashish A. Prabhu and Veeranki Venkata Dasu 11.1 Introduction .................................................................................................... 275 11.2 Causes and effects of toxic wastewater pollution ......................................... 276 11.3 Current interventions in toxic wastewater control ........................................ 277 11.3.1 Treatment using aquatic systems .......................................................... 277 11.3.2 Treatment using microalgae .................................................................. 278 11.3.3 Treatment using vermifiltration ............................................................ 278 11.3.4 Other interventions in toxic wastewater control .................................... 279 11.4 Wastewater reuse............................................................................................ 281 11.5 Conclusion...................................................................................................... 282 Acknowledgments..................................................................................................... 282 References................................................................................................................. 282

Contents xi

Chapter 12: Latest innovations in bacterial degradation of textile azo dyes ..........285 Shantkriti Srinivasan, Kanyaga Parameswari M and Siranjeevi Nagaraj 12.1 Introduction .................................................................................................... 285 12.2 Bacteria in degradation of azo dyes .............................................................. 287 12.2.1 Bacteria as source ................................................................................. 287 12.2.2 Mechanism of azo dye degradation by bacteria .................................... 288 12.2.3 Phases of treatment ............................................................................... 289 12.2.4 Recent studies in bacterial mediated azo dye degradation..................... 290 12.2.5 Analytical methods in azo dye degradation .......................................... 294 12.2.6 Analysis of efficiency of bacterial dye degradation by toxicity tests .... 295 12.3 Computational inputs in enhancing biodegradation ...................................... 296 12.3.1 Choice of strains: adapted versus nonadapted strains ............................ 296 12.3.2 In silico analysis as a valuable tool....................................................... 298 12.4 Alternative front-runners: fungi, yeast, and algae-mediated azo dye degradation ..................................................................................................... 300 12.5 Future perspective .......................................................................................... 301 References................................................................................................................. 302

Chapter 13: Development in wastewater treatment plant design .........................311 Bapi Mandal, Anwesha Purkayastha, Ashish A. Prabhu and Veeranki Venkata Dasu 13.1 Introduction .................................................................................................... 311 13.1.1 Conventional wastewater treatment technology .................................... 312 13.1.2 Recent advances achieved in wastewater treatment plant ..................... 315 13.2 Tertiary treatment........................................................................................... 320 13.3 Conclusion ...................................................................................................... 320 References................................................................................................................. 320 Chapter 14: Engineering biomaterials for the bioremediation: advances in nanotechnological approaches for heavy metals removal from natural resources ...........................................................................323 Magapu Solomon Sudhakar, Aakriti Aggarwal and Mahesh Kumar Sah 14.1 Introduction .................................................................................................... 323 14.2 Bioremediation ............................................................................................... 325 14.3 Nanotechnology and bioremediation ............................................................. 325 14.3.1 Nanomaterials used for removing pollutants ......................................... 326 14.3.2 Bioremediation of soil .......................................................................... 332 14.4 Conclusion ...................................................................................................... 333 Acknowledgement .................................................................................................... 334 References................................................................................................................. 334

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Chapter 15: Algal bacterial symbiosis and its application in wastewater treatment ......................................................................................341 Inigo Johnson, Sudeeptha Girijan, Binay Kumar Tripathy, Mohammad Abubakar Sithik Ali and Mathava Kumar 15.1 Introduction .................................................................................................... 341 15.2 The symbiotic process.................................................................................... 342 15.2.1 Exchange of information in the form of bioactive compounds for symbiosis .............................................................................................. 343 15.2.2 Exudates that can inhibit the microbes in the vicinity........................... 345 15.2.3 Exudates that can stimulate the microbes in the vicinity....................... 346 15.2.4 Factors affecting symbiotic systems...................................................... 349 15.3 Applications in wastewater treatment............................................................ 351 15.3.1 Types of reactors .................................................................................. 351 15.3.2 Nutrients removal ................................................................................. 353 15.3.3 Metal removal ...................................................................................... 356 15.3.4 Organic matter removal ........................................................................ 357 15.3.5 Emerging contaminants removal ........................................................... 359 15.3.6 Removal of refractory compounds ........................................................ 361 15.4 Energy generation .......................................................................................... 362 15.4.1 Algal biohydrogen production .............................................................. 363 15.4.2 Algal lipid production ........................................................................... 363 15.4.3 Microbial fuel cell reactor using algal bacteria interaction.................. 364 15.5 Conclusion and future directions ................................................................... 365 References................................................................................................................. 366

Chapter 16: Role of plant growth promoting rhizobacteria in mitigation of heavy metals toxicity to Oryza sativa L. .........................................373 Vishnu Kumar, Gayatri Singh, Rajveer Singh Chauhan and Geetgovind Sinam 16.1 Introduction .................................................................................................... 373 16.2 Different genera of plant growth promoting rhizobacteria ......................... 374 16.3 Plant growth promoting rhizobacteria role in heavy metals dynamics in the soil ........................................................................................................ 377 16.4 Plant growth promoting rhizobacteria’s role in controlling pathogens in rice.............................................................................................................. 379 16.5 Plant growth promoting rhizobacteria in remediation of the environment.................................................................................................... 380 16.6 Conclusion and future prospect ..................................................................... 384 References................................................................................................................. 384 Further reading.......................................................................................................... 389

Contents xiii

Chapter 17: Study of transport models for arsenic removal using nanofiltration process: recent perspectives ............................................................391 Robin Marlar Rajendran, Sangeeta Garg and Shailendra Bajpai 17.1 Introduction .................................................................................................... 391 17.1.1 Sources ................................................................................................. 391 17.1.2 Health effects ....................................................................................... 393 17.1.3 Permissible limit ................................................................................... 393 17.2 Chemistry of arsenic ...................................................................................... 395 17.3 Methods of arsenic removal from water/wastewater .................................... 395 17.3.1 Membrane technology .......................................................................... 396 17.4 Nanofiltration of arsenic ................................................................................ 397 17.4.1 Modeling of nanofiltration membranes for arsenic removal .................. 397 17.5 Conclusion and future perspective................................................................. 401 Nomenclature............................................................................................................ 401 Greek symbols .......................................................................................................... 402 Abbreviations ............................................................................................................ 402 References................................................................................................................. 402 Further Reading ........................................................................................................ 405

Chapter 18: Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms .................................................................407 B. van Veenhuyzen, C. Ho¨rstmann, J. Peens and H.G. Brink 18.1 Introduction .................................................................................................... 407 18.2 Remediation of Pb.......................................................................................... 408 18.2.1 Conventional methods for Pb remediation or recovery ......................... 408 18.2.2 Bioremediation of Pb ............................................................................ 408 18.2.3 Mechanisms of bioremediation ............................................................. 409 18.3 Case study of aqueous Pb biorecovery by local Pb-resistant organisms...... 411 18.3.1 Characterization of bacterial consortia .................................................. 411 18.3.2 Precipitate identification ....................................................................... 412 18.3.3 Microbiological and kinetic study ......................................................... 413 18.3.4 Case studies of varied operating conditions .......................................... 417 18.4 Conclusion and outlook ................................................................................. 421 Acknowledgment ...................................................................................................... 422 References................................................................................................................. 422

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Contents

Chapter 19: Microbial bioremediation of azo dye through microbiological approach .......................................................................................425 Celia Vargas-de la Cruz and Daniela Landa-Acun˜a 19.1 Introduction .................................................................................................... 425 19.2 Classification of dyes ..................................................................................... 426 19.3 Role of environmental parameters on microbial biodegradation and bioremediation of azo dye.............................................................................. 427 19.4 Effects of environmental parameters on azo dye degradation...................... 432 19.4.1 Temperature ......................................................................................... 433 19.4.2 Oxygen ................................................................................................. 433 19.4.3 Dye concentration ................................................................................. 434 19.4.4 Electron donor ...................................................................................... 435 19.4.5 pH ........................................................................................................ 435 19.4.6 Dye structure ........................................................................................ 435 19.4.7 Redox potential..................................................................................... 436 19.4.8 Redox mediator .................................................................................... 437 19.4.9 Decolorization by genetically modified organisms ............................... 438 19.5 Conclusion...................................................................................................... 438 References................................................................................................................. 438 Further reading.......................................................................................................... 441

Chapter 20: Novel process of ellagic acid synthesis from waste generated from mango pulp processing industries ............................................443 Murugan Athiappan, Shantkriti Srinivasan, Rubavathi Anandan and Janani Rajaram 20.1 Introduction .................................................................................................... 443 20.1.1 Waste from mango pulp processing industries ...................................... 443 20.2 Composition of mango wastes ....................................................................... 444 20.3 Types of tannins ............................................................................................. 444 20.4 Bioconversion of tannin to ellagic acid......................................................... 445 20.5 Microbes involved in the production of ellagic acid .................................... 447 20.6 Applications of ellagic acid ........................................................................... 448 20.7 Conclusion...................................................................................................... 451 References................................................................................................................. 451 Further reading.......................................................................................................... 454 Index ............................................................................................................ 455

List of contributors Aakriti Aggarwal Department of Biotechnology, Dr. B. R. Ambedkar National Institute of Technology, Jalandhar, India Komal Agrawal Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Bandarsindri, Kishangarh, Ajmer, India Hafiz Adeel Ahmad Shandong Provincial Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shandong University, Shandong, P.R. China Shakeel Ahmad Department of Soil and Environmental Sciences, Muhammad Nawaz Shareef University of Agriculture, Multan, Pakistan Mohammad Abubakar Sithik Ali Environmental and Water Resource Engineering Division, Department of Civil Engineering, IIT Madras, Chennai, India Rubavathi Anandan Department of Microbiology, Periyar University, Salem, India Murugan Athiappan Department of Microbiology, Periyar University, Salem, India Shailendra Bajpai Department of Chemical Engineering, Dr B R Ambedkar National Institute of Technology, Jalandhar, India Shivarudrappa Bhairappanavar Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India Ankita Bhatt Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Bandarsindri, Kishangarh, Ajmer, India Sayantani Biswas Department of Biotechnology and Medical Engineering, National Institute of Technology Rourkela, Odisha, India H.G. Brink Department of Chemical Engineering, Faculty of Engineering, Built Environment and Information Technology, University of Pretoria, Hatfield, Pretoria, South Africa I. Chakraborty Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Venkatesh Chaturvedi SMW College, MG Kashi Vidyapeeth, Varanasi, India Rajveer Singh Chauhan Department of Botany, Deen Dayal Upadhyaya Gorakhpur University, Gorakhpur, India Sushma Chityala Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India Jayashankar Das Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India

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xvi

List of contributors

Veeranki Venkata Dasu Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering Indian Institute of Technology Guwahati, Guwahati, India Chirayu Desai PD Patel Institute of Applied Science, Charotar Institute of Science and Technology, Changa, India Neelam Devpura Environmental Genomics and Proteomics Lab, UGC-Centre of Advanced Study, Post Graduate Department of Biosciences, Sardar Patel University, Satellite Campus, Bakrol, India Sangeeta Garg Department of Chemical Engineering, Dr B R Ambedkar National Institute of Technology, Jalandhar, India M.M. Ghangrekar Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Sudeeptha Girijan Environmental and Water Resource Engineering Division, Department of Civil Engineering, IIT Madras, Chennai, India C. Ho¨rstmann Department of Chemical Engineering, Faculty of Engineering, Built Environment and Information Technology, University of Pretoria, Hatfield, Pretoria, South Africa Kunal Jain Environmental Genomics and Proteomics Lab, UGC-Centre of Advanced Study, Post Graduate Department of Biosciences, Sardar Patel University, Satellite Campus, Bakrol, India Dharanidaran Jayachandran Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India Inigo Johnson Environmental and Water Resource Engineering Division, Department of Civil Engineering, IIT Madras, Chennai, India Jenny Johnson Environmental Genomics and Proteomics Lab, UGC-Centre of Advanced Study, Post Graduate Department of Biosciences, Sardar Patel University, Satellite Campus, Bakrol, India Kanyaga Parameswari M Department of Zoology, Cotton University, Panbazar, Guwahati, Assam, India Sufia K. Kazy Department of Biotechnology, National Institute of Technology Durgapur, Durgapur, India Dinesh Kumar Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India Mathava Kumar Environmental and Water Resource Engineering Division, Department of Civil Engineering, IIT Madras, Chennai, India Raghawendra Kumar Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India Vineet Kumar Environmental Microbiology and Biotechnology Laboratory, School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India; Department of Environmental Microbiology, School of Environmental Sciences, Babasaheb Bhimrao Ambedkar (A Central) University, Lucknow, India Vishnu Kumar Plant Ecology & Climate Change Science Division, CSIR-National Botanical Research Institute, Lucknow, India Daniela Landa-Acun˜a Laboratory of Microbial Ecology and Biotechnology, Department of Biology, Faculty of Sciences, National Agrarian University La Molina (UNALM), Lima, Peru Datta Madamwar Environmental Genomics and Proteomics Lab, UGC-Centre of Advanced Study, Post Graduate Department of Biosciences, Sardar Patel University, Satellite Campus, Bakrol, India

List of contributors xvii Bapi Mandal Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering Indian Institute of Technology Guwahati, Guwahati, India Siranjeevi Nagaraj Nencki Institute of Experimental Biology, Polish Academy of Sciences, Warsaw, Poland Shou-Qing Ni Shandong Provincial Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shandong University, Shandong, P.R. China Priti Raj Pandit Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India Labdhi Pandya Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India Ajit Patel J. & J. College of Science, Nadiad, India Kamlesh Patel P. G. Department of Biosciences, Sardar Patel University, Vallabh Vidyanagar, India Radhika Patel P. G. Department of Biosciences, Sardar Patel University, Vallabh Vidyanagar, India Vanita Patel V.P & R.P.T.P. Science College, Vallabh Vidyanagar, India Zarna Patel Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India J. Peens Department of Chemical Engineering, Faculty of Engineering, Built Environment and Information Technology, University of Pretoria, Hatfield, Pretoria, South Africa Kasturi Poddar Department of Biotechnology and Medical Engineering, National Institute of Technology Rourkela, Odisha, India Ashish A. Prabhu Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India Anwesha Purkayastha Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering Indian Institute of Technology Guwahati, Guwahati, India Janani Rajaram Department of Microbiology, Periyar University, Salem, India Robin Marlar Rajendran Department of Chemical Engineering, Dr. B. R. Ambedkar National Institute of Technology, Jalandhar, India Rohit Rathour PD Patel Institute of Applied Science, Charotar Institute of Science and Technology, Changa, India Ajoy Roy Department of Biotechnology, National Institute of Technology Durgapur, Durgapur, India Mahesh Kumar Sah Department of Biotechnology, Dr. B. R. Ambedkar National Institute of Technology, Jalandhar, India Pinaki Sar Department of Biotechnology, Indian Institute of Technology Kharagpur, Kharagpur, India Angana Sarkar Department of Biotechnology and Medical Engineering, National Institute of Technology Rourkela, Odisha, India Debapriya Sarkar Department of Biotechnology and Medical Engineering, National Institute of Technology Rourkela, Odisha, India

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List of contributors

Jayeeta Sarkar Department of Biotechnology, Indian Institute of Technology Kharagpur, Kharagpur, India S.M. Sathe Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India Maulin P. Shah Industrial Waste Water Research Laboratory, Division of Applied & Environmental Microbiology, Enviro Technology Limited, Ankleshwar, India Geetgovind Sinam Plant Ecology & Climate Change Science Division, CSIR-National Botanical Research Institute, Lucknow, India Ajay Kumar Singh Department of Environmental Microbiology, School of Environmental Sciences, Babasaheb Bhimrao Ambedkar (A Central) University, Lucknow, India Gayatri Singh Plant Ecology & Climate Change Science Division, CSIR-National Botanical Research Institute, Lucknow, India Shantkriti Srinivasan Department of Biotechnology, Kalasalingam Academy of Research and Education, Krishnankoil, India Magapu Solomon Sudhakar Applied Biotechnology Department, Sur College of Applied Sciences, Ministry of Higher Education, Sur, Sultanate of Oman Indu Shekhar Thakur Environmental Microbiology and Biotechnology Laboratory, School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India Onkar Tiwari Department of Biotechnology, Ministry of Science & Technology, New Delhi, India Binay Kumar Tripathy Environmental and Water Resource Engineering Division, Department of Civil Engineering, IIT Madras, Chennai, India Ujjval Trivedi P. G. Department of Biosciences, Sardar Patel University, Vallabh Vidyanagar, India B. van Veenhuyzen Department of Chemical Engineering, Faculty of Engineering, Built Environment and Information Technology, University of Pretoria, Hatfield, Pretoria, South Africa Celia Vargas-de la Cruz Latin American Center for Teaching and Research in food bacteriology (CLEIBA), Faculty of Pharmacy and Biochemistry, Major National University of San Marcos (UNMSM), Lima, Peru Nishchay Verma Department of Biotechnology and Medical Engineering, National Institute of Technology Rourkela, Odisha, India Pradeep Verma Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Bandarsindri, Kishangarh, Ajmer, India

Preface This book describes the state-of-the-art and possibilities of emerging technologies in environmental bioremediation, and reviews its various areas together with their related issues and implications. Considering the number of problems that define and concretize the field of environmental microbiology or bioremediation, the role of some bioprocesses and biosystems for environmental protection, control, and health based on the utilization of living organisms are analyzed. The book aims to provide a comprehensive view of advanced emerging technologies with environmental approaches for wastewater treatment, heavy metal removal, pesticide degradation, dye removal, waste management, microbial transformation of environmental contaminants, etc. With advancements in the area of environmental bioremediation, researchers are looking for new opportunities to improve quality standards and the environment. Recent technologies have given an impetus to the possibility of using renewable raw materials as a potential source of energy. Cost-intensive and ecofriendly technologies for producing high-quality products and efficient ways to recycle waste to minimize environmental pollution are the needs of hour. The use of bioremediation technologies through microbial communities is another viable option to remediate environmental pollutants, such as heavy metals, pesticides, and dyes. Since physicochemical technologies employed in the past have many potential drawbacks including their high cost and low sustainability, efficient biotechnological alternatives to overcome increasing environmental pollution are needed. Hence, environment-friendly technologies that can reduce the pollutants causing adverse hazards to humans and the surrounding environment are required. Environmental remediation, pollution prevention, detection, and monitoring are evaluated by considering the achievements, as well as the future prospects, in the development of biotechnology. Various relevant topics have been chosen to illustrate each of the main areas of environmental biotechnology: wastewater treatment, soil treatment, solid waste treatment, and waste gas treatment, dealing with both microbiological and process

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xx Preface engineering aspects. The distinct role of emerging technologies in environmental bioremediation in the future is emphasized by considering the opportunities to contribute to new solutions and directions in the remediation of contaminated environments, as well as minimizing future waste release and creating pollution-preventing alternatives. To take advantage of these opportunities, innovative new strategies, which advance the use of molecular biological methods and genetic engineering technologies, are examined. These methods would improve the understanding of existing biological processes in order to increase their efficiency, productivity, and flexibility. Examples of the development and implementation of such strategies are included. Also, the contributions of environmental biotechnology to the progress of a more sustainable society are revealed. Editors Maulin P. Shah, India Susana Rodriguez-Couto, Spain ¨ r, United States S. Sevinc¸ Sengo ¸

CHAPTER 1

Immobilization of anaerobic ammonium oxidation bacteria for nitrogen-rich wastewater treatment Shou-Qing Ni1, Hafiz Adeel Ahmad1 and Shakeel Ahmad2 1

Shandong Provincial Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shandong University, Shandong, P.R. China, 2 Department of Soil and Environmental Sciences, Muhammad Nawaz Shareef University of Agriculture, Multan, Pakistan

1.1 Introduction Worldwide, water plays a central role in the effort toward the economic development. The demand for fresh water increases globally for the ever-growing population. A total of 3% water is available as a freshwater source and the remaining is salt water (Ellis, 2004). Significantly, the agriculture sector demanded about 70% of water worldwide, and predicted that the water consumption for the production of energy and the industrial process is projecting (WWAP, 2017). The growth of population in the developing world increased the demand for water. However, the water availability is directly linked to water quality since different human activities demand different qualities of water. Overall, it is estimated that over 80% of wastewater are directly discharged without an acceptable standard (WWAP, 2017). If this scenario remains unchecked, it will further threaten human health and limit the sustainable development especially in the developing countries. Hence all the above-discussed points necessitate the proper management of wastewater. Mostly coagulation (chemical and electrical), adsorption, and biological technologies were applied for the removal of different contaminants from wastewater (Ahmad, Lafi, Abushgair, & Assbeihat, 2016). Nitrogen is an essential component of different proteins and amino acids. Additionally, it also plays a vital role as plant nutrient and applied in the form of fertilizer to improve the food productivity. Nitrogen in the wastewater discharged directly into rivers, canals, and lakes can cause severe threats to the water bodies and human health. Eutrophication of the lake, river, and other water bodies is the primary outcome of the nitrogenous wastewater (Quan, Khanh, Hira, Fujii, & Furukawa, 2011). The concentration of ammonia as low as Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00001-8 © 2020 Elsevier Inc. All rights reserved.

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2

Chapter 1

0.01 mg/L can cause mortalities of the shrimp larvae and other pathological effects. The consumed nitrate is converted to nitrite within the human body, and further nitrite is reduced to cancer-causing N-nitroso compounds (IOWA Environmental Council, 2016). Hence it is necessary to remove nitrogen from wastewater to overcome the eutrophication, decrease the nitrous oxide emission, and avoid human health issues. Customarily, nitrogenous species are biologically removed from wastewater through the ordinary process of nitrification
denitrification (N
DN) (Dongen, 2001). On the other hand, the filtering technique was also reported for the removal of nitrogen. The carbon-to-nitrogen (C/N) ratio is essential for the successful application of the conventional process. However, some wastewater loaded with high ammonium concentration and lower organic matter increases the cost of treatment due to addition of artificial carbon source for the growth of heterotrophic denitrifying bacteria (Kimura, Isaka, & Kazama, 2011a; Kimura, Isaka, Kazama, & Sumino, 2010; Quan et al., 2011). So advancement in the nitrogen removal process is desired under the conditions of the lower C/N ratio. In the late 20th century, a new mean of anaerobic ammonium oxidation (anammox) was reported (Mulder, van de Graaf, Robertson, & Kuenen, 1995). Later studies proved anammox as an active biological pathway to treat nitrogenous wastewater (Jetten et al., 2001). Autotrophic growth behavior of anammox decreased the cost of treatment compared to conventional process due to exclusion of the external carbon sources. Generally, the anammox reaction produced dinitrogen gas with little amount of nitrate using nitrite and ammonium as a substrate (Mulder et al., 1995). Autotrophic nature of anammox growth, sluggish growth behavior, and lower cell yield increased the start-up period of the wastewater treatment process (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, & Zhang, 2015; Tang et al., 2011). So, for the slow-growing anammox, more diversifying techniques such as the design of the reactor, restriction of cell mass by immobilization, and proper functioning conditions are needed (Bae et al., 2015), which will be helpful for the reduction of anammox initiation time and avoiding the washout of the precious microbial biomass. Suspended growth is advantageous because of suitable distribution of substrate in the reactor for the effective reaction (Bae et al., 2015). However, the major challenge in the suspended growth wastewater treatment is to retain the biomass in the reactor (Magrı´, Vanotti, & Szo¨gi, 2012). Plodding growth of anammox makes the cell vulnerable to washout from the reactor (Zhu, Yan, & Hu, 2014) due to N2 gas bubble, which causes the floating of biomass (Quan et al., 2011). Thus it is necessary to lessen the loss of slowgrowing anammox and to make anammox a dominate bacteria in the reactor microbial community (Chen et al., 2016) for sustainability of the removal process. Cell immobilization is a new technique that has attracted the scientific community as a competitive alternative to avoid cell washout from the reactor (Quan et al., 2011). The immobilization can protect microorganisms from inhibitory effects of the substrate as well as products. The techniques practiced for the immobilization of anammox cell include development of granules by natural process (Dapena-Mora, Campos, Mosquera-Corral, Jetten, & Me´ndez, 2004), attachment of anammox microbial cell on the exterior of carrier

Immobilization of anaerobic ammonium oxidation bacteria 3 (Biofilm) (Ni, Lee, Fessehaie, Gao, & Sung, 2010; Tsushima, Ogasawara, Kindaichi, Satoh, & Okabe, 2007), and entrapment of anammox biomass inside different gel carrier (Furukawa et al., 2009; Isaka, Date, Sumino, & Tsuneda, 2007). On contrary to the other methods of cell immobilization, entrapment of microbial cell inside the gel carrier is advantageous in terms of shorter start-up time (Ali et al., 2015; Bae et al., 2015), maintaining the proper density of cell, and protection against the unfavorable environment (Hsia, Feng, Ho, Chou, & Tseng, 2008; Zhu, Hu, & Wang, 2009). Also immobilization of microbial cell is preferred because of consistent and effective biomass retention (Bae et al., 2015). So immobilization of microbial biomass was found to have a greater potential to improve the performance of anammox bacteria for wastewater treatment. In the developed countries, such as United States and Japan, N
DN with gel immobilization have already been practicing (Isaka et al., 2007). This chapter (1) highlights the opportunities and challenges in using cell immobilization for enhancing the efficacy of wastewater nitrogen removal by anammox bacteria, (2) discloses the existing technologies for the immobilization of anammox and examines the sustainability of different gel carrier, (3) discusses the application of anammox with immobilization technology, and (4) highlights the bottlenecks to commercialize this technology in the lower income country.

1.2 Anammox bacteria and their metabolic process Anammox bacteria belong to phylum Planctomycetes and metabolized very slowly (Strous et al., 2006). Previously it was assumed that anammox oxidize ammonia with nitrite as an essential electron acceptor. Later it is reported that anammox could use alternative electron acceptor rather than nitrite, which confirmed that anammox have versatility in their metabolism (Liu et al., 2008). Guven et al. (2005) described that organic acids (propionate) have competition with ammonium for the nitrite as an electron acceptor. In the absence of nitrite or nitrate, anammox sludge failed to oxidize propionate, and propionate and ammonium are independent of each other. The later study proposed the name of that novel anammox as Anammoxoglobus propionicus (Kartal et al., 2007). With the freshly discovered anammox strain, it concluded that the anammox process has a comprehensive application for the removal of nitrogen and organic acid from wastewater. It was also stated that iron could be oxidized with nitrite as the electron acceptor (Strous et al., 2006). 2 1 Fe1 2 1 NO2 -Fe3 1 N2

(1.i)

Another study postulated the anaerobic realization of elemental sulfur and nitrogen gas from ammonium and sulfate (Fdz-Polanco, Fdz-Polanco, Garcia, Villaverde, & Uruen, 2001), but the exact mechanism of this conversion was missing. A few years later, purified strain by percoll density gradient centrifugation was applied for the removal of ammonium and sulfate,

4

Chapter 1

completely autotrophic removal of ammonium and sulfate was observed with pure culture and phylogenetic analysis confirmed that new strain had 92% similarity with Planctomycetes, provisionally named as Anammoxoglobus sulfate (Liu et al., 2008). A. sulfate used sulfate as an electron acceptor with ammonium as an electron donor to produce nitrite and elemental sulfur, when nitrite concentration becomes higher, traditional anammox process started which cause the removal of remaining ammonium and nitrite (Liu et al., 2008). Chemical mechanism of autotrophic sulfate reducing ammonium oxidation reaction is given in Eq. (1.ii), and further traditional anammox process removes the remaining ammonium with nitrite as electron acceptor as shown in Eq. (1.iii) (Liu et al., 2008). 1 2 SO22 4 1 NH4 -NO2 1 S 1 2H2 O

(1.ii)

2 NH1 4 1 NO2 -N2 1 2H2 O

(1.iii)

So in general anammox changes their strategy as per the environment. The metabolism of anammox is environment dependent and is governed by the present pollutants. The contaminant is ammonium, propionic acid, urea, or sulfate, which show the versatility of the metabolic process in the anammox species. The detailed mechanism of anammox metabolism was explored by Kartal et al. (2011) as shown in Fig. 1.1. The nitric oxide and hydrazine are the central intermediate while the nitrate

Figure 1.1 The moleculaisms of anammox metabolism (Kartal et al., 2011) included Nar enzyme, nitrate reductase; NirS, nitrite reduction; HZS enzyme, hydrazine synthesis; and HDH enzyme, hydrazine dehydrogenase.

Immobilization of anaerobic ammonium oxidation bacteria 5 is the byproduct of the anammox process. However, the mechanism and involved protein for the production of nitrate is an open question. It is proposed that the protein or enzyme present on the anammoxosome membrane or inside the anammoxosome are responsible for the process of nitrite in some other nitrogen compounds that may use NirS to produce nitric oxide while the remaining goes to Nar enzyme which produces nitrate. The production of nitrate is necessary for the production or consumption of extra electron to keep the electron balance.

1.3 Cell immobilization: a strategy to improve microbial wastewater treatment 1.3.1 What is cell immobilization? Artificial technique for the immobilization of cells is defined as follows: Cell immobilization is defined as the confinement of microorganisms within the specific material to enhance the time course of the process and make sure the presence of a microbial cell in the system. Leenen, Dos Santos, Grolle, Tramper, and Wijffels (1996)

Through immobilization, the retention time of the cell can be increased and eventually amplified the process efficiency. After the discovery of anammox, research was dedicated to address the washout problem of anammox. Different strategies were practiced ranging from the natural process of granulation, biofilm to artificial gel immobilization techniques. Gel immobilization seems to be a more feasible option to sustain the suitable amount of biomass in the reactor. The schematic diagram describing the immobilization of the cell is depicted in Fig. 1.2, where bacterial cell is fixed at the inside of the beads artificially.

1.3.2 Different approaches for cell immobilization In general, the immobilization of microorganisms can be divided into three methods (1) granulation, (2) natural adsorption to a support matrix (Klein & Ziehr, 1990), and (3) entrapment into cross-linked polymers. This means that the microorganism itself has the property to adhere on the solid surface, and this happens as long as the conditions are appropriate. Different approaches of immobilization are depicted in Fig. 1.3 with gel immobilization in detail. 1.3.2.1 Granulation In a particular case, the microorganisms in biological treatment systems are clustered together and aggregated with each other, which can serve as a support matrix and form particles with better properties (Pijuan, Werner, & Yuan, 2011; Verawaty, Pijuan, Yuan, & Bond, 2012). By different electron receptors in the metabolic process, granulation can be

6

Chapter 1

Figure 1.2 Schematic description of cell immobilization.

Figure 1.3 Schematic description of different approaches for cell immobilization.

Immobilization of anaerobic ammonium oxidation bacteria 7 divided into anaerobic granular sludge and aerobic granular sludge (De Bok, Plugge, & Stams, 2004). Compared with suspended sludge, both kinds of mature sludge particles have better settling ability because of the high density of biomass. The granules have a clear spherical structure with a diameter of 0.14
5 mm and fast settling speed of 18
100 m/h (Schmidt & Ahring, 1996). The granules settling ability is closely related to the particle size and density, while the surface hydrophobicity of granules is also higher than the activated sludge (AS) (Li et al., 2013; Liang, Gao, & Ni, 2017; Zheng, Yu, & Sheng, 2005). The granulation also leads to a variety of strains, which means that the coexistence of aerobic and anaerobic structure decreased specific surface area and concentrated carbohydrates and proteins in the extracellular polymeric substances and better tolerance to various types of high-strength and toxic wastewaters (Hamza, Iorhemen, & Tay, 2016; Liu, Chan, & Fang, 2002; Liu, Kang, Li, & Yuan, 2015; Lourenc¸o et al., 2015; Zheng et al., 2005; Zhu, Jin, Lin, Gao, & Xu, 2015). However, the factors including the size of the granule, the growth of nitrite oxidizing bacteria (competition with anammox for nitrite), and the decay of pellets hamper the application of this process. 1.3.2.2 Biofilm formation Microorganisms have property to adhere on the surface of the solid, and this happens as long as the conditions are appropriate. Generally in the combined nitration
anammox biofilm process, ammonium oxidizing bacteria (AOB) existed on the water biofilm interface while anammox presented deep inside the biofilm. AOB consumed the oxygen from the liquid and ultimately protected the anammox from the oxygen inhibition (Nielsen et al., 2018). Furthermore the slow growth bacteria can be retained in the reactor due to the attachment on the surface of the carrier to form the biofilm. The materials used to support the biofilm growth of bacteria included nonwoven fibers and porous carrier. Ahmad et al. (2017) studied the development of biofilm and reported that from the carbohydrate, protein amid and deoxyribonucleic acid/ribonucleic acid (DNA/RNA) were the responsible factors for the growth of biofilm. The size of the carrier is an essential parameter for the development of biofilm. Carrie size of 15 mm exhibited an excellent result compared to large and small carrier size (Ahmad et al., 2017). The outer 10 mm layers provide a suitable condition for the growth of AOB, and inner 5 mm zone is ideal for the anammox growth. 1.3.2.3 Gel entrapment Above the exploration of natural phenomena to attach the bacteria on the specific surface or making the clump of biomass, there is also an artificial approach to immobilize the biomass. The biomass washout in flocculent condition is indispensable and need to solve this issue. The incorporation of bacterial biomass within the gel material is a new and

8

Chapter 1

useful approach (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). The microbial cell and enzyme incorporation by different gel materials without affecting their activity increase the application potential of immobilization techniques in the wastewater treatment. This incorporation technique effectively avoided biomass washout and enhanced microbial activity (Bae et al., 2015; Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). The artificial incorporation technique is a simple and effective approach. Different natural and artificial materials were applied for the immobilization of anammox. The natural materials are easy to biodegrade due to organic in nature while synthetic materials are resistant to the biodegradation. So in the wastewater treatment, artificial materials are more attractive because of their stability and excellent durability against the shock loading. The detail of the well-reported materials is given in the following section.

1.4 Why is gel immobilization advantageous? The biofilm formation and granulation process based on the exploration of the natural properties of microorganisms have some demerits. Small granules may cause the saturation of oxygen in single-step PN-anammox process; nevertheless large granules limit the transfer of substrate in the core of pellets. The gas pocket developed inside the granules may cause the floating of granules and ultimately leads toward the lower efficiency of the system (Song et al., 2017). Similarly, the properties of carriers, especially surface properties would affect the development of biofilm as well as the thickness of biofilm and process efficiency. Large carrier size reduces the penetration of substrate deep inside the biofilm, and small carrier size causes the detachment of biofilm under the influence of large shear force (Ahmad et al., 2017). The biomass detachment from the specific surface under limited conditions reduces the process efficiency. The increase in the shear stress decreases the nitrogen removal efficiency due to the obliteration of the anaerobic zone and ultimately washout of anaerobic microbes caused by the detachment under the influence of higher shear stress. Therefore in the wake of above-described loopholes of granulation and biofilm technologies, gel immobilization seems a better alternative approach to shorten the start-up time of anammox. The comparison of the start-up period of different methods is depicted in Table 1.1. The retention of biomass was enhanced by immobilization of anammox within the gel material even at a short hydraulic retention time. The liquid and solid separation can reach, which reduces the chance of biomass washout from the system. Further the stirring can remove the formation of the gas pocket inside the beads. The gel cubes are easy to handle for further investigation of anammox characteristics such as anammox identification and visual observation of the anammox process (Isaka et al., 2007). Immobilization of anammox by different gel material is discussed in detail.

Immobilization of anaerobic ammonium oxidation bacteria 9 Table 1.1: Comparison of the start-up time between different immobilization techniques. Immobilization techniques

Start-up time Origin of Sludge the sludge concentration (days)

Removal Dominant efficiency strain

Removal rate (kg N/m3/d) Reference

Granulation

Enriched anammox sludge

2124 mg VSS/L

62

88%

NA

NA

Granulation

Anaerobic granular sludge

93900 mg/L MLSS

63

NA

1.16

0.72 g (dry SS)

20

NA

Candidatus jettenia, Brocadia, and kuenenia NA

3.7

Isaka et al. (2007)

0.33 g VSS/L

35

NA

NA

10.8

Ali et al. (2015)

2000 mg/L (MLVSS)

70

91%

NA

0.19

Gel entrapment Activated sludge

25.3 g/L (TSS)

25

89%

1.12

granulation

1700 mg/L (MLVSS)

24

81%

Candidatus Brocadia sinica NA

Lu, Ma, Ma, Shan, and Chang (2017) Bae et al. (2015)

Gel entrapment Enriched anammox sludge Gel entrapment Enriched anammox sludge Biofilm N/A

Enriched anammox biomass

NA

Li, Xu, Shao, Zhang, and Yang (2014) Wang et al. (2017)

Qiao, Kawakubo, Cheng, and Nishiyama (2009)

1.5 Gel materials used for the immobilization of anammox Nitrogen removal activities by immobilized cells were reviewed. It was concluded that anammox biomass immobilizing in the gel carrier were protected from colloidal solids present in the wastewater (Magrı´ et al., 2012). Diverse synthetic polymers including waterborne polyurethane (WPU) (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015), polyethylene glycol (PEG) (Isaka et al., 2007), and polyvinyl alcohol (PVA) (Magrı´ et al., 2012) mixed media of PVA/sodium alginate (SA) (Hsia et al., 2008) have been applied for the entrapment of anammox to reduce the inhibition of nitrite and best substitute to avoid the washout problems (Isaka, Date, Sumino, Yoshie, & Tsuneda, 2006; Magrı´ et al., 2012). Optimal temperature range for the immobilized biomass can be extended. Moreover microbial cells can easily separate from inhibitory substances in the surrounding environment.

10

Chapter 1

1.5.1 Polyvinyl alcohol and polyvinyl alcohol/sodium alginate PVA/SA was used for entrapment of microbial cells from the past decades. The carboxyl group of SA and the hydroxyl group of PVA would develop bonding which made the rugged structure of the carrier (Zhu et al., 2009). Different concentration of prepolymer and reinforcing procedure of gel carrier affect the micropores of the beads (Quan et al., 2011) and the exchange of the substrate between solid and liquid phases. Cells can be entrapped within the PVA by two mechanisms. The dissimilarity between these two methods is in the second stage of immobilization, which involved the hardening of beads, accomplished by freezing or chemical means (Magrı´ et al., 2012). The chemical way of immobilization is more sustainable than freezing method especially when the process temperature is adjusted above 30 C. However, the duration and concentration of crosslink solution affect the stiffness of the beads and ultimately process efficiency. It was reported that the insertion of SA improved the substrate transfer and microspores structure of the PVA beads (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). Most reported solutions used for the crosslink included calcium chloride (Ali, Oshiki, & Okabe, 2014; Zhu et al., 2009), calcium chloride/sodium nitrate (Hsia et al., 2008; Quan et al., 2011), and calcium chloride/boric acid (Bae et al., 2015). It was reported that anammox activity was less affected when calcium chloride was used as crosslink solution (Zhu et al., 2009). Conversely, the crosslinking time and the concentration of the solution are crucial factors if it consists of boric acid, which may increase the time of process start-up and stability of the beads. The porous structure of PVA/SA confirmed by the SEM micrograph facilitated the transformation of the substrate into the gel beads (Quan et al., 2011; Zhu et al., 2009). Microscopic observation pronounced the smoothed surface of PVA beads with a diameter of about 3
4 mm (Hsia et al., 2008). Entrapment of anammox by PVA and SA reported from a few years back (Hsia et al., 2008; Zhu et al., 2014). PVA of 4%
6% considered the most preferred concentration to immobilize the anammox (Han, 2016).

1.5.2 Waterborne polyurethane There is another material that is reportedly applied for the immobilization of anammox. A comparative study was conductive to verify the mechanical strength, stability, and long-term application of different materials including WPU, PV, PVA/SA, and SA (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). It was concluded that WPU has the best stability and mechanical strength compared to other synthetic and natural materials (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). WPU revealed good mechanical stability compared to PVA, PVA/SA, and SA, which showed the clear superiority of WPU for the immobilization of microbes. As for the concern of other materials such as carboxymethyl cellulose (CMC), PVA, and PEG, WPU also has

Immobilization of anaerobic ammonium oxidation bacteria 11 superiority in terms of mechanical stability (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). The apparent dissimilarity in the mechanical stability of different materials is due to the variance in the properties of immobilized materials and the way of crosslinking of immobilized granules (Chen et al., 2016). WPU, when compared with PEG, have a different chemical structure and relative molecular mass. The WPU have small space in their three-dimensional structure, which make WPU less absorbance of water during swelling and ultimately increase the mechanical strength of immobilized granules (Chen et al., 2016). Another significant parameter included the absence of the spherical shell on the WPU granules, which increases the mass transfer property of WPU granules (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). Moreover, WPU immobilized anammox have lower half-saturation constant (Ks), which increases the substrate affinity of immobilized granules and aid the anammox process in the competitive environment (Chen et al., 2016). During the continuous experiment, it was reported that WPU has an excellent sludge retaining ability and good mechanical strength (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). WPU provided a suitable environment for the proliferation of anammox inside the gel carrier, which is the strong evidence for the increased of anammox activity inside the gel carrier (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015).

1.5.3 Polyethylene glycol gel Anammox entrapped in the PEG prepolymer was also used for the removal of nitrogen from wastewater. The colorless PEG material makes an easy characterization of anammox in the changing of the beads color. In the initial stage of the experiment, the small light brownish red aggregates can be seen and later the whole cube turned into bright red (Isaka et al., 2007). The different aspects of anammox were examined under the immobilization of anammox within the PEG. The performance of immobilized bacteria depends on the distribution of ammonium inside the gel material. The calculation of ammonium distribution coefficient was carried out according to the equation reported in the study of Sumino, Nakamura, Mori, and Kawaguchi (1992). 0

C V o Cso 2 Cs1 Kp 5 s 5 Cs V 1 2 Vo Cs1

(1.1)

where Kp represents the coefficient, Vo and V1 portray the volume of solution before and after the addition of gel cubes, Cso and Cs1 are the ammonium concentration before and after the addition of gel cubes. So, it is all about the brief of gel material used in the immobilization of anammox. The ammonium distribution coefficient of 2.2 is ideal for the negatively charged PEG (Sumino et al., 1992).

12

Chapter 1

1.6 Application of cell immobilization in anammox and partial nitrification 1.6.1 Application of immobilized anammox The full-scale application of immobilized anammox was reported only with PEG (Isaka, Kimura, Matsuura, Osaka, & Tsuneda, 2017). However, others materials also had great potential to upgrade the immobilization to pilot scale and then full-scale application. The laboratory-scale study and the feasible option of anammox application in the future were disclosed in this section. The immobilization can effectively achieve the rapid start of anammox by alleviating the washout of slow growth anammox bacteria (Ali et al., 2015). Anammox bacteria were sheltered from the surrounding and the loss of anammox cell from the reactor was avoided by immobilizing in the PVA/SA (Hsia et al., 2008). The rapid startup of anammox process was characterized by the conditions of immobilization, initial higher concentration of anammox sludge, and retaining of microbial biomass due to entrapment (Quan et al., 2011). Recently AS immobilized within the PVA/SA gel beads was applied to initiate the anammox process, specifically in the countries with the absence of full-scale anammox treatment process (Cho, Choi, Jeong, Lee, & Bae, 2017). The anammox activity with precultured anammox bacteria (PAB) started very shortly after the operation of the system, but the activity of anammox started after 93 days with the AS. The PAB reached an average nitrogen removal rate (NRR) of 0.35 kg N/m3/d in the last phase of the operation, at the same time the NRR of 0.36 kg N/m3/d with AS was achieved with immobilization techniques. The bacterial composition on the base of 16S rRNA of both reactor was entirely different during the start-up period, but at the end of the operation, the anammox bacterial population was comparable in both reactors (Cho et al., 2017). Another point of importance for the application of anammox in the place where there is no full-scale application is to transfer the anammox sludge to the long distance. It was reported that anammox immobilized and stored at the temperature of 28 C for 17 h recovered their activity very quickly (Magrı´ et al., 2012). Preserved anammox at room temperature and transportation of PVA/SA immobilized sludge is another feasible option to promote the application of anammox (Ali et al., 2014). Therefore the immobilized anammox can be transferred to a long distance without the fatal damage of their activity. Immobilized anammox within the PVA
SA strategically increased the nitrogen loading rate from 0.32 6 0.04 to 1.26 6 0.04 within three phases (cell lysis phase, anammox start-up phase, and anammox stability phase) but failed to achieve the reported stoichiometry ratio probably due to AOB, which can be adopted lower dissolved oxygen or from the oxidant (super oxides or hydroxyl radical) resulted from the enrichment process of biomass (Bae et al., 2015).

Immobilization of anaerobic ammonium oxidation bacteria 13 After the lag phase, anammox performance increased and an average nitrogen conversion rate of 3.4 kg N/m3/d was achieved from 0.2 kg N/m3/d after 60 days, which indicated the growth of anammox in the carrier (Isaka et al., 2007). Furthermore the effect of initial biomass concentration was also verified and the start-up time was reduced by increasing the biomass concentration in the gel carrier, which indicated that the concentration of 1.4% w/v dramatically reduced the start-up time and 11 kg N/m3-carrier nitrogen conversion rate was achieved within 25 days, which was very short compared to 62 and 39 days at the concentrations of 0.24 and 0.52% w/v, respectively (Isaka et al., 2007). The color of PEG beads changed from light brownish red to bright red, confirming the growth of anammox in the gel carrier (Isaka et al., 2007). Compared to nitrifying bacteria, anammox bacteria can also grow in the core of bead (Isaka et al., 2007). Successful operation of anammox wastewater treatment depends on the activity of anammox biomass. Terminal restriction fragment length polymorphism analysis showed two peaks (40 and 285 bp) characterizing HPT-WU-N01 clone and HPT-WU-A01 clone, and clone HPT-WUA01 became the dominant anammox population in PEG prepolymer carrier (Date, Isaka, Sumino, Tsuneda, & Inamori, 2008). It was observed that anammox population highly correlated with anammox activity (Date et al., 2008). Up to 2 mg/L Zn, Ni, Co, and Cu did not affect the anammox activity but as the concentration increased, anammox activity showed a decreasing trend. The effects of Zn, Ni, Co, and Cu metals on anammox activity were reversible, while the effects of Mo were irreversible (Kimura and Isaka, 2014). The addition of methanol irreversibly inhibited the immobilized anammox (Isaka et al., 2008). The nitrogen gas bubbles formation is another point of consideration to avoid the washout of the sludge. Nitrogen gas bubbles were observed on the exterior of the carrier, but stirring can remove the gas from the carrier peripheral and settle down the carrier (Isaka et al., 2007). Salinity level of the wastewater can also affect anammox performance. It was reported that anammox tolerance to salt concentration increased with PVA immobilized sludge (Han, 2016). The concentration of inorganic carbon (IC) was directly proportional to the anammox activity immobilized by PEG (Kimura, Isaka, & Kazama, 2011b). As the concentration of IC increased, the activity of anammox also increased. Zhu et al. (2014) reported that nitrogen removal efficiency of immobilized anammox was higher when compared with suspended growth. Further details of different studies, objectives of the studies, and results with immobilization technology of anammox are given in Table 1.2.

1.7 Commercialization of immobilizing technology The immobilization technology for the removal of nitrogen practiced in modern countries like Japan and United States was reported (Isaka et al., 2007). However, the application of immobilization technology in the lower income countries is lacking.

Table 1.2: The studies of immobilized anammox for evaluation of the different factors affecting the anammox activity.

Material

Packaging ration

Reactor type

Reactor volume

PEG

30%

CSTR

1L

PVA

20%

CSTR

1.4 L

PVA
SA

40%

ASBR

2L

PEG

20%

CSTR

PEG

20%

CSTR

500 mL

PEG

20%

CSTR

500 mL

Objectives

Remarks

To evaluate the entrapped Comparatively lower nitrogen anammox activity removal achieved, which can be increased by further optimization of different conditions Anammox perform well Usefulness of anammox inside the gel beads biomass immobilized by Successful recovery of nitrite PVA to ammonium ratio with the Steadiness of anammox application of swine system using synthetic wastewater as well as partially nitrified swine wastewater Achieved stable performance To check the effects of of anammox, broadened pH various pH, HRT, and range, HRT was 25 days, and temperature on the temperature was 35 to immobilized anammox achieve a stable nitrogen To check the nitrogen removal removal performance Anammox activity affected by To scrutinize the influent the DO concentration higher DO for the than 2.5 mg/L but the effects stable anammox process are reversible To evaluate the inhibition Anammox activity was affected by the higher level of heavy and reversibility of metals, but the effects of Zn, anammox activity under different concentration of Ni, Cu, and Co were reversible, and the impact of Mo was heavy metals irreversible To examine the impact of Impact of nitrite was reversible but the time to recover the nitrite on anammox anammox activity depended on activity the duration of nitrite inhibition

Level of study and Reported value duration

References

NRR of 3.4 kg N/m3 reactor/d

Continuous test for about 100 days

Isaka et al. (2007)

NCE of 73% with partially nitrified swine wastewater

Continuous test for about 160 days

Magrı´ et al. (2012)

NCR of Continuous 0.58 kg N/m3/d test about 90 days

Zhu et al. (2014)

NCR 2.9 kg N/ m3/d

Continuous test

Kimura et al. (2011a)

NCR of 4.0 kg N/m3/d with Zn, Ni, Cu, and Co concentration below 2 mg/L ACR 1.8 kg N/ m3/d

Batch test

Kimura and Isaka (2014)

Continuous test

Kimura et al. (2010)

ARR of 0.455 kg/m3/L

WPU, PEG, PVA, and CMC

Not Serum 500 mL mentioned bottles as reactor

To check the stability of anammox for long-term operation immobilized by different gel carrier

Compared with all other materials, WPU was more stable in term of mechanical stability, shock loading, and anammox activity

WPU, SA, PVA, PVA
SA

5

Serum bottle as reactor

500 mL

In terms of mechanical stability NCR of 80.98% Continuous test with and shocked loading, WPU WPU, was the suitable material 100 days compared to PVA, SA, and PVA
SA

PEG

20%

CSTR

500 mL

To compare the mechanical stability, effects of shock loading on the performance of immobilized anammox by different material Evaluation of the different concentrations of methanol and ethanol on anammox activity

PVA

Capsule 1L 100 bioreactor capsule per reactor

Continuous test for about 100 days

PVA
SA

70%

Upflow column reactor

10 mL

To evaluate the anammox performance by immobilizing within the novel PVA capsule and nitrite inhibition Preservation of anammox and reactivation of anammox biomass immobilized by PVA
SA

PVA
SA

70%

Upflow column reactor

10 mL

PVA
SA

20%

Conical flask

1L

Anammox has more sensitivity for the methanol compared to ethanol. Methanol inhibited anammox activity irreversibly Improvement in the anammox process achieved, which showed the stability of a novel capsule bioreactor

Preservation at room temperature and entrapment by PVA
SA, anammox reactivated their activity very rapidly Achievement of the anammox To optimize the process at the concentration of concentration of anammox in the gel beads 0.33 g VSS/L to fasten the start-up of the anammox process On the base of the To retain the anammox biomass and evaluation of stoichiometric value and nitrogen removal process, it anammox nitrogen concluded that anammox removal as well as retained in the gel beads stoichiometric value successfully

NCR of 2.9 kg N/m3/d. Before the application of methanol Not reported

Continuous test with WPU, 65 days

Batch test

Chen, Li, Deng, et al.( 2015); Chen, Li, Tabassum, et al. (2015) Chen, Li, Deng, et al. (2015); Chen, Li, Tabassum, et al. (2015) Isaka et al. (2008)

Chou, Tseng, and Ho (2012)

NRR of 7 kg N/ Batch test m3/d within 35 days

Ali et al. (2014)

NRR 7 kg N/ m3/d

Batch test

Ali et al. (2015)

Not reported

Continuous test

Hsia et al. (2008)

(Continued)

Table 1.2: (Continued) Level of study and Reported value duration

Packaging ration

Reactor type

Reactor volume

Objectives

Remarks

Na-CMC, PVA, SA, and PVA
SA

20%

Conical flask

250 mL

To optimize the characteristics of different material for the successful application of the anammox process

PEG

30%

CSTR

1L

To evaluate the change in the anammox population in the coexisting environment

PVA
SA

18.5%

CSTR

14 L

PEG

20%

To validate the activated sludge as a potential source of the anammox process and long-term constancy To verify the effects of inorganic carbon on the activity of immobilized anammox

Batch test Only PVA, SA was unfit for the Not reported application of anammox and Na-CMC have good transferability, but have inferior long-term stability was inferior. The mixture of PVA
SA was a suitable option for the successful application Not reported Continuous On the base of real-time PCR test result, anammox bacteria associated with clone HPTWU-N03 were the dominant population Anammox bacteria enriched by 0.36 kg N/m3/d Continuous using activated sludge. test about Anammox activity starts after 1.5 year 93 days

Material

500 mL

The decrease in the anammox activity observed under low concentration of inorganic carbon. The higher nitrogen conversion rate observed at 60 mg/L inorganic carbon

NCR of 7.0 kg N/m3/d

Continuous test

References Zhu et al. (2009)

Date et al. (2008)

Cho et al. (2017)

Kimura et al. (2011b)

ACR, Ammonium conversion rate; ARR, ammonium removal rate; CMC, carboxymethyl cellulose; DO, dissolved oxygen; NCE, nitrogen conversion efficiency; NCR, nitrogen conversion rate; NRR, nitrogen removal rate; PVA, polyvinyl alcohol; SA, sodium alginate; WPU, waterborne polyurethane.

Immobilization of anaerobic ammonium oxidation bacteria 17 To apply anammox immobilization technology on a large scale in lower income countries, some barriers need to be addressed. For instance, the capital cost of such a considerable quantity of beads, the project, and operation cost should be considered carefully. A commercial scale application needs a large number of gel beads about 20
1000 m3 (Sumino et al., 1992). Additionally further study about the substrate diffusion inside the bead, bead dissolution behavior, and life of the beads in the full-scale application are also required. Although immobilization shows good performance on the laboratory-scale, the large-scale application for an extended period has some other limitations. It was reported that the PVA is not equally distributed in the water under lower temperature (Tacx, Schoffeleers, Brands, & Teuwen, 2000). Thus higher temperature is needed for equal distribution of PVA into water for stable beads formation, which increases the cost and the risk of human safety. The higher temperature is needed for dissolution of PVA solution, which can burn the skin because of fallen drop. Self-condensing and water swelling properties of several hydroxyl groups in PVA (Tacx et al., 2000) may cause the adherence of PVA granules with each other and influence the mass transfer (Chen, Li, Deng, et al., 2015; Chen, Li, Tabassum, et al., 2015). Therefore these factors limit the application of PVA
SA gel beads on a large scale. The dissolution of PVA is a matter of time but possibly the time of decay can be extended to a long period. Moreover, PEG with different molecular weight is available in the market with different price. Using higher molecular weight of PEG increases the capital cost and seems difficult in the lower income country or developing country for adopting the application. Another critical point of concern to the environment is the disposal of the gel material. The start-up of the experiment depends on the packaging ratio as well as the sludge inside of the beads (Isaka et al., 2007). Increasing the concentration of sludge inside the beads may decrease the stability of the process due to the growth of bacteria. The decrease of the stability of the material and ultimate breakage of the gel beads also raise the question about the disposal of these beads.

1.8 Conclusion This chapter discussed about different immobilization technologies (natural and artificial) in the application of anammox. Also, the materials employed for the immobilization of anammox. The full-scale application of anammox with immobilization had been practiced by using PEG material. However, the research on the PVA
SA is still at infancy stage and need further exploration. Although the material and methods are similar in the process of immobilization, the concentration of sludge was different, which causes the difference in the conclusion. Overall immobilization is the technique that stops the washout of anammox bacteria. However, the full-scale application of immobilization technology in the developing country needs to solve some questions. Further disposal of the beads is an important concern in terms of environment and need environment study.

18

Chapter 1

Acknowledgments The authors acknowledge the support from the National Natural Science Foundation of China (21777086), Natural Science Foundation for Distinguished Young Scholars of Shandong Province (JQ201809), and Shandong Provincial Water Conservancy Research and Technology Promotion Project (SDSLKY201802).

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11024. Ahmad, M., Liu, S., Mahmood, N., Mahmood, A., Ali, M., Zheng, M., & Ni, J. (2017). Effects of porous carrier size on biofilm development, microbial distribution and nitrogen removal in microaerobic bioreactors. Bioresource Technology, 234, 360
369. Available from https://doi.org/10.1016/j.biortech.2017.03.076. Ali, M., Oshiki, M., & Okabe, S. (2014). Simple, rapid and effective preservation and reactivation of anaerobic ammonium oxidizing bacterium “Candidatus Brocadia sinica. Water Research, 57, 215
222. Available from https://doi.org/10.1016/j.watres.2014.03.036. Ali, M., Oshiki, M., Rathnayake, L., Ishii, S., Satoh, H., & Okabe, S. (2015). Rapid and successful start-up of anammox process by immobilizing the minimal quantity of biomass in PVA-SA gel beads. Water Research, 79, 147
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CHAPTER 2

Accelerated bioremediation of petroleum refinery sludge through biostimulation and bioaugmentation of native microbiome Jayeeta Sarkar1, Ajoy Roy2, Pinaki Sar1 and Sufia K. Kazy2 1 2

Department of Biotechnology, Indian Institute of Technology Kharagpur, Kharagpur, India, Department of Biotechnology, National Institute of Technology Durgapur, Durgapur, India

2.1 Introduction In the modern fossil fuel driven economy, refineries engaged in the conversion of crude oil into usable petrochemical products hold immense importance. In spite of their invaluable role in the economy, the huge amount of sludge waste generated during the refining processes causes economic and environmental concerns (Bhattacharyya & Shekdar, 2003; Hu, Li, & Zeng, 2013). The heterogeneous nature of refinery sludges, which are mostly composed of varying proportions of complex aliphatic and aromatic hydrocarbons, nitrogen, sulfur, and/or oxygen containing compounds (NSO), asphaltenes, heavy metals, etc., and their hydrophobic nature lead to their recalcitrance in the environment. The presence of toxic components such as polyaromatic hydrocarbons (PAHs), benzene, toluene, ethylbenzene, xylene (BTEX), as well as metals, such as lead and chromium, have led environmental authorities like United States Environmental Protection Agency (USEPA) and Central pollution control board (CPCB), India to mark these wastes as “hazardous” and issue strict guidelines on their disposal. It is of utmost concern for industries to efficiently manage and remediate such wastes and therefore various physical and chemical strategies, like pyrolysis, solvent extraction, oxidative thermal treatment, centrifugation, air sparging, electrode emulsification, land filling, land farming, solidification/stabilization, soil vapor extraction, incineration, have been employed (Jasmine & Mukherji, 2015; Megharaj, Ramakrishnan, Venkateswarlu, & Sethunathan, 2011; Wang et al., 2010). However, major concerns regarding usage of such strategies entail high economic expense as well as incomplete conversion of pollutants into, sometimes, more harmful intermediates, or just the transfer of pollutants from one phase (e.g., soil) to another (e.g., air) instead of remediation. Since early reports on the ability of prokaryotic microorganisms (bacteria and archaea) to utilize

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00002-X © 2020 Elsevier Inc. All rights reserved.

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hydrocarbons as a source of carbon and energy, these were targeted as agents for remediation (Dean-Ross, Moody, & Cerniglia, 2002; El-Din, Moussa, Moawad, & Sharaf, 2014; Wang et al., 2011). Biodegradation-based remediation has several advantages over conventional physicochemical technologies (viz., being envirofriendly, cost-efficient, and wholesome in terms of mineralization of contaminants) (Cappello et al., 2007; Kumar & Khanna, 2010; Suja et al., 2014). Bioremediation entails three different principles: natural attenuation (NA) (wherein the native community’s ability to degrade the pollutant is solely trusted upon); biostimulation (BS) (wherein nutritionally imbalanced environment is supplemented with organic or inorganic nutrients to support the native community to function faster); and bioaugmentation (BA) (where specialized microbes are added to the specific environment to aid pollutant removal) (Łebkowska et al., 2011; Li et al., 2009; Tyagi, da Fonseca, & de Carvalho, 2011; Vinas, Sabate´, Espuny, & Solanas, 2005; Yu, Wong, Yau, Wong, & Tam, 2005). However, tapping the potential of native microorganisms needs a thorough knowledge about the identity, abundance, and genetic/functional potential of the native community residing at the site of contamination (Dibble & Bartha, 1979; Margesin & Schinner, 1997; Ro¨ling et al., 2004). Researchers have reported various hydrocarbon-degrading microbes, specifically bacteria and archaea from different types of hydrocarbon-associated environments (HAEs) since 1906 (Bastin, 1926; So¨hngen, 1906). Knowledge of petroleum microbiology was enriched with culture-dependent and cultureindependent investigations on petroleum reservoir/oil field/pipeline souring and crude oil degradation; accidental oil spills on sea or land; enhanced microbial oil recovery from oil sand deposits; natural seeps on ocean beds, etc. The most predominant phyla representing hydrocarbon degraders in refinery waste were Proteobacteria followed by Actinobacteria, Acidobacteria, Bacteroidetes, Chloroflexi, Firmicutes, Spirochaetes, Thermotogae, and Caldiserica (Das & Kazy, 2014; de Vasconcellos et al., 2010 and references therein), which work in tandem. Culture-based and enrichment studies have delineated a wide range of microbes capable of the degradation of hydrocarbons of complex structures and have confirmed the selective ability of such microbes in utilizing certain classes of hydrocarbons (Prince, Amande, & McGenity, 2018). Metabolic and genetic studies over the years have deciphered pathways for the activation of chemically inert hydrocarbons under aerobic and anaerobic conditions (Abbasian, Lockington, Mallavarapu, & Naidu, 2015; Rojo, 2009). Recognition of the 16S rRNA gene as a molecular marker and the “big plate count anomaly” (discovery of the unculturable majority) brought breakthroughs in such studies (Staley & Konopka, 1985). The advent of high-throughput next-generation sequencing technologies caused a huge dip in the cost of sequencing and opened new horizons in petroleum microbial ecology research (Levy & Meyers, 2016). Amplicon sequencing of 16S rRNA genes deciphered hitherto unknown diversity from HAEs, the function of which are still being deciphered (Hu et al., 2016). Several studies have also reported the shift in native microbial community composition with respect to bioremediation treatments (Bell et al., 2016; Militon et al., 2010; Roy et al., 2018). Whole metagenome shotgun metagenomics and

Accelerated bioremediation of petroleum refinery sludge 25 transcriptomics deciphered the genetic repertoire from a vast range of HAEs and linked that to the performance of native microbiota under environmental stress (Vigneron et al., 2017). Microbes from HAEs have been found to possess versatile metabolic potentials including aerobic and/or anaerobic hydrocarbon degradation, nitrate-, sulfate-, iron-reduction, syntrophy, fermentation, methanogenesis, methane-oxidation, etc. (An et al., 2013; Daffonchio et al., 2013; Das & Kazy, 2014; Hazen, Rocha, & Techtmann, 2013; Head, Gray, & Larter, 2014; Roy et al., 2018). Metabolic interactions empower microbial communities toward complete mineralization of hydrocarbons, wherein metabolites produced by certain groups are utilized by other groups of microorganisms as a carbon substrate and energy source, depleting the intermediates and driving the thermodynamically unfavorable reactions in the direction toward mineralization. One of the most studied interactions involves the interplay of fermentative and hydrocarbonoclastic microbes producing volatile fatty acids, which in turn are utilized by syntrophic organisms like Smithella and Syntrophus to produce acetate, formate, and hydrogen. Acetoclastic and hydrogenotrophic methanogens convert these substrates to methane and CO2, thus driving the reactions toward degradation under anaerobic conditions in an association termed “syntrophic methanogenic hydrocarbon degradation” (Gieg, Fowler, & Berdugo-Clavijo, 2014; Gray et al., 2011; Tan, Dong, Sensen, & Foght, 2013). On-site bioremediation employing native resident microbial populations could be greatly inhibited by nutrient limitations and/or unfavorable environmental factors like temperature, moisture content, pH, availability of electron donors and/or acceptors, pollutant concentration, etc. within the contaminated sites (Smith et al., 2015). The scarcity of essential nutrients, such as nitrogen, phosphorous, and terminal electron acceptors (TEAs), has been recognized as a critical factor that could inhibit microbial bioremediation performance. Lack of suitable and readily available nutrients in the contaminated sites may lead to the implementation of engineered bioremediation strategies. BS and BA in oilcontaminated environments have been considered as efficient and effective approaches for the restoration of polluted sites (Suja et al., 2014). Previous researchers have reported successful bioremediation of various contaminated sites, wherein BS with the addition of required nutrients resulted in an improved activity of indigenous microorganisms and therefore accelerated the rate of hydrocarbon degradation (Ghaly, Yusran, & Dave, 2013; Smith et al., 2015; Suja et al., 2014). Nitrate amendment has been found to be an effective BS strategy due to its thermodynamic favorability as TEA in the absence of oxygen, which could facilitate the degradation of contaminants (Sarkar et al., 2016). Nevertheless, at high hydrocarbon contamination, the lack of efficient hydrocarbon-degrading microbes could reduce BS efficacy (Almeida et al., 2013). This can be overcome by the addition of specialist microorganisms to the contaminated sites (BA strategy) (Jasmine & Mukherji, 2015). However, this process might also be ineffective owing to the strong competition and predation of autochthonous organisms on the exogenously introduced microbes. Therefore

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the application of native microbial populations in bioremediation strategy development has been preferred as they could easily adapt and acclimatize in the same environment (Suja et al.,2014). Researchers have attempted on-site bioremediation of several kinds of HAEs employing different strategies including NA, BS, and BA as well as combinations of BS and BA approaches (Roy et al., 2018; Sarkar et al., 2016; Smith et al., 2015; Wu et al., 2016). Bioremediation attempts on refinery sludge have yielded varied results but most laboratory-based studies have failed to translate to field scale (Bell et al., 2016; Militon et al., 2010; Stefani, Bell, Marchand, de la Providencia, & El Yassimi, 2015). A lack of understanding of the genetic construct of indigenous microbial communities and effects of stimulants as well as physicochemical conditions of sites and contaminant chemistry hinder the development of efficient sludge bioremediation strategy.

2.2 Petroleum refinery waste: composition and hazard Crude oil recovered from oil reservoirs/fields is processed through refineries to produce diesel, petrol, wax, lubricants, and other petrochemical products. Petroleum refineries import crude oil of various composition (in terms of its hydrocarbon content, viscosity, Air Pollution Index, etc.) and generate usable products through thermal fractional distillation to separate hydrocarbon fractions and subsequent chemical modifications (Hu et al., 2013; Sharma et al., 2016). During the upstream (extraction, transportation, storing of crude oil) and downstream (refining operations) processes, petroleum refineries generate enormous amounts of residual waste sludge (Bhattacharyya & Shekdar, 2003; da Silva, Alves, & de Franc¸a, 2012; Hu et al., 2013; Johnson & Affam, 2018). Refinery generated sludges are categorized into three categories, viz., oily sludge, bio sludge, and chemical sludge (Johnson & Affam, 2018; Kriipsalu, Marques, & Maastik, 2008). Sludge generated through all of these operations, viz., tank bottom and effluent treatment plant oily sludge, are sent for further treatment (for residual oil recovery) or disposal. Oily sludge is characterized as a viscous mixture of hydrocarbons, soil, and water and its composition depends on the density and viscosity of the crude oil and the refining capacity (Katsivela, Moore, & Kalogerakis, 2005; Overcash & Pal, 1979). The Ministry of Petroleum and Natural Gas, Govt. of India estimates 249.4 million metric tons of crude oil was refined in 23 refineries across India in the year 2018 (Ministry of Petroleum and Natural Gas, 2018). Indian Oil Cooperation Limited (IOCL) is one of the major stakeholders and registered crude throughput of 69 MMT and a total of 55,656 MT of sludge generation in its different establishments (Sustainability report, IOCL, 2018). Oily sludge has been categorized as a hazardous waste and reported to be composed of heavier hydrocarbons, water, and other sediments (Cerqueira et al., 2011; Hu et al., 2013). The hydrocarbon content of oily sludge varies and is a complex, heterogenous mix of petroleum hydrocarbons containing diverse compounds such as aliphatics, aromatics, NSO, and asphaltenes (Gallego et al., 2007; Head, Jones, & Larter, 2003; Hu et al., 2013; Mishra, Jyot, Kuhad, & Lal, 2001;

Accelerated bioremediation of petroleum refinery sludge 27 Reddy et al., 2012; Roy et al., 2014). The total petroleum hydrocarbons (TPHs) content itself can also be highly variant, ranging from 5% to 86% of the sludge mass (very commonly between 15% and 50%) while the water content could be as high as 30% 85%. Other hazardous constituents of the waste include heavy metals that can occur at concentration ranges of 7 80 mg/kg for zinc (Zn), 0.001 0.12 mg/kg for lead (Pb), 32 120 mg/kg for copper (Cu), 17 25 mg/kg for nickel (Ni), and 27 80 mg/kg for chromium (Cr) (Da Rocha, Dantas, Duarte, Duarte, & Da Silva, 2010; Rolda´n-Carrillo, Castorena-Corte´s, Zapata-Pen˜asco, Reyes-Avila, & Olguı´n-Lora, 2012). Another key component of oily sludge that makes it of grave environmental concern is PAHs. Previous studies have reported several PAHs to be toxic, mutagenic, and carcinogenic to various degrees (Guazzaroni et al., 2013). Other carcinogenic pollutants in the sludge include the BTEX fraction (benzene, toluene ethylbenzene, and xylene) (Almeida et al., 2013; Cerqueira et al., 2011; Guazzaroni et al., 2013; Hu et al., 2013; Liu et al., 2009; Megharaj et al., 2011; Pepi et al., 2009). The physicochemical properties along with other sludge characteristics, such as solubility, lipophilicity, and polarity, are of vital importance in order to understand the fate of these contaminants in various environmental matrices (Jones et al., 2008; Masciandaro, Macci, Peruzzi, Ceccanti, & Doni, 2013; Tyagi et al., 2011; Van Hamme, Singh, & Ward, 2003). Additionally, the ability to design effective cleanup and remediation strategies hinges heavily on a detailed understanding of the sludge properties. Improper disposal of sludge due to lack of regulations into neighboring ecosystems has led to severe loss in fertility, nutrient imbalance, and stunted growth of flora and fauna (Saeki, Sasaki, Komatsu, Miura, & Matsuda, 2009; Varjani & Upasani, 2016). Harmful chemicals such as PAH and BTEX infiltrate air or ground/surface water environments causing damage to animal and human health in its vicinity (Baheri & Meysami, 2002; Dı´ez, Jover, Bayona, & Albaige´s, 2007; Mason, Hazen, & Borglin, 2012). Overall, the complex nature of the sludge and the prevalence of several toxic bioaccumulants make oil contamination a severe hazard to the ecosystem (Almeida et al., 2013; Marco-Urrea, Garcia-Romera, & Aranda, 2015; Safdari et al., 2018; Sharma et al., 2016). The quantity of sludge produced and its components (toxic, mutagenic PAHs, heavy metals, as well as aliphatics and aromatics) render them a very serious environmental threat. CPCB, USEPA, and Organization for Economic Cooperation and Development (OECD) organizations have designated oily sludge wastes as hazardous owing to its composition. All of this necessitates the development of efficient and suitable disposal and cleanup strategies, the failure of which might have severe detrimental effects to the contaminated ecosystem. In the subsequent sections of this chapter a brief account on understanding the biological and nonbiological strategies that are typically employed for degradation of hydrocarbon and petrochemical waste is presented.

28

Chapter 2

2.3 Microbiology of hydrocarbon-associated environments In order to understand and devise proper bioremediation strategy it is important to know “who is there?” in terms of microbial community composition and “what are they able to do?,” that is, their metabolic potential. Evaluation of whether the native microorganisms are sufficient for efficient degradation provided the presence of suitable physicochemical conditions and nutrients, will determine the kind of bioremediation strategy to be used. In the next section we highlight some key findings from previous studies detailing hydrocarbon metabolism by microbes, followed by the microbial community composition reported from various HAEs. Ever-evolving microbes have figured out strategies not only to survive under hydrocarbon stress, but some also depend solely on hydrocarbons as their carbon and energy source. These microbes have been reported to be ubiquitous; however, the genes related to degradation, primarily carried in plasmids, are enriched in the presence of hydrocarbons. Zobell et al. (1946) pioneered research on petroleum microbiology by establishing the capacity of microbes to “eat” oil. The presence of microbes in extreme environments, such as oil fields, was first reported by researchers looking for reasons for the “souring of oil fields.” It was reported that microbes from an oil reservoir in Illinois basin were capable of sulfate reduction by Bastin (1926). It was the result of the investigation of the causative agent for “souring” (conversion of sulfate to sulfide and hydrogen sulfide, causing reductions in the value of the oil produced). These findings led to the development of the entire field of bioremediation, with a plethora of literature trying to exploit the capabilities of these microbes, which is discussed thoroughly in the next section. Biodegradation of organic compounds has been an area of active research for a long time, however, knowledge about the limits of abilities possessed by the microbes to withstand any extreme conditions are still being tested. Research on the metabolic diversity of microbes over the past 100 years has established the role of microbes, specifically bacteria and fungi, in hydrocarbon degradation. Several studies have previously attempted to characterize the microbial populations that inhabit hydrocarbon-contaminated sites. Effective bioremediation strategy development requires a thorough understanding of the composition and dynamics of the indigenous microbial community and their ability to function under physiological stress. Numerous investigators have elucidated the microbial community structure of different petroleum hydrocarbonimpacted environments, such as underground petroleum hydrocarbon plumes adjacent to the former refinery, soil contaminated with crude oil, oily sludge, oil naturally mixed with formation water, employing clone library analysis (Allen et al., 2007; Das & Kazy, 2014; Silva et al., 2013; Zhang, Mo¨rtelmaier, & Margesin, 2012). Amplified ribosomal DNA restriction analysis (ARDRA) is one of the primitive methods designed for community profiling. ARDRA was used as a powerful tool that could provide preliminary phylogenetic information and the grouping of potential degrading microorganisms, thereby reducing the

Accelerated bioremediation of petroleum refinery sludge 29 time exhausted in selecting microorganisms for further biodegradation studies (Oliveira, da Rocha Calixto, Felippe, & de Franca, 2013). Bacterial diversity of bioremediated oilcontaminated beach samples was analyzed by Ro¨ling et al. (2002) using 16S rRNA gene clone library and ARDRA. Characterization of microbial diversity of production wells was carried out using denaturing gradient gel electrophoresis (DGGE) and 16S rRNA gene libraries by Dellagnezze et al. (2016). DGGE was used to study the dynamics of the petroleum refinery sludge community during bioremediation (Sarkar et al., 2016). The microbial community diversity of oil or water samples from the production wells of Shengli oil fields injected with nutrients were studied by culture-independent molecular techniques, which included a 16S rRNA gene clone library, terminal restriction fragment length polymorphism, and DGGE by Xingbiao, Yanfen, Sanqing, Zhiyong, and Yanhe (2015). Stable isotope probing (SIP) involves a stable isotope, for example, 13C-labeled substrate, to microbial communities and whose utilization is of interest in deciphering a key biogeochemical process (Dumont & Murrell, 2005; Wellington, Madgavkar, & Ryan, 2003). Bell et al. (2011) demonstrated the role of Alphaproteobacteria, especially Sphingomonas and other members of Sphingomonadaceae family, in nitrogen incorporation in hydrocarbon-contaminated Arctic soil through 15N-DNA SIP combined with large-scale sequencing for its possible bioremediation by biostimulating the hydrocarbon-degrading community. Fluorescence in situ hybridization (FISH) has also been extensively used for quantification of the presence and relative abundance of microbial populations within a community. Catalyzed reporter deposition-FISH, an improved version of FISH, has been used to elucidate the microbial community of hydrocarbon-contaminated aquifers, the production water of oil fields, etc. (Lenchi et al., 2013; Tischer et al., 2012). The estimation of the microbial population by calculating the copy number using quantitative polymerase chain reaction in oil-contaminated samples was previously reported from PAHcontaminated soil, production water, and injection water (Gao et al., 2015; Peng, 2015). Numerous next-generation sequencing platforms, for example, 454 pyrosequencing (Roche), HiSeq, MiSeq, Solexa (Illumina), and Ion Torrent, have been widely applied by many investigators (Shokralla, Spall, Gibson, & Hajibabaei, 2012). The application of these novel, high-throughput, and ultradeep sequencing methods has immense potential in analyzing microbial community composition and function in contaminated environments. The pairing of metagenomics with an emergent pyrosequencing technology marked the beginning of a sequencing revolution. In rapid succession 454 (Roche), Illumina, SOLiD (Life Technologies), and Pacific Biosciences drove the cost of sequencing down to ,US$0.1 per Mb, which democratized metagenomic analysis, with a concomitant explosion of studies across a wide range of environments to investigate the diversity and functional capacity of all domains of life. The Roche 454 GS-FLX Titanium is capable of generating more than 1 million reads per run of 1000 bp (average 500 bp) length within 23 h; the average run generates 750 Mbp of sequencing data. Pyrosequencing was used for unraveling the

30

Chapter 2

microbial community composition of oil-contaminated sediment cores, beach sands exposed to Deepwater Horizon oil spills, oil-contaminated soil, oil field samples, tailing ponds, etc. (Bell et al., 2011; Fowler, Toth, & Gieg, 2016; Kostka et al., 2011; Lamendella et al., 2014; Liao, Wang, & Huang, 2015; Peng, 2015). Ha˚velsrud, Haverkamp, Kristensen, Jakobsen, and Rike (2012) and Sutton et al. (2013) used pyrosequencing for the microbial community of oilcontaminated pockmarked sediments and refueling station samples, respectively. Illuminabased platforms cost around US$0.50 per Mb, but have a longer run time than 454 pyrosequencing. Currently this feature is being addressed by the MiSeq Illumina machine, which has been developed in order to run smaller jobs at a much faster rate with relatively high throughput. Illumina allows sample preparation sizes of ,20 ng DNA (similar to 454 pyrosequencing). Wu et al. (2016) analyzed the community shift during microcosm-based bioremediation of oil-contaminated soil using the Illumina MiSeq platform. Recently Gao et al. (2018) used the Illumina HiSeq2500 platform to demonstrate the community composition of a contaminated shoreline. Yergeau et al. (2012) used Ion Torrent to describe the community composition of tailing ponds. Bell, Yergeau, Juck, Whyte, and Greer (2013), Bell, Yergeau, Maynard, et al. (2013), and Joshi et al. (2014) employed the Ion Torrent sequencing platform to determine the microbial community of diesel-contaminated Arctic soil and petroleum muck, respectively. Microbial community shift during BS -based bioremediation of refinery sludge was also elucidated by Sarkar et al. (2016) the 318 Ion Express chip in a Personal Genome Machine of Ion Torrent (Life Technologies). A similar platform was used by Stagars, Mishra, Treude, Amann, and Knittel (2017) to explore microbial community response to simulated petroleum seepage in Caspian Sea sediments. Recently Guerra et al. (2018) used the Ion Torrent Personal Genome Machine for analyzing the metagenome of drill cuttings. Tools such as Newbler (Roche) and MIRA 4 are commonly used in metagenomics for performing reference-based assemblies, whereas Velvet and Abyss, which were among the first to perform de novo assembly, are still being widely used at present (Sharpton, 2014). The difficulty of assembling and annotating the data, due to short read lengths, has been the primary challenge to analyze high-throughput metagenomic/metatranscriptomic data (Wooley, Godzik, & Friedberg, 2010). With the recent advancement in bioinformatics sequencing, assembly errors detection and correction have significantly improved. Many packages are now available which provide pipelines to bring these new algorithms into the lab (Cole et al., 2008; Schloss et al., 2009). Annotation of metagenomic sequences starts with a series of preprocessing steps that prepare the read for annotation. The dereplication step removes the sequences that are more than 95% identical and the final step of preprocessing is performed by the MG-RAST pipeline (Oulas et al., 2015). MG-RAST uses similarity to compare three known databases— Greengenes (DeSantis et al., 2006), Ribosomal Database Project-RDP (Cole et al., 2008), and SILVA (Quast et al., 2013)—to predict rRNA genes. The 16S rRNA gene data derived from these sequencing technologies were further analyzed by using software tools like Mothur (Schloss et al., 2009), QIIME (Quantitative Insights

Accelerated bioremediation of petroleum refinery sludge 31 Into Microbial Ecology) (Caporaso et al., 2010), and SILVAngs (Quast et al., 2013). The most recent of the three is SILVAngs, which provides a fully automated analysis pipeline for data derived from 16S rRNA gene amplicon sequencing. The analysis workflow is based on (1) alignment of reads; (2) quality assessment and filtering of reads; (3) dereplication, whereby identical sequences are filtered out to avoid overestimation; (4) clustering and operational taxonomic unit (OTU) picking using a priori defined thresholds; and (5) taxonomic assignment of OTUs using the SILVA rDNA database (Quast et al., 2013). The next stage of the annotation pipeline involves functional assignment to the predicted protein coding genes. Some widely used data repositories to obtain annotation for metagenomic datasets include functional annotation databases such as KEGG (Ogata et al., 1999), COG/KOG (Tatusov, Galperin, Natale, & Koonin, 2000), SEED (Overbeek et al., 2005), and eggNOG (Powell et al., 2014). Phylogenetic investigation of communities by reconstruction of unobserved states (PICRUSt) connects the taxonomic classification from metaprofiling results with metabolic information (Langille et al., 2013). PICRUSt predicts the functional potential of the community from ribosomal 16S rRNA gene data, deriving information from whole genome inventory of closely related organisms. However, it only works adequately for those environments where the results have large numbers of organisms with annotated reference genomes available. Finally, PICRUSt was designed to analyze prokaryotes, ignoring a large amount of metabolic features performed by eukaryotes. Characterization of native microbial communities, diversity, metabolic potential, and response (shift in community composition) toward biostimulatory agents are considered as prerequisites for bioremediation technology development. Microbial diversity across various such hydrocarbon-affected environments is shown in Fig. 2.1A D. The abundance of Proteobacteria in petroleum hydrocarbon-rich environments has been observed by numerous investigators (Das & Kazy, 2014; Gao et al., 2015, 2018; Joshi et al., 2014; Kostka et al., 2011; Lamendella et al., 2014; Lenchi et al., 2013; Liao et al., 2015; Mason et al., 2012; Sarkar et al., 2016; Tan et al., 2013; VanMensel et al., 2017; Wu et al., 2016; Zhang et al., 2012). Lamendella et al. (2014) reported the maximum abundance of Alphaproteobacteria in samples with maximum hydrocarbon content. The abundance of Alphaproteobacteria in production and injection wells has been previously reported by Lenchi et al. (2013). Within Alphaproteobacteria, species belonging to Brevundimonas, Sphingopyxis, Ochrobactrum, Hyphomonas, Novosphingobium, Paracoccus, and Sphingomonas have been reported to have degradation potential (Abed, Al-Kindi, & Al-Kharusi, 2015; Gao et al., 2015). A number of purple nonsulfur Alphaproteobacterial members like Rubribacterium, Rhodomicrobium, Porphyrobacter, and Rhodospirillum were reported from oil-contaminated cyanobacterial mats (Abed, Al-Sabahi, Al-Maqrashi, Al-Habsi, & Al-Hinai, 2014). Various Alphaproteobacterial members like Novosphingobium and Paracoccus were also capable of nitrate reduction. The presence of sequences related to nitrogen-fixing have been found in

32

Chapter 2

Figure 2.1: Distribution of microbial phyla in different hydrocarbon-contaminated environments. (A) Deepwater Horizon spill, tailing pond, and oil field samples; (B) injection, production, and formation water samples; (C) oil-contaminated soil samples; and (D) refinery waste.

Accelerated bioremediation of petroleum refinery sludge 33

Figure 2.1: (Continued)

34

Chapter 2

Alphaproteobacteria, for example, Azospirillum, Rhizobium, Bradyrhizobium, and Mesorhizobium have been found to inhabit hydrocarbon-rich sites (Abed et al., 2014; Gao et al., 2015; Guazzaroni et al., 2013). Betaproteobacterial dominance was noted in refueling station samples, production water, injection water, and crude oil-contaminated soil (Abed et al., 2015; Bell et al., 2011; Lenchi et al., 2013; Morais, Pylro, Clark, Hirsch, & To´tola, 2016; Sutton et al., 2013). Among Betaproteobacteria, various hydrocarbon-degrading genera, for example, Thiobacillus (nitrate reducer and thiosulfate oxidizer), Limnobacter (thiosulfate oxidizer), Methyloversatilis (methylotrophs), Acidovorax (nitrate reducer), and Polaromonas (chemolithotrophy), were abundant in hydrocarbon-rich environments (Guazzaroni et al., 2013; Lenchi et al., 2013). Rhodocyclaceae (Betaproteobacteria) members have also been known to utilize nitrate, perchlorate, Fe(III), and other metals as electron acceptors during the biodegradation of aromatic hydrocarbons under anaerobic conditions (Das & Kazy, 2014). Within Proteobacterial classes, Gammaproteobacteria was one of the most abundant groups in hydrocarbon-contaminated sites (Gao et al., 2015; Guazzaroni et al., 2013; Lamendella et al., 2014; Morais et al., 2016; Sun et al., 2018). Most of the well-studied hydrocarbon-degrading genera, including Pseudomonas, Pseudoxanthomonas, Marinobacter, Marinobacterium, Acinetobacter, Alkanivorax, Stenotrophomonas, and Shewanella, were members of Gammaproteobacteria (Abed et al., 2015; Atlas, 1981; Gao et al., 2015; Lenchi et al., 2013; Varjani, 2017). Sulfate-reducing Deltaproteobacterial members like Desulforhabdus, Desulfomonile, Desulfosarcina, Desulfotignum, Desulfacinum, Desulfotalea, Desulfotomaculum, Desulfurivibrio, Desulfobacterium, Desulfatibacillum, Desulfomicrobium, Desulfovibrio, Desulfococcus, Desulfosalsimonas, and Desulfovermiculus have been reported from petroleum industry waste and contaminated sites (Abed et al., 2015; Gao et al., 2015; Joshi et al., 2014; Mukherjee et al., 2017). Syntrophic organisms (Syntrophaceae, Geobacteraceae, etc.) catalyze the degradation of hydrocarbon substrates to produce H2 and/or acetate, which could subsequently be utilized by methanogens (Methanobacteriaceae, Methanoregulaceae, and Methanosaetaceae) (Gieg et al., 2014; Tan et al., 2015). Such interactions are quite common in hydrocarbon-impacted ecosystems (An et al., 2013; Das & Kazy, 2014; Head et al., 2014; Ismail, Ijah, Riskuwa, & Allamin, 2014; Joshi et al., 2014; Mukherjee et al., 2017; Sarkar et al., 2016). Members of Epsilonproteobacteria including Sulfurospirillum and Arcobacter have been known to be associated with the cycling, oxidization, and reduction of sulfur and nitrogen, but were least represented when compared to other members of Proteobacteria (Abed et al., 2015; Gao et al., 2015). The predominance of Proteobacteria in hydrocarbon-rich environments could be attributed to the fact that the members of the phylum were known to be metabolically versatile and phylogenetically most diverse within the bacterial domain (Silva et al., 2012). Firmicutes has been reported as one of the most frequently reported phyla in hydrocarbon-impacted environments (Gao et al., 2015, 2018; Head et al., 2014; Lenchi et al., 2013; Sarkar et al., 2016; Tan et al., 2015; Wu et al., 2016). Koo, Mojib, Huang, Donahoe, and Bej (2015) observed Firmicutes as a late

Accelerated bioremediation of petroleum refinery sludge 35 responder during oil contamination, that is, their population increased after perturbation with crude oil. Among Firmicutes, different genera like Bacillus, Clostridium, Paenibacillus, Planomicrobium, Streptococcus, and Exiguobacterium were reported from production water, refueling station samples, oily sludge, etc. and are considered to be highly versatile in the context of hydrocarbon degradation (Das & Kazy, 2014; Lenchi et al., 2013; Sarkar et al., 2016; Sun et al., 2018; Sutton et al., 2013; Tan et al., 2015). Thiosulfatereducers from Clostridiaceae and Peptostreptococcaceae families have been previously isolated from production waters in Northeastern India (Agrawal & Lal, 2009). Higher abundance of Actinobacteria was also documented in different oil-contaminated sites, for example, Dietzia, Rhodococcus, Arthrobacter, Brachybacterium, Tessaracoccus, Georgenia, Corynebacterium, Kocuria, Microbacterium, Propionobacterium, and Mycobacterium (Abed et al., 2014; Gao et al., 2015; Kostka et al., 2011; Lenchi et al., 2013; Liao et al., 2015; Sutton et al., 2013; Wu et al., 2016). Members of Acidobacteria have been considered to be abundant in oil-associated environments (Abed et al., 2015; Abercron et al., 2016; Joshi et al., 2014; Mukherjee et al., 2017). Acidobacteria members are also known to be highly adaptable, metabolically flexible, and widely distributed in hydrocarbon-contaminated environments (Das & Kazy, 2014; Militon et al., 2010). Understanding the microbial community composition, structure, and dynamics within a contaminated niche has been recommended as critical for designing in situ bioremediation strategies. Ecologically hydrocarbon-metabolizing microorganisms were known to be widely distributed. Under anaerobic sulfate-reducing conditions, members of Chloroflexi were known to be involved in the anaerobic fermentation of alkanes and other hydrocarbon metabolism and even had connections with methanogenesis through reverse electron transport (Xia, Wang, Wang, Chin, & Zhang, 2016). Abundance of Chloroflexi, specifically Anaerolineae, was reported from refinery waste, oil-contaminated soil, and water (Sarkar et al., 2016; Silva et al., 2013; Sutton et al., 2013). Thermotogae, Bacteroidetes, Spirochaetes, Synergistetes, and Deferribacteres were reported in varying abundance from similar environments (Abed et al., 2015; Sutton et al., 2013). Deferribacteres are acetoclastic iron-reducing bacteria that have been reported from production water of mesothermic petroleum reservoirs (Silva et al., 2013). Siles and Margesin (2018) reported Gammaproteobacteria and Bacteroidia classes to be actively involved in TPH removal in the NA and BS of hydrocarbon-contaminated Alpine soil. Hu et al. (2016) have demonstrated moderate abundance of various candidate division phyla like OP1, OP9, OP11, OD1, etc. in oil reservoirs (Table 2.1). Among the archaeal phyla predominately Euryarchaeota and few genera of Crenarchaeota have been reported from tailing ponds, refueling station soil, the Deepwater Horizon oil spill, petroleum muck, etc. (Joshi et al., 2014; Mason et al., 2014; Sutton et al., 2013). Both hydrogenotrophic and acetoclastic methanogens like Methanobacterium, Methanosaeta, Methanomicrobiales, and Methanocella(members of Euryarchaeota) have been reported

36

Chapter 2 Table 2.1: Details of different hydrocarbon-contaminated environments investigated for microbial community distribution.

Sample name

Description

Reference

5F, 5S, 5V, and 5C

Allen et al. (2007)

OS-71b JP-8 MBR_10 SU3 TPM1-1

Underground petroleum hydrocarbon plume adjacent to the former refinery Beach sands exposed to Deepwater Horizon spill JP-8 fuel spill Arctic soil Membrane reactor treating oil refinery waste Tailings pond Oil-contaminated pockmarked sediments

SCHO SMDW

Soil highly contaminated with oil Salt marsh from Deepwater Horizon spill

PPDW MBR_13 SCAD PNF1, PNF2, PFOH1, PFOH2, and PFS1 A2_RS, B3_RS, and B4_RS PTS1 and GMR75 BPCL PM EBOil

Proximal plume from Deepwater Horizon spill Membrane reactor treating oil refinery waste. Enrichment cultures from oil sands mining trailing Production water

Kostka et al. (2011) Bell et al. (2011) Silva et al. (2012) Yergeau et al. (2012) Ha˚velsrud et al. (2012) Zhang et al. (2012) Beasley and Nanny (2012) Mason et al. (2012) Silva et al. (2013) Tan et al. (2013) Lenchi et al. (2013)

Refueling station

Sutton et al. (2013)

BP074 JBT60, JBT70, SYT, and JBT1 HB DQ XJ SL Lu3084 and Lu3096 Lu3065 NA CT BR A3, B3, and C3 DB2 GR1

Oil naturally mixed with formation water Oily sludge Petroleum muck Oil-contaminated sediments cores

Silva et al. (2013) Das and Kazy (2014) Joshi et al. (2014) Lamendella et al. (2014) DW impacted sediment Mason et al. (2014) The oil-contaminated soil from abandoned well mined; the Peng (2015) oil-contaminated soil from recent (1 year) oil-spilled site Huabei oil field sample (China) Liao et al. (2015) Daqing oil field sample (China) Karamay oil field sample (China) Shengli oil field sample (China) Injection well, production well, production well Gao et al. (2015)

Soil contaminated with crude oil Control set without any treatment of soil contaminated with crude oil Bioremediated soil contaminated with crude oil Oil-contaminated beach sediment Digboi Refinery waste Guwahati Refinery waste

Wu et al. (2016)

Gao et al. (2018) SRR2646856 SRR2645287

from oily sludge and production water (Das & Kazy, 2014; Gao et al., 2015; Sarkar et al., 2016). Dolfing (2014) suggested the abundance of hydrogenotrophic methanogens over acetoclastic methanogens during crude oil biodegradation. Crenarchaeota members have been known to be chemolithotrophic—capable of utilizing ammonia or other reduced

Accelerated bioremediation of petroleum refinery sludge 37 inorganic compounds—or chemoorganotrophic—capable of utilizing simple or complex organic compounds (Lliro´s, Casamayor, & Borrego, 2008). The distribution of microbial community members in different petroleum hydrocarbon-contaminated environments has been well studied (Fig. 2.1 and Table 2.2). Table 2.2: Bioremediation studies on petroleum-contaminated samples with high TPH content. Initial TPH concentration (g/kg)

Sl no. Remediation approach

Percentage TPH reduction (%)

1

BA

2

BA and BS (composting)

160.30 372.50 Refinery oily sludge, India 370 China

Mandal et al. (2012) Ouyang et al. (2005)

3

BS [C:N:P ratio 5 100/ 1.74/0.5] using inorganic nutrients

95% within 2 12 months 46% 53% degradation within 56 days 32% 51% after 30 days

334

Rolda´n-Carrillo et al. (2012)

4

Combined biostimulation and bioaugmentation (fungal nutrient amendment) BS (biopile)

Up to 91% degradation

110

Oil sludge from a natural gas processing facility in Tabasco, Mexico Oil tank bottom sludge

60% degradation in 3 months

300

80.6% within 1 year BA (Bacillus subtilis, Bacillus megaterium, Achromobacter xylosoxidans, Pseudomonas fluorescens, Candida tropicalis, and Rhodotorula dairenensis) Slurry-based BA and BS 24% biostimulation

220

Contaminated soil from refinery in SE Spain Bottom sludge of oil separating tank in Zhongyuan oil field, Puyang, China

Marı´n, Moreno, Herna´ndez, and Garcı´a (2006) He, Duan, and Liu (2014)

Weathered oily waste (PB401) from a 10year-old disposal site in Poza Bellota, Tabasco, Me´xico Artificial contamination with Barauni refinery sludge, Bihar, India

Machin-Ramirez et al. (2008)

Oil sludge contaminated soil from Shengli oil field, Shandong Province, Northern China

Liu, Luo, Teng, Li, and Ma (2010)

5

6

7

8

9

90.2% within 120 BA [Acinetobacter days baumannii (S19, S26, S30), Burkholderia cepacia (P20), Pseudomonas sp. (S24)], and nutrients BS (manure) 58.2% in 360 days

130

99.2

250

Contaminant site

Reference

Adetutu et al. (2015)

Mishra et al. (2001)

(Continued)

38

Chapter 2 Table 2.2: (Continued) Initial TPH concentration (g/kg)

Sl no. Remediation approach

Percentage TPH reduction (%)

10

BS (inorganic nutrients)

11

BS and BA

12

BS and BA (microcosms, artificial contamination)

15 80 70% 90% during 2 months, regardless of their initial concentrations C14 23, 46% 67% 22 55 for PD and 28% 42% for KH in presence of inorganic nutrients with or without bacteria within 90 days 25 80% decrease with combined BSBA in 28 weeks

13

Biopile with the aid of biostimulation BS (column fixed bed reactor)

14

85% in 76 days

13

10.4 C12 20, 80% C21 35, 38%, C35 40, 44%. after 2 years 61.5% with 0.5% 36.6 surfactant in 90 days

15

BS (Tween 80)

16

Combined BS and bioremediation

5% in 112 days

61

17

BS (compost, NPK fertilizer)

88.8

18

BA and BS

19

BS (acetate and methanol)

20

BS (compost)

.90% with compost, .70% with NPK fertilizer in 2 months 58% with BA and 48% with BS in 8 weeks 64% with 10 mM acetate, 53% with 20 mM methanol after 30 weeks 34% in 12 weeks

20.2

Contaminant site

Reference

Oily sludge from Haifa oil refinery, Israel

Admon, Green, and Avnimelech (2001)

Sediment samples from oilcontaminated land at Mina Al-Fahal, Muscat, and Oman

Abed et al. (2014)

Crude oilcontaminated soil samples from Jorhat, Assam, India Diesel-contaminated site in Algeria Industrial site contaminated by progressive leakage

Roy et al. (2014)

Contaminated soil from Tayoltita, Mexico

de la Cueva, Rodrı´guez, Cruz, Contreras, and Miranda (2016) Wu et al. (2017)

Crude oilcontaminated site from oil well in China Artificially contaminated soil

Petroleumcontaminated soil from oil field, China 5.6 [C10 C40] Marine oilcontaminated sediment from Hong Kong 0.9 Alkane-contaminated soil

Chemlal et al. (2013) Militon et al. (2010)

Nwankwegu, Orji, and Onwosi (2016) Wu et al. (2017) Zhang and Lo (2015)

Wallisch et al. (2014)

BA, Bioaugmentation; BS, biostimulation; TPH, total petroleum hydrocarbon.

Culture-independent molecular methods have revealed complex assemblages of diverse bacteria and archaea, which are capable of hydrocarbon degradation, nitrate/sulfate/ironreduction, fermentation, and methane metabolism in various HAEs (An et al., 2013; Daffonchio et al., 2013; Das & Kazy, 2014; Hazen et al., 2013; Head et al., 2014). Considerable work on petroleum microbiology has shown the abundance of

Accelerated bioremediation of petroleum refinery sludge 39

Figure 2.2 Schematic representation of metabolic processes in hydrocarbon-associated environments (HAEs). The boxes marked with star represent aerobic pathways and other boxes represent anaerobic pathways.

hydrocarbon-degrading organisms, particularly bacteria and archaea capable of utilizing a wide range of hydrocarbons as nutritional resources (Alonso-Gutie´rrez et al., 2009; An et al., 2013; Fowler et al, 2016; Hazen et al., 2013; Head et al., 2014). A comprehensive study on microbial diversity of refinery sludge from different refinery sludge collected from refineries across Assam, India, using high-throughput next-generation sequencing revealed the dominance of strictly anaerobic, fermentative, thermophilic, sulfate-reducing bacteria (SRB) affiliated to Coprothermobacter, Fervidobacterium, Treponema, Syntrophus, Thermodesulfovibrio, Anaerolinea, Syntrophobacter, Anaerostipes, Anaerobaculum, etc.; along with methanogenic Methanobacterium, Methanosaeta, Thermoplasmatales, etc. archaeal groups (Roy et al., 2018) (Fig. 2.2). Isolation and cultivation of microorganisms from oily sludge sample has revealed that a great variety of microorganisms reside in such environments. It also provides information about the community composition, and biodegradation profiles can shine a light on microbial metabolic preferences, yielding correlations between microbial products and typical biomarkers of degraded oils, as well as further potential applications in bioremediation studies or the recovery of reservoirs (Afifi, Motamedi, Alizadeh, & Leilavi, 2015; Cerqueira et al., 2011; Das & Kazy, 2014; Obi et al., 2016; Pal et al., 2017; Sarkar et al., 2017). Several studies of

40

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the microbiology of petroleum reservoirs suggest that only strictly anaerobic microorganisms could be considered as truly indigenous to these ecosystems. Nevertheless, aerobic, facultative anaerobic, and microaerophilic microorganisms have been found in such environments. Hydrocarbon-degrading genera have been reported from several petroleum impacted environments, some of which are discussed below (Atlas, 1981; Fathepure, 2014; Leahy & Colwell, 1990; Varjani, 2017; Van Hamme et al., 2003; ZoBell, 1946). Alcanivorax sp., Cycloclasticus sp., Oleiphilus sp., Oleispira sp., Thalassolituus sp., and some members of the genus Planomicrobium were reported from hydrocarbonimpacted marine samples (Head et al., 2003). Twenty-three bacterial strains were isolated from hydrocarbon-contaminated sites of Guwahati city, out of which a Bacillus cereus strain was identified which was capable of degrading anthracene, naphthalene, pyrene, and benzo(a)pyrene (Das, Bhattacharya, Banu, & Kotoky, 2017). Even though a wide range of metabolically diverse bacteria have been successfully isolated from petroleum oil sludge samples, there is a significant difference between the microbial diversity present in the environment and the fraction that can be recovered through culturing. Biosurfactantproducing Bacillus subtilis A1 was isolated from a crude oil reservoir (Parthipan et al., 2017). Gordonia sp. nov. Q8 have previously been reported from oil field produced water PAHs or crude oil-contaminated water and soil (Qi, Wang, Lv, Lun, & Zheng, 2017). Communities of microorganisms rather than single strains are most important in bioremediation as their metabolic diversity and sometimes metabolic redundancy may contribute to process robustness. Gallego et al. (2007) isolated 30 strains from oil refinery tank bottom sludge. Among them consortia including Pseudomonas, Acinetobacter, Bacillus, and Rhodotorula were screened on the basis of cometabolic effects, colonization on oil components, emulsification properties, and degradative capabilities, which were further used for alkanes and aromatic degradation. Roy et al. (2014) also isolated 39 bacterial strains belonging to the genera Lysinibacillus, Brevibacillus, Bacillus, Paenibacillus, Stenotrophomonas, Alcaligenes, Delftia, Achromobacter, and Pseudomonas from crude oil-contaminated soil. Among these isolates a consortium consisting of three strains of Pseudomonas aeruginosa and a strain of Achromobacter xyloxidans was used for bioremediation of crude oil-contaminated soil samples in a microcosm-based study. Obi et al. (2016) isolated Pseudomonas, Bacillus, Stenotrophmonas, Achromobacter, Bordetella, Ochrobactrum, Advenella, Brucella, Ochrobactrum, Advenella, Mycobacterium, Mesorhizobium, Klebsiella, Pusillimonas, and Raoultella from a composting bin containing crude oil. Among these isolates Pseudomonas stood out as the best hydrocarbon degrader. It is well-known that the majority of microbes in environmental samples cannot be cultured at present in laboratory media, which are biased for the growth of specific microorganisms. The design of efficient bioremediation systems requires a set of careful studies of the site-specific microbial communities and of their catabolic potentials. Recently Kumari, Regar, and Manickam (2018) reported the degradation of numerous polycyclic aromatic hydrocarbons (PAHs) present in crude oil by

Accelerated bioremediation of petroleum refinery sludge 41 Microbacterium esteraromaticum, P. aeruginosa, Pseudomonas mendocina, Ochrobactrum anthropi, and Stenotrophomonas maltophilia, which were further used as consortia. The isolation and whole-genome sequencing of a potent hydrocarbonoclastic strain, Franconibacter pulveris DJ34, from Duliajan oil fields, Assam, showed the utilization of and genetic machinery for degradation of diverse petroleum hydrocarbons and electron acceptors, metal resistance, and biosurfactant production (Pal et al., 2017). The combined use of molecular biology techniques with traditional cultivation-based methods has been employed recently in petroleum microbiological studies, providing a more precise perspective of the microbial community diversity in such extreme environments. Sarkar et al. (2017) established an enrichment-based isolation technique to get culturable access to the microbial population of oily sludge. On the other hand, describing microbial species from a petroleum environment or investigating their ability to use hydrocarbons as carbon sources, the application of traditional culturing techniques to recover these microorganisms from the environment is essential. The culture-based methods, such as classical plating and most probable number, focused on isolating and identifying microorganisms from hydrocarbon-impacted environments and the descriptions of the microbial communities have been solely based on functional characteristics (Van Hamme et al., 2003). CLPP, which can be performed by the BIOLOG system, has been used widely to analyze microbial communities based on the ability of microbes to utilize different carbon sources (Lehman, Colwell, Ringelberg, & White, 1995; Shi, Bischoff, Turco, & Konopka, 2005; Wang, Wang, Wang, Li, & Guo, 2012). An assemblage of diverse bacterial groups capable of the degradation of major constituents of the waste as well as being involved in biochemical cycling of nitrogen, iron, carbon dioxide methane, sulfur, etc. have been reported from several petroleum-rich sites. Genespecific or shotgun metagenome sequencing or microarray-based analyses, have demonstrated the cooccurrence of catabolic genes related to microbial aerobic and anaerobic hydrocarbon biodegradation and associated nutrient metabolism. Such associations of microorganisms might be contributing in the concerted metabolic interactions for biodegradation of hydrocarbons in complex environments. Microorganisms residing in the contaminated habitat often have the requisite degradation potential. The presence of enzymes capable of degrading hydrocarbons within the microbial community allows them to thrive in the inhospitable conditions (Fuentes, Me´ndez, Aguila, & Seeger, 2014). The key enzymes involved in hydrocarbon degradation pathways are oxygenases, which catalyze the addition of molecular oxygen to the substrate. Dioxygenases catalyze the addition of two hydroxyl groups, whereas monoxygenases catalyze the introduction of one atom of oxygen into the hydrocarbon (Rojo, 2009). During anaerobic degradation, activation of hydrocarbons could be achieved by coupling CO2 or fumarate to hydrocarbons, with nitrate or sulfate as TEAs (Callaghan, Tierney, Phelps, & Young, 2009). The genes responsible for the biodegradation of hydrocarbons are usually clustered,

42

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comprising catabolic genes encoding catabolic enzymes, transport genes encoding proteins for active uptake of the compounds, and regulatory genes responsible for the regulation of the expression of both catabolic and transport genes (Lin & Cai, 2008; Peixoto, Vermelho, & Rosado, 2011; Peng et al., 2008). Operon alkBFGHJKL encodes for the enzymes involved in converting alkanes into acetyl-coenzyme A (Acetyl-CoA), while alkST encodes for rubredoxin reductase and the positive regulator for the alkBFGHJKL operon (Van Hamme et al., 2003). alkB and CYP153 encode alkane hydroxylases and cytochrome P450 that are responsible for aerobic alkane degradation in oil-polluted environments. The functional diversity of the alkB gene in oil-contaminated sites showed their affiliation with Gordonia, Mycobacterium, Rhodococcus, Rhodobacter, Burkholderia, Acinetobacter, Legionella, and Marinobacter. Similarly, CYP153 genes were found in the sequenced genomes of Bradyrhizobium, Rhodopseudomonas, Caulobacter, and Novosphingobium (Nie, Liang, Fang, Tang, & Wu, 2014). Wasmund, Burns, Kurtbo¨ke, and Bourne (2009) also studied the diversity of the alkB gene in hydrocarbon seeps which indicated close resemblance with Gammaproteobacteria. Among the aerobic PAH-degrading genes, the operon nahAaAbAcAdBFCED encodes the pathway for conversion of naphthalene to salicylate and nahGTHINLOMKJ encodes for the conversion of salicylate to acetaldehyde and pyruvate via catechol meta-cleavage. These genes have been reported from Pseudomonas, Burkholderia, and Rhodococcus. Genes involved in phenanthrene degradation, pbhABCD, have been reported from Sphingomonas paucimobilis, Nocardiodes sp., and Mycobacterium sp. (Van Hamme et al., 2003). Anaerobic hydrocarbon genes include bssA, which codes for benzylsuccinate synthase—a key enzyme that initiates anaerobic toluene degradation by attaching fumarate to the methyl group of toluene to form benzyl succinate (Kazy, Monier, & Alvarez, 2010). The enzyme benzylsuccinate synthase has been reported from Azoarcus sp. and Thauera aromatic (Van Hamme et al., 2003). The ecosystem of a hydrocarbon-impacted environment also comprises SRBs, nitrate-reducing bacteria, and methanogens (Gao et al., 2015). Sarkar et al. (2016) showed the presence of nitrogen fixers in refinery waste. Dissimilatory reduction of nitrate to N2 or intermediate products is catalyzed by several enzymes working in tandem, some of which are nitrite reductase (nirS) and nitrate reductase (narG), while N2 fixation is normally catalyzed by a nitrogenase enzyme whose iron subunit is encoded by the nifH gene (Sarkar et al., 2016). Nitrate reductase (narG) and ammonia oxidation genes have been reported from bioremediation of hydrocarbon-contaminated groundwater (Yagi et al., 2010). Diversity analysis of dissimilatory sulfite reductase genes (dsrA and dsrB, involved in dissimilatory sulfite reduction) indicated the presence of a sulfate-reducing microbial population within the community. Obi Chioma et al. (2016) elucidated a sulfate-reducing community of production water samples of Indian oil fields using DGGE of the dsrB gene to reveal the abundance of Desulfovibrio, Desulfotomaculum, Desulfomicrobium, Desulfosporosinus, and Desulfitobacterium. The sulfate-reducing microbial community of the Gulf of Mexico was investigated by high-throughput sequencing of dsrAB genes to reveal their affiliation with

Accelerated bioremediation of petroleum refinery sludge 43 Desulfobacteraceae and Syntrophobacteraceae members (Vigneron et al., 2017). The methanogenic community in the oily sludge was characterized by the analysis of a functional marker, the mcrA gene encoding the alpha subunit of the enzyme involved in the final methane-forming step during methanogenesis (Das & Kazy, 2014; Gao et al., 2015; Lutton, Wayne, Sharp, & Riley, 2002). These genes indirectly demonstrate the functional potential of the anaerobic microbial population of the hydrocarbon-impacted environments (Gao et al., 2015; Sarkar et al., 2016). A clone library of the mcrA gene from anaerobic digestors showed close affiliation with Methanobacteriaceae, Methanothermaceae, and Methanosaetaceae family members (Wallisch et al., 2014). Methanolobus and Methanosarcina reported mcrA genes from organic carbon-dominated deep Fennoscandian Bedrock fluids (Purkamo et al., 2015). The tools mentioned previously provided us with knowledge about either the microbial community composition or its functional potential, however, oftentimes knowledge about both helps in comprehending the dynamics of a microbiota. Omics approaches can overcome the limitations imposed by classical approaches, enabling a broader perspective of the taxonomy and functional variety of environmental microorganisms and access to their metabolic potential (Handelsman, Rondon, Brady, Clardy, & Goodman, 1998).

2.4 Microbial bioremediation of waste sludge The chemical complexity of oily sludge makes its treatment and disposal extremely challenging. Physicochemical remediation methods have been conventionally and extensively used, including techniques such as pyrolysis, solvent extraction, oxidative thermal treatment, centrifugation, air sparging, electrode emulsification, land filling, land farming, solidification/stabilization, soil vapor extraction, incineration, and transfer of hydrocarbon from one phase to other (air) (Jasmine & Mukherji, 2015; Wang et al., 2010). These approaches are cost intensive contributing to further economic constraints (Rayu, Karpouzas, & Singh, 2012; Varjani, 2017). Another roadblock in the use of these technologies entailed incomplete conversion of contaminants and the production of more persistent secondary contaminants. Compared to these physicochemical treatment technologies, microorganism-based bioremediation technology has been considered to be an effective and economically sound strategy for the treatment of hydrocarbon-contaminated wastes (Cappello et al., 2007; Dellagnezze, Gomes, & de Oliveira, 2018; Kumar & Khanna, 2010; Miri, Naghdi, Rouissi, Kaur Brar, & Martel, 2019; Varjani, 2017). Development of an appropriate remediation strategy depends on the physicochemical properties of the polluted matrix and on the degree and age of the contaminated sites. The oily sludge biodegradation can be affected by a variety of factors, such as site environmental conditions, the indigenous microbial population, availability of nutrients, contaminant chemistry, and availability. Microbes harbor genetic machinery necessary for either

44

Chapter 2

complete or partial conversion of organic as well as inorganic compounds to harmless end products under aerobic as well as oxygen-deficient conditions (An et al., 2013). These “microbial cell factories” assert their suitability as agents for bioremediation due to their nutritional versatility, smaller and adaptable genomes, diverse catabolic potentials, as well as their ability to survive and perform under harsh environmental conditions. Technologies utilizing bioremediation are mainly of two types: (1) in situ, wherein direct treatment of contaminants occur at the site of contamination (bioventing, on-site BS); and (2) ex situ, where the contaminated matrix is extracted from its site and treated, for example, bioreactors, composting, land farming, anaerobic digestion, biosorption. Both have their advantages and disadvantages based on the nature and extent of contamination. In situ bioremediation has emerged as a potential alternative cleanup technology that utilizes the inherent potential of native microbial communities to metabolize, and thus degrade, hydrocarbons as an energy as well as a carbon source, resulting in removal of the contaminants from the wastes or waste-impacted environment (Das & Chandran, 2011; Ivshina et al., 2015; Pal, Banat, Almansoori, & Haijra, 2016). Based on the principle, bioremediation technologies can be classified as (1) NA, (2) BS, or (3) BA. Natural degradation of hydrocarbon compounds has given researchers as well as industry the idea of NA. In this process, the indigenous microbial community is trusted fully to degrade the contaminants. Although it entails the least financial involvement, the time required for this process to occur restricts its use. However, this strategy is used sometimes as external intervention of any sorts can topple the balance in some sensitive environments. The BS approach uses the knowledge about the discrepancy in nutrient concentrations in the contaminated sample and tries to normalize it by the addition of nutrients to stimulate the activity of indigenous microorganisms, while BA utilizes microorganisms, either indigenous or known potent degraders, to enhance the rate of biodegradation of pollutants. Hydrocarbon composition, physicochemical (viz., temperature, pH, moisture content, nutrient availability) characteristics, and native microbial community composition influence the selection of the biodegradation strategy for remediation (Gillespie & Philp, 2013; Kleinsteuber, Schleinitz, & Vogt, 2012; Lu et al., 2014; Smith et al., 2015; Zhang et al., 2012).

2.4.1 Accelerated bioremediation The rapid expansion of petroleum and allied industries has added to the economic prosperity of India (and many other countries as well), but has led to the generation of large volumes of hazardous oily sludge that warrant immediate attention (Hu et al., 2013; Sarkar et al., 2017). Sustainable and affordable disposal/remediation of all petroleum hydrocarbon-rich waste sludge/contaminants is a prime technological impediment. Microbial bioremediation has emerged as the most feasible, yet effective solution to attain complete pollutant degradation (Cerqueira et al., 2011). Oily sludge waste management includes a three-tiered strategy. First, reduction of the quantity of oily sludge produced from petroleum industries by

Accelerated bioremediation of petroleum refinery sludge 45 employing suitable technologies. Second, recovery of valuable fuel from the existing sludge. Third, safe disposal of the unrecoverable residues or oily sludge, where neither of the first two tiers could be applied (Hu et al., 2013). Present physicochemical technologies for remediation of contaminated sites and refinery waste include solidification/stabilization, soil vapor extraction, incineration, solvent extraction, and chemical treatments. However, these conventional physicochemical approaches were generally found to be expensive due to the cost of transportation of large volumes of contaminated materials (waste) for ex situ treatment, viz., soil washing, chemical inactivation (use of potassium permanganate and/or hydrogen peroxide as a chemical oxidant to mineralize nonaqueous contaminants such as petroleum), and incineration (Rayu et al., 2012; Varjani, 2017). The remediation process is often impaired due to the incomplete conversion of the parent compounds, producing metabolites which are more persistent and equally or more toxic to nontarget organisms, that is, they generate secondary contaminants. The most common in situ remediation used for treating contaminated soils was biopiles, composting, and land farming (Hu et al., 2013). Armed with the knowledge of the participants in biodegradation and their metabolic potential toward hydrocarbon degradation, the studies conducted using bioremediation approaches, as well as their successes and pitfalls are elaborated in the next sections under two broad categories of BS and BA. Some studies on BS, BA, and BSBA (combination of both BS and BA) have been tabulated and presented in Table 2.2. The successes and failures of these studies have added to the knowledge base for development of suitable strategies for bioremediation of hydrocarbon-contaminated environments. Different strategies are involved in bioremediation to remediate petroleum hydrocarbon- contaminated waste (Fig. 2.3).

Figure 2.3 Schematic diagram indicating various bioremediation strategies.

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Chapter 2

2.4.1.1 Biostimulation BS is a bioremediation strategy that relies on the indigenous ability of the microbiota, however, then aim is to achieve a nutrient balance for their proper growth and functioning. Supplementation with one or more rate-limiting nutrients (e.g., nitrogen (N), phosphorous (P), sulfur (S)) in hydrocarbon-loaded environments has been aimed at accelerating the growth of microorganisms, most of which are capable of the degradation of the hydrocarbons. Biostimulants can be water-soluble inorganic nutrients, oleophilic fertilizers, slow-release fertilizers, as well as oxygenation. Aeration/oxygenation is cost intensive; however, it helps in the delivery of oxygen to aerobic degraders which are better and more efficient hydrocarbon degraders. This has been achieved through tilling, forced aeration of soil, as well as the addition of bulking agents which increase the porosity of soil and thus help the diffusion of oxygen, and has been successfully applied in several cases. Rhykerd, Crews, McInnes, and Weaver (1999) reported 82% TPH degradation in 12 weeks in the treatment of oil-contaminated soil. Among all the electron acceptors substituted for oxygen 22 2 31 (NO2 3 ; Fe ; SO4 ; CO2 ), NO3 is the best alternative since it solubilizes easily, is less expensive (Barbaro et al., 1992), and its physicochemical properties, such as energy yield and redox potential, are the same as those of oxygen. Many reports cite the beneficial effects of N addition on biodegradation rates (Walworth & Reynolds, 1995). Some studies have also shown the ill effects of inundating N levels (Chaineau, Rougeux, Yepremian, & Oudot, 2005; Oudot, Merlin, & Pinvidic, 1998) which might expedite osmotic potential depression leading to the reduction of the rate of microbial activity. Microbes require a threshold level of the essential nutrients carbon (C), nitrogen (N), and phosphorus (P) in order to grow and several reports have vouched for the optimal ratio being 100:10:1 (C:N:P), although variations also exist (Dias et al., 2012; Dibble & Bartha, 1979; Zhao, Selvam, & Wong, 2011). Excess carbon in the form of hydrocarbons causes a paucity of essential nutrients, which has also been found as the critical factor in reducing microbial metabolism in several laboratory studies (Atlas, 1981; Ron & Rosenberg, 2014; Varjani, Thaker, & Upasani, 2014). BS has been reported to be superior to BA in various cases where the indigenous community was more resilient and effective in hydrocarbon degradation (Kauppi, Sinkkonen, & Romantschuk, 2011). Organic wastes, such as biochar, sawdust, crop residues, leaf litter, compost, have also been successfully utilized to remediate hydrocarbon-contaminated samples (Simons et al., 2013; Qin, Gong, & Fan, 2013). Optimization of the C:N:P ratio for the growth and performance of bacteria has been thoroughly studied in the literature, however, the results varied considerable based on the type of contaminant. While 51% TPH degradation for an oily sludge from Mexico was reported within 30 days with a C:N:P ratio of 100:1.74:0.5 (Rolda´n-Carrillo et al., 2012), reductions of TPH content from 1042 to 470 mg/kg and 133 to 93 g/kg in 15 days in a slurry reactor were reported for ratios of 100:1.76:1.73 and 100:0.6:0.5, respectively. ´ lvarez, Balbo, Mac Cormack, & Ruberto, 2015; Machin-Ramirez et al., 2008). (A

Accelerated bioremediation of petroleum refinery sludge 47 Amending nitrate levels has been recommended as one of the most efficient BS approaches, which could facilitate efficient oxidation of hydrocarbons, thus resulting in bacterial growth 1 and hydrocarbon catabolism (Bell et al., 2016; Dashti et al., 2015). Among NO2 3 and NO4 , the former was found to perform better, due to acid production during ammonium metabolism (Chen & Strous, 2013; Ramstad & Sveum, 1995; Wrenn, Haines, Venosa, Kadkhodayan, & Suidan, 1994; Xia et al., 2016). Børresen and Rike (2007) reported elevated hexadecane mineralization on addition of NH4Cl and NaCl at various concentrations compared to an unamended microcosm, although an elongated lag period was observed in the microbial community. Two different BS strategies, the addition of bulking agent and nutrients, on oil sludge showed positive effects of bulking agent addition on microbial diversity concomitant to TPH reduction. Negative effects of BS using urea were reported, probably due to the acidification of sludge and a decrease in microbial diversity (Wang et al., 2012). Metagenomic investigation of hydrocarbon degradation and the shift in microbial community composition in a bioremediation treatment site in Canadian high Arctic soil indicated the role of γ-Proteobacteria members (Pseudomonas) during degradation, who dwindled in abundance after bioremediation treatment for a year, the community was then dominated by members of α-Proteobacteria and Actinobacteria (Yergeau et al., 2012). Intrinsic biodegradation ability of the indigenous microorganisms of a refinery sludge from Guwahati refinery, Assam (with very high TPH content 400 g/kg) was enhanced significantly ( . 80% reduction in TPH by 90 days) with nitrate amendment. Preferred utilization of both higher- ( . C30) and middle-chain (C20 30) length hydrocarbons was evident from gas chromatography-mass spectrometry data, concomitant to hydrocarbon degradation, lowering of dissolved O2, and an increase in oxidation reduction potential. Marked increase in abundance of N2 fixing, nitrate-reducing aerobic/facultative anaerobic members (e.g., Azovibrio, Pseudoxanthomonas, and Commamonadaceae members) was also evident in N-amended microcosm (Sarkar et al., 2016). Successful bioremediation events have been reported for various hydrocarboncontaminated sites, wherein the amendment of appropriate inorganic nutrients (N and/or P) resulted in enhanced growth and activity of efficient indigenous microorganisms, thus expediting bioremediation (Adetutu et al., 2015; Agarry, Aremu, & Aworanti, 2013; Bento, Camargo, Okeke, & Frankenberger, 2005; Gillespie & Philp, 2013; Kleinsteuber et al., 2012; Rahman, Rahman, & Rahman, 2003; Roy et al., 2014; Smith et al., 2015; Stroud, Paton, & Semple, 2007; Zhang et al., 2012; Zhang & Lo, 2015). 2.4.1.2 Bioaugmentation BA uses extraneously added microbes, singly or as a consortium, to enhance and aid the metabolic potential of contaminated sites (Alvarez and Illman, 2005; El Fantroussi & Agathos, 2005; Iwamoto & Nasu, 2001). Bioaugmenting agents are characterized by their ability to not only degrade the contaminant of interest but also be resilient to the prevailing environmental conditions, as well as their genetic stability and survivability in competition

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Chapter 2

with native microbiota (Gentry, Rensing, & Pepper, 2004). Initial studies used single cultures with versatile and potent hydrocarbon degradation ability, for example, Acinetobacter sp. capable of degrading C10 C40 n-alkanes; biosurfactant-producing strains which enhance the bioavailability of pollutants; the utilization of Gram-negative Pseudomonas, Flavobacterium, Sphingomonas, and Achromobacter strains, as well as Gram-positive bacteria species such as Mycobacterium, Bacillus, and Rhodococcus, for BA (El Fantroussi & Agathos, 2005; Gentry et al., 2004; Throne-Holst, Wentzel, Ellingsen, Kotlar, & Zotchev, 2007). Several in situ BA -based studies have been conducted and have showed efficient TPH degradation (Bhattacharya, Sarma, Krishnan, Mishra, & Lal, 2003; Gao et al., 2014; Katsivela et al., 2005; Mansur et al., 2014; Mishra et al., 2001; Sarma, Bhattacharya, Krishnan, & Lal, 2004). However, the inability of a single microbe to degrade a wide range of hydrocarbons and the complex nature of wastes led researchers to develop mixtures of microbes (consortia) for enhanced degradation (Heinaru et al., 2005). Bento et al. (2005) reported the degradation of 73% 75% of the light (C12 C23) and heavy (C23 C40) fractions of TPH after bioaugmenting with a bacterial consortium, as opposed to 46% by NA. A consortium composed of Mycobacterium fortuitum, B. cereus, Microbacterium sp., Gordonia polyisoprenivorans, Microbacteriaceae bacterium, and Fusarium oxysporum, was found to degrade 96% 99% of PAHs (anthracene, phenanthrene, and pyrene) in 70 days (Jacques et al., 2008). Sprocati et al. (2012) demonstrated the use of microbial consortia of Delftia tsuruhatensis, B. cereus, Pseudomonas resinovorans, Pseudomonas fluorescerens, Exiguobacterium sp., Pseudomonas jessenii, Arthrobacter sp., and Rhodococcus erythropolis in the bioremediation of diesel and heavy metal cocontaminated soil. BA with a consortium (composed of Aeromonas hydrophila, Alcaligenes xylosoxidans, Gordonia sp., Pseudomonas fluorescens, Pseudomonas putida, Rhodococcus equi, S. maltophilia, and Xanthomonas sp.) showed 89% TPH degradation of diesel oil contaminated soil in a year (Szulc et al., 2014). Another consortium composed of Ochrobactrum sp., S. maltophilia, and P. aeruginosa reported 84% crude oil degradation (Varjani, Rana, Jain, Bateja, & Upasani, 2015). An indigenous community amended with Bacillus subitilis showed enhancement in TPH degradation (Tao et al., 2017). A TPHdegrading consortium consisting of Pseudomonas stutzeri, Pseudomonas SZ-2, Bacillus SQe2, and Acinetobacter SZ-1 obtained from a petroleum-polluted soil in combination with NH4NO3 and K2HPO4 remediated TPH from soil from 61 to 57.9 g/kg in 112 days (Wu et al., 2016). A consortium of M. esteraromaticum, O. anthropi, S. maltophilia, O. anthropi, P. mendocina, and P. aeruginosa along with biosurfactant enhanced PAH degradation by up to 10% (Kumari et al., 2018). Recently, Siles and Margesin (2018) reported that the addition of N, P, and K resulted in an increased TPH removal potential of the native microbiome. An evaluation of BS (using N.P and N 1 P) and BA (biosurfactant-producing Bacillus strains) as a remediation technique for refinery sludge, revealed 46% 55% TPH removal through BS with an abundant presence of fermentative, hydrocarbondegrading, sulfate-reducing, CO2-assimilating, and methanogenic microorganisms

Accelerated bioremediation of petroleum refinery sludge 49 (Bacillus, Coprothermobacter, Rhodobacter, Pseudomonas, Achromobacter, Desulfitobacter, Desulfosporosinus, T78, Methanobacterium, Methanosaeta, etc.). Combined BS and BA enhanced TPH reduction by 57% 75% (Roy et al., 2018). These studies highlight the efficiency of BA, however, lab-based BA studies have often failed to translate to field applications, owing to their survivability as well as loss of function. Considering the theoretical limitation of bioremediation posed by the requirement of “combined capacity of catabolic pathways,” BA with catabolically superior bacteria has been identified as being more advantageous in achieving improved remediation (Dueholm et al., 2015). As a result, the biodegradation of oil hydrocarbons typically needs a microbial consortium with a succession of species. It was reported that the employment of a mixed bacterial culture was more advantageous in comparison with pure cultures due to synergistic interactions among microbial species (Gallego et al., 2007; Ghaly et al., 2013; Hu et al., 2013; Tyagi et al., 2011). Different microorganisms have showed potential to utilize various components of petroleum hydrocarbons, but there is no single bacterial strain available that could completely mineralize all the components of the oily sludge (Van Hamme et al., 2003; Megharaj et al., 2011; Roy et al., 2014; Sprocati et al., 2012; Varjani, 2017). The biostimulation and BA -based technologies have also been widely used for the treatment of xenobiotic compounds, long- and medium-chain alkanes, mono- and polycyclic aromatics hydrocarbons-contaminated soils and waters, which could be a potential solution for hydrocarbon-associated pollution (Akinde, Iwuozor, & Obire, 2012; Roy et al., 2014; Sarkar et al., 2016; Sprocati et al., 2012; Sun et al., 2012; Wu et al., 2016).

2.5 Factors affecting bioremediation Bioremediation relies on the degradative power of microbes and thus the major factors affecting growth and survivability of microbes directly affect the rate of biodegradation. Major factors that dictate the biodegradation of hydrocarbons include (1) environmental factors (presence of TEAs, nutrient availability, salinity, pressure, temperature, pH, water availability, and osmotic stress) (Smith et al., 2015); (2) the nature of the contaminant (solubility, concentration, hydrophobicity, volatility, molecular mass); and (3) the native microbial community composition and its functional ability (genetic complement, gene regulation and expression, surface hydrophobicity, metabolic diversity and flexibility, substrate uptake or adherence mechanisms, tolerance to metals and other toxic xenobiotics, chemotaxis, biofilm formation) (Atlas, 1981; Banat et al., 2010; Bordoloi & Konwar, 2009; Botalova, Schwarzbauer, Frauenrath, & Dsikowitzky, 2009; Cerqueira et al., 2011; Couling, Towell, & Semple, 2010; Fuentes et al., 2014; Hu et al., 2013; Kleinsteuber et al., 2008; Leahy & Colwell, 1990; Varjani, 2017). Biodegradation by indigenous microbes using the available nutrients and electron acceptors is the most important process contributing to NA in contaminated aquifers. To monitor NA and assess in situ biodegradation of hydrocarbons, a detailed understanding of the microbial activities is required (Kleinsteuber et al., 2012).

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The success of bioremediation using native microbiomes is greatly constrained by imbalanced nutrient levels and/or physicochemical conditions prevailing at the contaminated sites (Lu et al., 2014; Smith et al., 2015). Particularly, the lack of essential nutrients (e.g., nitrogen (N), phosphorous (P), TEA) has been found to be the critical factor in reducing the rate of microbial metabolism and thus natural bioremediation. The lack of appropriate and readily available nutrients necessitates engineered bioremediation strategies including BS and/or BA that might accelerate the bioremediation process. Microbes require a threshold level of the essential nutrients carbon (C), nitrogen (N), and phosphorus (P) in order to grow and several reports have vouched for the optimal ratio of 100:10:1 (C:N:P), although variations also exist (Dias et al., 2012; Dibble & Bartha, 1979; Zhao et al., 2011). An excess of C in the form of hydrocarbons causes a paucity of essential nutrients [e.g., nitrogen (N), phosphorous (P), sulfur (S)] and has often been found to be the critical factor in reducing the rate of microbial metabolism in several labbased studies (Atlas, 1981; Evans et al., 2015; Ron & Rosenberg, 2014; Varjani et al., 2014). Bacterial growth is sensitive to changes in pH, with optimal growth reported in most cases under neutral conditions (Dibble & Bartha, 1979; Leahy & Colwell, 1990). Biodegradation, which involves these microbes, has thus been reported to be optimal in the pH range of 6.5 8.5. Temperature is one of the controlling factors in hydrocarbon degradation as it impacts not only the growth rate of the microbes but also the gas solubility and physical and chemical properties of hydrocarbon components. (Leahy & Colwell, 1990; Megharaj et al., 2011). While low temperature increases the viscosity of oil, decreasing the volatilization of lighter fractions (Leahy & Colwell, 1990), higher temperature increases the solubility of hydrocarbons and their rate of diffusion as well as decreasing viscosity (Aislabie, Balks, Foght, & Waterhouse, 2004). The hydrocarbon contaminants pose a challenge since they are highly heterogenous. The degradability of hydrocarbons has been reported in the literature as follows: alkanes . isoalkanes . low molecular weight aromatics . cyclic alkanes (naphthenes) . polyaromatics . polar compounds . asphaltenes. The storage and long-term aging of sludge have been reported to complexify the problem by reducing bioavailability as well as causing the simultaneous loss of volatile and gain in recalcitrant fractions in the sludge. Initial concentrations of hydrocarbons present (usually measured as TPH) have been reported to inhibit the functioning of potent degraders. This clearly highlights that the structure and initial concentration of hydrocarbon compounds, as well as the fraction of each component present in the sludge, determine biodegradability. Aliphatic, aromatic, and asphaltene components of crude oil comprise the major components of refinery sludge, however, composition and relative abundance of these fractions vary considerably between samples. While aliphatics are easier to degrade, aromatics, branched aliphatics, and polyaromatic compounds pose harder challenges. Several enzyme systems have been discovered which work under aerobic as well as anaerobic conditions to degrade these complex organics to simple inorganic products in energetically favorable reactions.

Accelerated bioremediation of petroleum refinery sludge 51

2.6 Future scope Refinery sludge microbiomes comprise diverse hydrocarbon-degrading microbes with vast metabolic potential (fermentation, sulfate reduction, syntrophism, nitrogen fixing, and methanogenesis). Technological advances in the area of metagenomics have enhanced the knowledge base of microbial composition and their genetic machinery in sludge samples. Proper utilization of this knowledge, as well as metaproteomic and metatranscriptomic insights, will help in the development of suitable bioremediation strategies for refinery waste. The heterogeneity of waste composition and prevailing environmental factors could be further modified to enhance degradation. Moreover, value addition and improved costeffectiveness of strategies could be achieved through bioproducts (enzymes, biosurfactants), biogas generation, and energy recovery systems.

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Szulc, A., Ambro˙zewicz, D., Sydow, M., Ławniczak, Ł., Piotrowska-Cyplik, A., Marecik, R., & Chrzanowski, Ł. (2014). The influence of bioaugmentation and biosurfactant addition on bioremediation efficiency of diesel-oil contaminated soil: Feasibility during field studies. Journal of Environmental Management, 132, 121 128. Tan, B., Dong, X., Sensen, C. W., & Foght, J. (2013). Metagenomic analysis of an anaerobic alkane-degrading microbial culture: Potential hydrocarbon-activating pathways and inferred roles of community members. Genome, 56(10), 599 611. Tan, B., Fowler, S. J., Laban, N. A., Dong, X., Sensen, C. W., Foght, J., & Gieg, L. M. (2015). Comparative analysis of metagenomes from three methanogenic hydrocarbon-degrading enrichment cultures with 41 environmental samples. ISME Journal, 9(9), 2028 2045. Tao, K., Liu, X., Chen, X., Hu, X., Cao, L., & Yuan, X. (2017). Biodegradation of crude oil by a defined coculture of indigenous bacterial consortium and exogenous Bacillus subtilis. Bioresource Technology, 224, 327 332. Tatusov, R. L., Galperin, M. Y., Natale, D. A., & Koonin, E. V. (2000). The COG database: A tool for genomescale analysis of protein functions and evolution. Nucleic Acids Research, 28(1), 33 36. Throne-Holst, M., Wentzel, A., Ellingsen, T. E., Kotlar, H. K., & Zotchev, S. B. (2007). Identification of novel genes involved in long-chain n-alkane degradation by Acinetobacter sp. strain DSM 17874. Applied and Environmental Microbiology, 73(10), 3327 3332. Tischer, K., Zeder, M., Klug, R., Pernthaler, J., Schattenhofer, M., Harms, H., & Wendeberg, A. (2012). Fluorescence in situ hybridization (CARD-FISH) of microorganisms in hydrocarbon contaminated aquifer sediment samples. Systematic and Applied Microbiology, 35(8), 526 532. Tyagi, M., da Fonseca, M. M. R., & de Carvalho, C. C. (2011). Bioaugmentation and biostimulation strategies to improve the effectiveness of bioremediation processes. Biodegradation, 22(2), 231 241. Van Hamme, J. D., Singh, A., & Ward, O. P. (2003). Recent advances in petroleum microbiology. Microbiology and Molecular Biology Reviews, 67(4), 503 549. VanMensel, D., Chaganti, S. R., Boudens, R., Reid, T., Ciborowski, J., & Weisener, C. (2017). Investigating the microbial degradation potential in oil sands fluid fine tailings using gamma irradiation: A metagenomic perspective. Microbial Ecology, 74(2), 362 372. Varjani, S. J. (2017). Microbial degradation of petroleum hydrocarbons. Bioresource Technology, 223, 277 286. Varjani, S. J., Rana, D. P., Jain, A. K., Bateja, S., & Upasani, V. N. (2015). Synergistic ex-situ biodegradation of crude oil by halotolerant bacterial consortium of indigenous strains isolated from on shore sites of Gujarat, India. International Biodeterioration & Biodegradation, 103, 116 124. Varjani, S. J., Thaker, M. B., & Upasani, V. N. (2014). Optimization of growth conditions of native hydrocarbon utilizing bacterial consortium “HUBC” obtained from petroleum pollutant contaminated sites. Indian Journal of Applied Research, 4(10), 474 476. Varjani, S. J., & Upasani, V. N. (2016). Biodegradation of petroleum hydrocarbons by oleophilic strain of Pseudomonas aeruginosa NCIM 5514. Bioresource Technology, 222, 195 201. Vigneron, A., Alsop, E. B., Cruaud, P., Philibert, G., King, B., Baksmaty, L., . . . Tsesmetzis, N. (2017). Comparative metagenomics of hydrocarbon and methane seeps of the Gulf of Mexico. Scientific Reports, 7(1), 16015. Vinas, M., Sabate´, J., Espuny, M. J., & Solanas, A. M. (2005). Bacterial community dynamics and polycyclic aromatic hydrocarbon degradation during bioremediation of heavily creosote-contaminated soil. Applied and Environmental Microbiology, 71(11), 7008 7018. Wallisch, S., Gril, T., Dong, X., Welzl, G., Bruns, C., Heath, E., . . . Schloter, M. (2014). Effects of different compost amendments on the abundance and composition of alkB harboring bacterial communities in a soil under industrial use contaminated with hydrocarbons. Frontiers in Microbiology, 5, 96. Walworth, J. L., & Reynolds, C. M. (1995). Bioremediation of a petroleum-contaminated cryic soil: Effects of phosphorus, nitrogen, and temperature. Soil and Sediment Contamination, 4(3), 299 310.

Accelerated bioremediation of petroleum refinery sludge 63 Wang, L., Li, Y., Yu, P., Xie, Z., Luo, Y., & Lin, Y. (2010). Biodegradation of phenol at high concentration by a novel fungal strain Paecilomyces variotii JH6. Journal of Hazardous Materials, 183(1 3), 366 371. Wang, X., Wang, Q., Wang, S., Li, F., & Guo, G. (2012). Effect of biostimulation on community level physiological profiles of microorganisms in field-scale biopiles composed of aged oil sludge. Bioresource Technology, 111, 308 315. Wang, X. B., Chi, C. Q., Nie, Y., Tang, Y. Q., Tan, Y., Wu, G., & Wu, X. L. (2011). Degradation of petroleum hydrocarbons (C6 C40) and crude oil by a novel Dietzia strain. Bioresource Technology, 102(17), 7755 7761. Wasmund, K., Burns, K. A., Kurtbo¨ke, D. I., & Bourne, D. G. (2009). Novel alkane hydroxylase gene (alkB) diversity in sediments associated with hydrocarbon seeps in the Timor Sea, Australia. Applied and Environmental Microbiology, 75(23), 7391 7398. Wellington, S., Madgavkar, A., & Ryan, R. (2003). U.S. Patent Application No. 10/279,226. Wooley, J. C., Godzik, A., & Friedberg, I. (2010). A primer on metagenomics. PLoS Computational Biology, 6(2), e1000667. Wrenn, B. A., Haines, J. R., Venosa, A. D., Kadkhodayan, M., & Suidan, M. T. (1994). Effects of nitrogen source on crude oil biodegradation. Journal of Industrial Microbiology, 13(5), 279 286. Wu, M., Dick, W. A., Li, W., Wang, X., Yang, Q., Wang, T., . . . Chen, L. (2016). Bioaugmentation and biostimulation of hydrocarbon degradation and the microbial community in a petroleum-contaminated soil. International Biodeterioration & Biodegradation, 107, 158 164. Wu, M., Li, W., Dick, W. A., Ye, X., Chen, K., Kost, D., & Chen, L. (2017). Bioremediation of hydrocarbon degradation in a petroleum-contaminated soil and microbial population and activity determination. Chemosphere, 169, 124 130. Wu, M., Ye, X., Chen, K., Li, W., Yuan, J., & Jiang, X. (2017). Bacterial community shift and hydrocarbon transformation during bioremediation of short-term petroleum-contaminated soil. Environmental Pollution, 223, 657 664. Xia, Y., Wang, Y., Wang, Y., Chin, F. Y. L., & Zhang, T. (2016). Cellular adhesiveness and cellulolytic capacity in Anaerolineae revealed by omics-based genome interpretation. Biotechnology for Biofuels, 9 (111). Available from https://doi.org/10.1186/s13068-016-0524-z. Xingbiao, W., Yanfen, X., Sanqing, Y., Zhiyong, H., & Yanhe, M. (2015). Influences of microbial community structures and diversity changes by nutrients injection in Shengli oilfield, China. Journal of Petroleum Science and Engineering, 133, 421 430. Yagi, J. M., Suflita, J. M., Gieg, L. M., DeRito, C. M., Jeon, C. O., & Madsen, E. L. (2010). Subsurface cycling of nitrogen and anaerobic aromatic hydrocarbon biodegradation revealed by nucleic acid and metabolic biomarkers. Applied and Environmental Microbiology, 76(10), 3124 3134. Yergeau, E., Lawrence, J. R., Sanschagrin, S., Waiser, M. J., Korber, D. R., & Greer, C. W. (2012). Nextgeneration sequencing of microbial communities in the Athabasca River and its tributaries in relation to oil sands mining activities. Applied and Environmental Microbiology, 78(21), 7626 7637. Yu, K. S. H., Wong, A. H. Y., Yau, K. W. Y., Wong, Y. S., & Tam, N. F. Y. (2005). Natural attenuation, biostimulation and bioaugmentation on biodegradation of polycyclic aromatic hydrocarbons (PAHs) in mangrove sediments. Marine Pollution Bulletin, 51(8 12), 1071 1077. Zhang, D. C., Mo¨rtelmaier, C., & Margesin, R. (2012). Characterization of the bacterial archaeal diversity in hydrocarbon-contaminated soil. Science of the Total Environment, 421, 184 196. Zhang, Z., & Lo, I. M. (2015). Biostimulation of petroleum-hydrocarbon-contaminated marine sediment with co-substrate: Involved metabolic process and microbial community. Applied Microbiology and Biotechnology, 99(13), 5683 5696. Zhao, Z., Selvam, A., & Wong, J. W. C. (2011). Effects of rhamnolipids on cell surface hydrophobicity of PAH degrading bacteria and the biodegradation of phenanthrene. Bioresource Technology, 102(5), 3999 4007. Zobell, C. E. (1946). Action of microo¨rganisms on hydrocarbons. Bacteriological Reviews, 10(1 2), 1.

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Further reading Abbasian, F., Lockington, R., Megharaj, M., & Naidu, R. (2016). A review on the genetics of aliphatic and aromatic hydrocarbon degradation. Applied Biochemistry and Biotechnology, 178(2), 224 250. Agarwal, P., & Sharma, D. K. (2009). Studies on the production of biosurfactant for the microbial enhanced oil recovery by using bacteria isolated from oil contaminated wet soil. Petroleum Science and Technology, 27 (16), 1880 1893. Andreote, F. D., Jime´nez, D. J., Chaves, D., Dias, A. C. F., Luvizotto, D. M., Dini-Andreote, F., . . . de Melo, I. S. (2012). The microbiome of Brazilian mangrove sediments as revealed by metagenomics. PLoS One, 7 (6), e38600. Bao, Y. J., Xu, Z., Li, Y., Yao, Z., Sun, J., & Song, H. (2017). High-throughput metagenomic analysis of petroleum-contaminated soil microbiome reveals the versatility in xenobiotic aromatics metabolism. Journal of Environmental Sciences, 56, 25 35. Brennerova, M. V., Josefiova, J., Brenner, V., Pieper, D. H., & Junca, H. (2009). Metagenomics reveals diversity and abundance of meta-cleavage pathways in microbial communities from soil highly contaminated with jet fuel under air-sparging bioremediation. Environmental Microbiology, 11(9), 2216 2227. Bushnell, L. D., & Haas, H. F. (1941). The utilization of certain hydrocarbons by microorganisms. Journal of Bacteriology, 41(5), 653. de Lorenzo, V. (2008). Systems biology approaches to bioremediation. Current Opinion in Biotechnology, 19 (6), 579 589. DeLong, E. F., Preston, C. M., Mincer, T., Rich, V., Hallam, S. J., Frigaard, N. U., . . . Chisholm, S. W. (2006). Community genomics among stratified microbial assemblages in the ocean’s interior. Science, 311(5760), 496 503. Gogoi, B. K., Dutta, N. N., Goswami, P., & Mohan, T. K. (2003). A case study of bioremediation of petroleumhydrocarbon contaminated soil at a crude oil spill site. Advances in Environmental Research, 7(4), 767 782. Guo, J., Li, J., Chen, H., Bond, P. L., & Yuan, Z. (2017). Metagenomic analysis reveals wastewater treatment plants as hotspots of antibiotic resistance genes and mobile genetic elements. Water Research, 123, 468 478. Hasanuzzaman, M., Ueno, A., Ito, H., Ito, Y., Yamamoto, Y., Yumoto, I., & Okuyama, H. (2007). Degradation of long-chain n-alkanes (C36 and C40) by Pseudomonas aeruginosa strain WatG. International Biodeterioration & Biodegradation, 59(1), 40 43. Ibarbalz, F. M., Orellana, E., Figuerola, E. L., & Erijman, L. (2016). Shotgun metagenomic profiles have a high capacity to discriminate samples of activated sludge according to wastewater type. Applied and Environmental Microbiology, 82(17), 5186 5196. Keller, A. H., Schleinitz, K. M., Starke, R., Bertilsson, S., Vogt, C., & Kleinsteuber, S. (2015). Metagenomebased metabolic reconstruction reveals the ecophysiological function of Epsilonproteobacteria in a hydrocarbon-contaminated sulfidic aquifer. Frontiers in Microbiology, 6, 1396. Kircher, M., & Kelso, J. (2010). High-throughput DNA sequencing—concepts and limitations. Bioessays, 32(6), 524 536. Margesin, R., Ha¨mmerle, M., & Tscherko, D. (2007). Microbial activity and community composition during bioremediation of diesel-oil-contaminated soil: Effects of hydrocarbon concentration, fertilizers, and incubation time. Microbial Ecology, 53(2), 259 269. Mukherjee, A. K., & Bordoloi, N. K. (2012). Biodegradation of benzene, toluene, and xylene (BTX) in liquid culture and in soil by Bacillus subtilis and Pseudomonas aeruginosa strains and a formulated bacterial consortium. Environmental Science and Pollution Research, 19, 3380 3388. Napp, A. P., Pereira, J. E. S., Oliveira, J. S., Silva-Portela, R. C., Agnez-Lima, L. F., Peralba, M. C., . . . Vainstein, M. H. (2018). Comparative metagenomics reveals different hydrocarbon degradative abilities from enriched oil-drilling waste. Chemosphere, 209, 7 16.

Accelerated bioremediation of petroleum refinery sludge 65 Ren, Z. J. (2014). Microbial metabolism and community structure in response to bioelectrochemically enhanced remediation of petroleum hydrocarbon-contaminated soil. Environmental Science & Technology, 48(7), 4021 4029. Ruberto, L., Dias, R., Lo Balbo, A., Vazquez, S. C., Hernandez, E. A., & Mac Cormack, W. P. (2009). Influence of nutrients addition and bioaugmentation on the hydrocarbon biodegradation of a chronically contaminated Antarctic soil. Journal of Applied Microbiology, 106(4), 1101 1110. Sette, L. D., Simioni, K. C., Vasconcellos, S. P., Dussan, L. J., Neto, E. V., & Oliveira, V. M. (2007). Analysis of the composition of bacterial communities in oil reservoirs from a southern offshore Brazilian basin. Antonie Van Leeuwenhoek, 91(3), 253 266. Sierra-Garcı´a, I. N., Correa Alvarez, J., de Vasconcellos, S. P., Pereira de Souza, A., dos Santos Neto, E. V., & de Oliveira, V. M. (2014). New hydrocarbon degradation pathways in the microbial metagenome from Brazilian petroleum reservoirs. PLoS ONE, 9(2), e90087. Singh, A., Van Hamme, J. D., & Ward, O. P. (2007). Surfactants in microbiology and biotechnology: Part 2. Application aspects. Biotechnology Advances, 25(1), 99 121. Temperton, B., & Giovannoni, S. J. (2012). Metagenomics: Microbial diversity through a scratched lens. Current Opinion in Microbiology, 15(5), 605 612. Tissot, B. P. (1984). Recent advances in petroleum geochemistry applied to hydrocarbon exploration. AAPG Bulletin, 68(5), 545 563. Tissot, B. P., & Welte, D. H. (1984). Petroleum formation and occurrence (2nd ed., 699 pp.). Berlin: Springer. Todorova, N. H., Mironova, R. S., & Karamfilov, V. K. (2014). Comparative molecular analysis of bacterial communities inhabiting pristine and polluted with polycyclic aromatic hydrocarbons Black Sea coastal sediments. Marine Pollution Bulletin, 83(1), 231 240. USEPA. (2013). Integrated Risk Information System (IRIS). Washington, DC: United States Environment Protection Agency. Wang, H., Wang, B., Dong, W., & Hu, X. (2016). Co-acclimation of bacterial communities under stresses of hydrocarbons with different structures. Scientific Reports, 6, 34588. Available from https://doi.org/10.1038/ srep34588. Wang, R. F., Wennerstrom, D., Cao, W. W., Khan, A. A., & Cerniglia, C. E. (2000). Cloning, expression, and characterization of the katggene, encoding catalase-peroxidase, from the polycyclic aromatic hydrocarbon-degrading bacterium Mycobacterium sp. strain PYR-1. Applied and Environmental Microbiology, 66(10), 4300 4304. Wilkins, D., Rao, S., Lu, X., & Lee, P. K. (2015). Effects of sludge inoculum and organic feedstock on active microbial communities and methane yield during anaerobic digestion. Frontiers in Microbiology, 6, 1114. Wolicka, D., Suszek, A., Borkowski, A., & Bielecka, A. (2009). Application of aerobic microorganisms in bioremediation in situ of soil contaminated by petroleum products. Bioresource Technology, 100(13), 3221 3227. Yu, K., & Zhang, T. (2012). Metagenomic and metatranscriptomic analysis of microbial community structure and gene expression of activated sludge. PLoS One, 7(5), e38183. Zerbino, D. R., & Birney, E. (2008). Velvet: Algorithms for de novo short read assembly using de Bruijn graphs. Genome Research, 18(5), 821 829. Zhao, B., & Poh, C. L. (2008). Insights into environmental bioremediation by microorganisms through functional genomics and proteomics. Proteomics, 8(4), 874 881.

CHAPTER 3

Degradation and detoxification of waste via bioremediation: a step toward sustainable environment Komal Agrawal and Pradeep Verma Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Bandarsindri, Kishangarh, Ajmer, India

3.1 Introduction The utilization of the term sustainable development has been used in the mainstream by many companies, meaning that the process of production and the end product are sustainable, although there is still confusion as to what is a sustainable product (Baumann, Boons, & Bragd, 2002; Berchicci & Bodewes, 2005). In this chapter we have adapted the view of Ottman, Stafford, and Hartman (2006) which states that “although no consumer product has a zero impact on the environment, in business, the terms ‘green product’ or ‘environmental product’ are used commonly to describe those that strive to protect or enhance the natural environment by conserving energy and/or resources and reducing or eliminating use of toxic agents, pollution, and waste.” The production of an environmentally sustainable product by a company includes the consideration of three major factors, that is, material, energy utilization, pollution generated and its impact on the environment. The other factors to be considered are the manufacturing process, utilization of the product, and its disposal. It also has to be taken into consideration if not all the steps are environmentally friendly, for example, one of the steps includes a process which has a certain level (low to high) of impact on the environment. Various examples include packaging of eatables in plastic, which after utilization is discarded, that is, onetime use plastic, and the disposal of detergents in the form of wastewater after utilization. Thus it becomes necessary to address the challenges associated with every step of the process, that is, material, energy, and pollution generation, from different dimensions, which can bring about considerable differences in the products which are utilized in and

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00003-1 © 2020 Elsevier Inc. All rights reserved.

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disposed of into the environment (Dangelico & Pujari, 2010). Thus considering the above points, various bioremediation strategies have been developed that either degrade and/or detoxify the pollutant, thereby having less negative impact on the flora and the fauna and thus favoring a sustainable environment. Thus the present chapter deals with bioremediation methods, various pollutants and their treatments, as well as their limitations and future prospects.

3.2 Bioremediation and the role of bioavailability The notion of bioremediation gained attention a few decades back and it has become a topic of concern throughout the world. The bioremediation primarily focuses on detoxifying and/or degrading the contaminant or the pollutants, however, this step depends on the accessibility and bioavailability of the pollutants (Antizar-Ladislao, Lopez-Real, & Beck, 2005). The pollutant can gain entry into the soil, binding with the solid phase of soil via sorption, precipitation, or complexation, which renders it inaccessible for effective bioremediation. The pollutant can aggregate with the soil forming “composite units” thereby limiting the bioavailability of the pollutant to microorganisms. The concept of bioavailability and biodegradation of the pollutants depends on two factors (Singh, Sethunathan, Megharaj, & Naidu, 2008): first, the release of the pollutants from the sorbed phase to the aqueous phase for their biodegradation by the microorganisms (Harms & Zehnder, 1995; Megharaj, Ramakrishnan, Venkateswarlu, Sethunathan, & Naidu, 2011; Shelton & Doherty, 1997); and second, the occurrence of biodegradation in the sorbed phase (Singh et al., 2003). This suggests that if the pollutant becomes accessible to the microorganism degradation and/or detoxification becomes feasible and effective (Megharaj et al., 2011).

3.2.1 Surfactants Surfactants are amphiphilic in nature, containing water-soluble (hydrophilic), oil-soluble, and water-insoluble (hydrophobic) compounds. Hydrophilic groups may be anionic, cationic, nonionic, and zwitterionic. The synthetic surfactants include anionics (sulfate, sulfonate), cationics (quaternary ammonium group), nonionics (polyoxyethylene, sucrose, or polypeptide), and the hydrophobic parts of paraffins, olefins, or alcohols, etc. The most common chemicals used as surfactants are Triton X-100 and Tween 80. Surfactants are applied to the pollutant to increase the mass transfer of organic compounds for effective bioremediation (Laha, Tansel, & Ussawarujikulchai, 2009; Rosen, 1989). The significant mechanism of surfactant remediation is the minimization of the interfacial tension, the solubilization of the hydrophobic organic compound, and the phase transfer of the organic compound from a soil-sorbed to a pseudo aqueous phase (Laha et al., 2009). Surfactants increase the mobilization and biodegradation of polycyclic aromatic hydrocarbons

Degradation and detoxification of waste via bioremediation 69 generated by incomplete combustion of organic materials. Nonionic surfactants can increase the rate of degradation of naphthalene (obtained by petroleum distillation) and phenanthrene (found in cigarette ash) (Megharaj et al., 2011). Many factors are involved in the selection of surfactants for field application; they should be economically feasible, effective at minimum concentration, have low toxicity, have low adsorption in soil, lead to low soil dispersion, and have low surface tension (Mulligan, Yong, & Gibbs, 2001). Mostly the food-grade (Shiau, Sabatini, & Harwell, 1995), plant-based (Roy, Kommalapati, Mandava, Valsaraj, & Constant, 1997), or natural surfactants (Conte, Agretto, Spaccini, & Piccolo, 2005) are preferred rather than synthetic surfactants as they have a higher rate of degradation, lower toxicity, and are publicly more acceptable than synthetic surfactants. The microbial surfactants are used in food processing industries, the bioremediation of pollutants, and the cosmetic and pharmaceutical industries. (Christofi & Ivshina, 2002). The various classes of biosurfactants produced by microorganisms are many and researchers have shown interest in searching for new types of biosurfactants produced by microorganisms which can be used for bioremediation or have industrial benefits (Satpute, Banat, Dhakephalkar, Banpurkar, & Chopade, 2010). The selection of biosurfactants depends on their application, that is, the enhancement of bioavailability of persistent organic pollutants, however, this requires cautious planning based on the information about the fate and behaviors of the surfactants and their target pollutants. Many precautions should be taken prior to utilization, that is, prevention of groundwater contamination via leaching. Therefore the screening of microorganisms should be done carefully, so that it has the dual capability to produce the surfactant and to catabolize the pollutant (Megharaj et al., 2011).

3.2.2 Biodegradation Diversity exists both with respect to the pollutants that exist in the environment and the microorganisms which have the capabilities to degrade and/or detoxify the contaminants (Ramakrishnan, Megharaj, Venkateswarlu, Naidu, & Sethunathan, 2010; Ramakrishnan, Megharaj, Venkateswarlu, Sethunathan, & Naidu, 2011; Watanabe, 2001). However, it has to be considered that the microbial diversity to degrade the pollutant may be limited to certain contaminants as these are recalcitrant chemicals consisting of substituent or structural elements that are rarely present in nature (Pieper & Reineke, 2000). On the contrary, the ability of microorganisms to degrade the organic pollutant varies depending upon the type of microflora and the concentration of the pollutant, for example, mycobacteria degrade and bioremediate aged polycyclic aromatic hydrocarbons from the bound soil due to the presence of lipophilic surfaces and they also have catabolic efficiency toward polyaromatic hydrocarbons (PAHs) with up to five benzene rings (Bogan, Lahner, Sullivan, & Paterek, 2003).

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3.2.3 In situ and ex situ bioremediation The concept of bioremediation can be of two types: in situ or ex situ bioremediation. The former involves the treatment of the pollutant on the site, whereas the latter involves the transportation of the material to a different site for treatment (Aggarwal, Means, Hinchee, Headington, & Gavaskar, 1990), thus there are various advantages of in situ over ex situ bioremediation. In the case of in situ bioremediation the organic pollutant is biodegraded to an attenuated transformation product CO2 and H2O. As the treatment occurs on-site the process is cost-effective, requires low maintenance, and is an environment-friendly and sustainable approach, whereas in the case of ex situ bioremediation the pollutant is required to be transported to the treatment site, which requires extra cost, thereby making the process less economically feasible compared to in situ bioremediation. As a result of which the in situ method of treatment is preferred over the ex situ method, which would involve the effective conservation of land and water bodies (Jorgensen, 2007). In situ bioremediation can be divided into three categories: 1. Bioattenuation—“depends on the natural process of degradation”; 2. Biostimulation—“intentional stimulation of degradation of chemicals is achieved by addition of water, nutrient, electron donors, or acceptors”; and 3. Bioaugmentation—“microbial members with proven capabilities of degrading or transforming the chemical pollutants are added” (Madsen, 1991). Thus for effective bioremediation several factors need to be taken into consideration, such as the condition of the polluted site, the existing/native microflora, and the type, concentration, and toxicity of the pollutants present on the site (Megharaj et al., 2011).

3.3 The degradation and/or detoxification of pollutants The growth of industry has led to the production of pollutants which are discharged into the environment. These pollutants then contaminate the site, harm the flora and fauna, and imbalance the ecosystem. Thus in the following section, the two prominent types of pollutant, that is, heavy metals and colored/dyed effluents, and their treatment are discussed.

3.3.1 Heavy metal pollutant Industrialization has led to an upsurge in the discharge of toxic effluent into the environment, which over time has accumulated in the ecosystem. Although a sustainable environment has gained interest amongst the researchers, the industrial sector, which contributes to the economic sector of country in which it thrives, gives more importance to industrialization than environmental-related issues. Thus the industrialization and release of

Degradation and detoxification of waste via bioremediation 71 the effluent has polluted the water bodies and the soil tremendously, especially heavy metals in urban sectors. These heavy metals that have been released into the environment are toxic at low concentrations (e.g., mercury, copper, cadmium, etc.) and pose a major threat. Because heavy metals do not break down into nontoxic end products they can have a long-lasting effect on the ecosystem (Salem, Eweida, & Farag, 2000). It has also to be noted that certain heavy metals are required in small quantities for optimum growth and development, however, at higher concentrations they exert toxic effects or hinder the normal functioning of the system. Therefore it becomes a necessity for the bioremediation of the heavy metals from the environment as it will help in reestablishing a healthy and clean ecosystem. The microorganisms effectively utilize the heavy metal up to a certain concentration, however, beyond a certain limit as stated above it cause remarkable changes in the microbial community (Doelman, Jansen, Michels, & van Til, 1994) by the blockage or displacement of important functional groups and metal ions or by the changes in the active conformation of the biological molecules (Li & Tan 1994; Wood & Wang 1983). Despite the limitations, biological methods are preferred over physical and chemical methods, of which 35% and 16% constitute microbial and phytoremediation, respectively (Environmental Agency (EA), 2015; U.S. E.P.A., 2007). The degradation and/or detoxification of heavy metals by the native microflora is the most safe and effective technology to bioremediate the pollutants generated due to human activities, for example, mining, release of dye effluent (Garbisu & Alkorta, 2001). 3.3.1.1 Sources There are various sources of heavy metal pollutants. Some occur naturally, such as from the weathering of the minerals, volcanic eruption and activity along with corrosion, whereas the anthropogenic sources of heavy metal pollutants include mining, electroplating (Fig. 3.1), smelting, biosolids (manures from livestock, compost, and municipal sewage sludge), utilization of pesticides, and discharge of phosphate fertilizers (Dixit et al., 2015; Fulekar, Singh, & Bhaduri, 2009; Sumner, 2000; Wuana & Okieimen, 2011). The disturbance of the geochemical cycle of nature by the accumulation of heavy metals has resulted in soil, air, and water being polluted beyond a critical level which can pose harmful effects on the health of flora and fauna, including humans (D’amore, AlAbed, Scheckel, & Ryan, 2005). 3.3.1.2 Bioremediation of heavy metals It is well-known that microorganisms are ubiquitous in nature and are prevalent on sites with heavy metal contamination. The native microflora mineralize the organic pollutants to nontoxic end products, such as CO2 and H2O, or they are broken down into primary metabolites which can be easily utilized by the microorganisms as a primary substrate. They can degrade the pollutant either by the production of an enzyme or by developing a resistance mechanism. The removal can occur via various mechanism, for example,

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Figure 3.1 Various natural and anthropogenic sources responsible for heavy metal pollution in the environment.

biosorption, metamicrobe interaction, bioaccumulation, biotransformation, biomineralization, and bioleaching, and the various methods by which the microorganisms sustain the environment are the oxidizing, binding, immobilizing, volatizing, and/or transformation of heavy metals. The activity of the native microflora can be used in specific locations by implementing the designer microbe approach. It can be effectively implemented if the mechanisms controlling the growth and activity of the microorganisms in contaminated sites, and the metabolic activity along with the adaptability to the changes are studied in detail (D’amore et al., 2005; Dixit et al., 2015). 3.3.1.3 Adsorption The microorganisms can take up the heavy metal present in the contaminated sites by biosorption at the binding sites present in the cellular structure of the microorganisms without the involvement of energy. Of all the reactive compounds associated with the bacterial cell, extracellular polymeric substances have been of increasing interest as they have considerable effects on the acid base properties and metal adsorption (Guine´ et al., 2006). The binding of heavy metal complexes to extracellular polymeric substances can occur by various mechanisms, such as proton exchange and microprecipitation of metals (Comte, Guibaud, & Baudu, 2008; Fang et al., 2010).

Degradation and detoxification of waste via bioremediation 73 3.3.1.4 Microorganisms for the detoxification of heavy metals It has been known that heavy metals cannot be destroyed totally but can be converted to less toxic and water-soluble forms (Garbisu & Alkorta, 2001). The microorganisms mobilize the heavy metals by leaching, chelation, methylation, and redox transformation of toxic metals. The microorganisms use heavy metals and trace elements as the terminal electron acceptor or reduce the contaminants by the detoxification mechanism, thereby removing the metal contaminants present in the environment. The microorganisms derive energy from the metal redox reactions in order to bioremediate the heavy metals by enzymatic and nonenzymatic processes, thus they have developed two mechanisms, that is, resistance in bacteria (detoxification) and active efflux pumping of the toxic metal from cells (Silver & Phung, 1996). Bioremediation involves an oxidation reduction reaction between the microorganisms and pollutants. The microorganisms act as the oxidizing agent for the heavy metal, resulting in the loss of an electron which then are accepted by alternative electron acceptors (nitrate, sulfate, and ferric oxides). The mechanism differs in aerobic and anaerobic systems (Table 3.1); in the former oxygen acts as the electron acceptor and in the latter microbes oxidize organic contaminants by reducing electron acceptors (Dixit et al., 2015).

3.3.2 Dyes Dyes have been used since the ancient civilizations when the dyes were made from natural components. In the present era these have been replaced by synthetic dyes and constitute a major part of the economy and are used in the textile, pharmacy, food, and cosmetic industries (Saratale, Saratale, Chang, & Govindwar, 2011). It has been estimated that 70% of dyes belong to the azo group (Balapure, Bhatt, & Madamwar, 2015) and 17% 20% Table 3.1: Bacterial and fungal strains showing efficient biosorbence for heavy metal removal. Microbial group

Microorganism as biosorbent

Initial metal ion Metal concentration (mg/L)

Sorption capacity (mg/g)

References

Bacteria

Bacillus laterosporus

Cd Cr (VI) Cd Cr (VI) Pb

1000 1000

159.5 72.6

Zouboulis, Loukidou, and Matis (2004)

1000 1000

142.7 62

7.2

2.3

Kang et al. (2015)

Pb

350

107.1

Cu

100

34

Akar, Tunali, and Kiran (2005) Fu, Li, Zhu, Jiang, and Yin (2012)

Bacillus licheniformis Fungi

Enterobacter cloacae Botrytis cinereal Rhizopus oryzae

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Chapter 3

constitute the dye wastes (Kant, 2012), most of which reach the environment directly as effluent (Carmen & Daniela, 2012). The Indian subcontinent, China, and the European Union are home to approximately 70% of the world’s dyeing industries (Ghaly, Ananthashankar, Alhattab, & Ramakrishnan, 2013), which raises health concerns for the population residing in these countries as most of the dyes are synthetic in origin. Thus addressing the issue and generating an ecofriendly and economically feasible technique which can address all the limitations of the existing physical and chemical techniques (e.g., high cost and energy consumption, end product is not completely nontoxic, and generation of toxic sludge) is a necessity and should also be a global priority as well (Rawat, Mishra, & Sharma, 2016). 3.3.2.1 Physical and chemical removal of dye effluent The removal of dye stuff from the environment is crucial for environmental sustainability thus various physical, chemical, and biological techniques have been developed. Adsorption, flocculation, membrane filtration (Fig. 3.2), and microbial treatment, along with the advanced oxidation process (AOP), have been used for the decolorization of dye ¨ stu¨n, Birgu¨l, & Yonar, effluent (Chaudhari, Bhatt, Bhargava, & Seshadri, 2011; Solmaz, U 2009). The physical and chemical treatment process passes the dye effluent from one phase to the other without its complete removal, degradation, and/or detoxification (Erswell, Brouckaert, & Buckley, 1988). These methods have numerous drawbacks like high cost, less economic feasibility, and high energy consumption. Among the various physiochemical techniques, adsorption is the most effectively used technique (Crini, 2006) due to its ease of implementation and insensitivity to toxic pollutants. However, the major limitation is the selection of a suitable adsorbent. Once a suitable material is selected for effective decolorization, adsorption is followed by sorption, and the combination of the two techniques has been named biosorption (Gupta, Khamparia, Tyagi, Jaspal, & Malviya, 2015).

Figure 3.2 Schematic representation of various physical treatment methods employed for the removal of dyes.

Degradation and detoxification of waste via bioremediation 75 3.3.2.2 Biological treatment These treatment methods involve microorganisms, which under both aerobic and anaerobic conditions can remove a huge range of dye pollutants (Aksu, 2005). The AOP involves the generation of reactive free radicals after the oxidation (of most) of the complex pollutants which are present in the dye effluent, and converts them into nontoxic end products (Kestiog˘ lu, Yonar, & Azbar, 2005). The strain reported for the effective decolorization of the dye effluent includes Shewanella sp. strain KMK6, which significantly reduced the color and chemical oxygen demand of the dye mixture with the end products being nontoxic in nature (Kolekar et al., 2013). Other strains reported for dye decolorization include Pseudomonas fluorescens strains (Sz6 and SDz3) (Zabłocka-Godlewska, Przysta´s, & Grabi´nska-Sota, 2014) and Lysinibacillus sp. (Saratale et al., 2013). The bacterial system can be used either singly or in combination with a plant bacterial synergistic system where Glandularia pulchella (Sweet) Tronc, Pseudomonas monteilii ANK, and their consortia are used to decolorize the dye mixture and the consortia exhibits 100% decolorization (Kabra, Khandare, Waghmode, & Govindwar, 2012). Along with the bacterial system the fungal system has also been effectively used for dye decolorization. Some of the examples include thermophilic fungus. Thermomucorindicae seudaticae isolated from compost was successfully used to decolorize the azo anthraquinone dye mixture at 55 C (Taha, Adetutu, Shahsavari, Smith, & Ball, 2014). The utilization of fungal cultures has gained importance as the oxidoreductases are effectively used for dye decolorization. Thus the attention has diverted toward the development of treatment methods where an enzyme-based system is developed for the removal or transformation of the phenolic compounds present in the wastewater (Duran & Esposito, 2000; Husain & Jan, 2000). Phenols are oxidized by peroxidases and phenoxy radicals are generated, producing oligomeric and polymeric products (Ward, Hadar, Bilkis, Konstantinovsky, & Dosoretz, 2001; Ward, Hadar, & Dosoretz, 2001), which are easily precipitated. This effect increases the molecular weight followed by precipitation, thereby resulting in detoxification and acting as a method for the treatment of industrial waste which otherwise would not be possible by the conventional physical and chemical methods. Another enzyme, Mn peroxidase (Table 3.2), effectively catalyzes the oxidation of various phenols, aromatic amines, and dyes. The action of the enzyme depends on two factors, that is, divalent manganese and the type of buffer used (Duran & Esposito, 2000; Ehara, Tsutsumi, & Nishida, 1998). Laccase reduces oxygen to water with the simultaneous oneelectron oxidation of many aromatic substrates, such as phenols and aromatic amines (Bourbonnais, Paice, Freiermuth, Bodie, & Borneman, 1997; Robles, Lucas, de Cienfuegos, & Ga´lvez, 2000). The four strains Trametes modesta, Trametes hirsuta, Trametes versicolor, and Sclerotium rolfsii were compared for their ability to produce laccase. Further they were also used for the decolorization of synthetic dyes, that is, anthraquinone,

76

Chapter 3 Table 3.2: The potential of various enzymes for the bioremediation of various dyes.

Enzyme

Source

Dye

Laccase

Trichoderma asperellum

Malachite Green

Peroxidase

Trametes Acid Black 172 pubescens Oudemansiella Congo Red canarii Soybean hulls Methyl Orange Soybean

Manganese peroxidase

Lignin peroxidase (LiP)

Citrus limon Ganoderma lucidum IBL-05 Ganoderma lucidum Irpex lacteus F17 Schizophyllum commune Pleurotus ostreatus (PLO9) Raoultella ornithinolytica OKOH-1

Decolorization (%) 97.18

References

68.84 6 6.68

Shanmugam, Ulaganathan, Swaminathan, Sadhasivam, and Wu, (2017) Zheng et al. (2016)

80

Iark et al. (2019)

81.4

Chiong, Lau, Lek, Koh, and Danquah (2016) Alneyadi and Ashraf (2016)

2100 Mercaptobenzothiazole Direct Yellow 4 89.47 Sandal reactive dyes 78.14 92.29

Nouren et al. (2017) Bilal and Asgher (2015)

Drimaren Blue CL-BR

70

Malachite Green

96

Xu, Guo, Gao, Bai, and Zhou (2017) Yang, Zheng, Lu, and Jia (2016)

Sandal-fix Foron blue E2BLN RBBR

89.71

Sofia et al. (2016)

40

Oliveira, da Luz, Kasuya, Ladeira, and Junior (2018)

Congo Red

65.03

Falade, Mabinya, Okoh, and Nwodo (2019)

azo, indigo, and triarylmethane. The decolorization of the dye depends on the source of enzyme preparation and the structure of dye (Nyanhongo et al., 2002). Lignin peroxidase (LiP) has the capability to mineralize various recalcitrant aromatic compounds and oxidize various polycyclic aromatic and phenolic compounds (Karam & Nicell, 1997; Wesenberg, Kyriakides, & Agathos, 2003).

3.4 Role of genetic engineering in bioremediation Conventional methods have been used for the decolorization of synthetic dyes, however, biological methods are preferred over the prior methods. The biological system has been used for in situ or ex situ bioremediation. The enzymatic machinery of the enzyme also has its utility in the degradation and/or detoxification of the toxic pollutant. However, there are certain limitations which can be overcome by the application of systems biology (SB) in different sectors of bioremediation (Dangi, Sharma, Hill, & Shukla, 2019).

Degradation and detoxification of waste via bioremediation 77

3.4.1 Bioremediation through microbial systems biology The approach of SB is new and has tremendous potential as it is used to study the biological system and explore the intricate networks and their interaction during different biological processes at various levels, that is, molecular, cellular, population, community, and ecosystem levels (Pavlopoulos et al., 2015). This technology can be effectively used for the better implementation of environmental bioremediation (Chakraborty, Wu, & Hazen, 2012; Noble, Pro¨schel, & Mayer-Pro¨schel, 2011). It can provide substantial information on gene expression, biosynthetic pathways, proteins, and secondary metabolites both with and without stress as a result of which better understanding and application would be possible. The most commonly used techniques are the “omics” which comprises genomics, proteomics, transcriptomics, and metabolomics (Banerjee, Singh, & Shukla, 2016; Dangi, Dubey, & Shukla, 2017; Kumar & Shukla, 2014; Rayu, Karpouzas, & Singh, 2012). 3.4.1.1 Genomics Bioremediation, when implemented, has certain limitations such as the cellular pathway, proteins, and the encoded genes of microorganisms used during mineralization are not known (Zhao & Poh, 2008). Thus the concept of genomics comes into play where it provides detailed information on the genetic material (DNA and RNA) when the microorganisms are under stress or in the presence of the pollutant. The concept of genomics involves genome sequencing and bioinformatic studies using various tools. It has been estimated that the genomes of 270,567 organisms have been sequenced, and more than 46,000 projects related to genome sequencing are in progress globally (http://www. genomesonline.org) (Dangi et al., 2019). 3.4.1.2 Metagenomics It has been known that only a small part of the microflora is culturable and the rest are categorized as uncultured microflora. The study of the uncultured microorganisms has been undertaken around the world. However, with the advancement of technology the study of metagenomics involves direct DNA isolation from microbial samples of the environment that are later sequenced and analyzed without obtaining a pure culture in the lab (Baweja, Nain, Kawarabayasi, & Shukla, 2016; Garza & Dutilh, 2015). Thus metagenomics can help to study and characterize the unknown microflora, in order to analyze degradation pathways which were previously unknown. The information so obtained can then be used for the further studying of impacts on the biodegradation of various pollutants and to design a more competent strain for its effective application in bioremediation. 3.4.1.3 Transcriptomics The study of transcriptomics is the study of “differential expression of genes which are upregulated and downregulated in response to environmental pollutants.” The function of

78

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previously unannotated genes can be inferred by transcriptomics and techniques such as microarray and RNA sequencing. In order to study transcriptomes of the microbial community it requires the isolation and enrichment of cellular mRNA, second synthesis of cDNA from mRNA, and lastly either sequencing of complete cDNA or the use of microarrays for the hybridization of cDNA (Zhu, Ma, Jin, Wu, & Sun, 2017). 3.4.1.4 Proteomics and metaproteomics The study of proteomics is more complex than the genomics and transcriptomics, as the microorganism’s genome remains relatively constant, however, the proteome differs from “cell to cell” and “time to time” (Nzila et al., 2018). The study of proteins in the bacterial system (proteomics) and in the environmental sample (metaproteomics) can be used for the effective detection and identification of the proteins that are important to the microorganisms and can be effectively used for the bioremediation of the pollutants as well. Metaproteomics can also be used for the bacterial samples which are not cultured in the laboratory (Aydin et al., 2017; Herbst et al., 2016), that is, environmental samples which will represent the native population existing in the ecosystem (Wang, Kong, Li, & Xie, 2016).

3.5 Limitations and future prospect Bioremediation of pollutants from the environment has become a major concern, however, certain limitations are associated with various methods. The physical and chemical treatments involve a high amount of energy, cost, and the end product generated is not completely nontoxic. On the other hand, using the biological method native microorganisms can bioremediate the pollutant more effectively, however, the major drawback is that most of the microorganisms are nonculturable in the lab and the detailed mechanism of the enzymatic machinery is not well documented. The above limitations can be overcome by the implementation of SB. These new technologies can be used for the detailed study of the nonculturable microflora inhabiting the polluted sites and the metabolic pathways which are responsible for the effective biodegradation of pollutants.

Acknowledgment PV is thankful to DBT (Grant No. BT/304/NE/TBP/2012; Grant No. BT/PR7333/PBD/26/373/2012), KA is thankful to Central University of Rajasthan, Ajmer, India for providing financial support.

Competing interests All the authors declare that they have no competing interests.

Degradation and detoxification of waste via bioremediation 79

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Degradation and detoxification of waste via bioremediation 83 Sofia, P., Asgher, M., Shahid, M., & Randhawa, M. A. (2016). Chitosan beads immobilized Schizophyllum commune IBL-06 lignin peroxidase with novel thermo stability, catalytic and dye removal properties. Journal of Animal and Plant Sciences, 26(5), 1451 1463. ¨ stu¨n, G. E., Birgu¨l, A., & Yonar, T. (2009). Advanced oxidation of textile dyeing effluents: Solmaz, S. A., U Comparison of Fe 1 2/H2O2, Fe 1 3/H2O2, O3 and chemical coagulation processes. Fresenius Environmental Bulletin, 18(8), 1424 1433. Sumner, M. E. (2000). Beneficial use of effluents, wastes, and biosolids. Communications in Soil Science and Plant Analysis, 31(11-14), 1701 1715. Taha, M., Adetutu, E. M., Shahsavari, E., Smith, A. T., & Ball, A. S. (2014). Azo and anthraquinone dye mixture decolourization at elevated temperature and concentration by a newly isolated thermophilic fungus, Thermomucor indicae-seudaticae. Journal of Environmental Chemical Engineering, 2(1), 415 423. U.S. E.P.A. (2007). Treatment technologies for site cleanup: Annual Status Report. Washington, DC: United States Environmental Protection Agency (EPA). Wang, D. Z., Kong, L. F., Li, Y. Y., & Xie, Z. X. (2016). Environmental microbial community proteomics: Status, challenges and perspectives. International Journal of Molecular Sciences, 17(8), 1275. Ward, G., Hadar, Y., Bilkis, I., Konstantinovsky, L., & Dosoretz, C. G. (2001). Initial steps of ferulic acid polymerization by lignin peroxidase. Journal of Biological Chemistry, 276, 18734 18741. Ward, G., Hadar, Y., & Dosoretz, C. G. (2001). Inactivation of lignin peroxidase during oxidation of the highly reactive substrate ferulic acid. Enzyme and Microbial Technology, 29, 34 41. Watanabe, K. (2001). Microorganisms relevant to bioremediation. Current Opinion in Biotechnology, 12(3), 237 241. Wesenberg, D., Kyriakides, I., & Agathos, S. N. (2003). White-rot fungi and their enzymes for the treatment of industrial dye effluents. Biotechnology Advances, 22(1-2), 161 187. Wood, J. M., & Wang, H. K. (1983). Microbiol resistance to heavy metals. Environmental Science &Technology, 17, 582 590. Wuana, R. A., & Okieimen, F. E. (2011). Heavy metals in contaminated soils: A review of sources, chemistry, risks and best available strategies for remediation. ISRN Ecology, 2011. Article ID 402647. Xu, H., Guo, M. Y., Gao, Y. H., Bai, X. H., & Zhou, X. W. (2017). Expression and characteristics of manganese peroxidase from Ganoderma lucidum in Pichia pastoris and its application in the degradation of four dyes and phenol. BMC Biotechnology, 17(1), 19. Yang, X., Zheng, J., Lu, Y., & Jia, R. (2016). Degradation and detoxification of the triphenylmethane dye malachite green catalyzed by crude manganese peroxidase from Irpex lacteus F17. Environmental Science and Pollution Research, 23(10), 9585 9597. Zabłocka-Godlewska, E., Przysta´s, W., & Grabi´nska-Sota, E. (2014). Decolourisation of different dyes by two Pseudomonas strains under various growth conditions. Water, Air, & Soil Pollution, 225(2), 1846. Zhao, B., & Poh, C. L. (2008). Insights into environmental bioremediation by microorganisms through functional genomics and proteomics. Proteomics, 8(4), 874 881. Zheng, F., Cui, B. K., Wu, X. J., Meng, G., Liu, H. X., & Si, J. (2016). Immobilization of laccase onto chitosan beads to enhance its capability to degrade synthetic dyes. International Biodeterioration & Biodegradation, 110, 69 78. Zhu, Y., Ma, N., Jin, W., Wu, S., & Sun, C. (2017). Genomic and transcriptomic insights into calcium carbonate biomineralization by marine actinobacterium Brevibacterium linens BS258. Frontiers in Microbiology, 8, 602. Zouboulis, A. I., Loukidou, M. X., & Matis, K. A. (2004). Biosorption of toxic metals from aqueous solutions by bacteria strains isolated from metal-polluted soils. Process Biochemistry, 39(8), 909 916.

Further reading Kile, D. E., & Chiou, C. T. (1989). Water solubility enhancements of DDT and trichlorobenzene by some surfactants below and above the critical micelle concentration. Environmental Science & Technology, 23 (7), 832 838.

CHAPTER 4

Fungal laccases: versatile green catalyst for bioremediation of organopollutants Ajit Patel1, Vanita Patel2, Radhika Patel3, Ujjval Trivedi3 and Kamlesh Patel3 1

J. & J. College of Science, Nadiad, India, 2V.P & R.P.T.P. Science College, Vallabh Vidyanagar, India, 3P. G. Department of Biosciences, Sardar Patel University, Vallabh Vidyanagar, India

4.1 Introduction Enzymes are proteinic in nature and function as specialized catalysts during chemical reactions. Their ability to perform very specific chemical transformations has made them increasingly useful in industrial processes (Li, Yang, Yang, Zhu, & Wang, 2012). Microbial enzymes have more widespread applications in industries and medicine due to their stability, catalytic activity, and ease of production and optimization than plant and animal enzymes. The enzymes have found profound utility in various industries (e.g., food, agriculture, chemicals, and pharmaceuticals) due to the reduced processing time and low energy input, as well as their cost-effectiveness, nontoxicity, and ecofriendly characteristics. Enzymatic oxidations have great potential to substitute for chemical methods because they are very specific and efficient catalysts as well as ecologically more sustainable. They are also capable of degrading toxic chemical compounds of industrial and domestic wastes (phenolic compounds, nitriles, amines, etc.) either via degradation or conversion (Singh, Kumar, Mittal, & Mehta, 2016). As compared to chemical reactions, enzymatic reactions exhibit a higher level of catalytic efficiency, more specificity, an absence of side-reactions, and they are also biodegradable, easily removed from contaminated streams, easily standardized in commercial preparations, and generally operate at mild conditions of temperature, pressure and pH. Oxidoreductases are enzymes that catalyze the transfer of electrons from a donor to an acceptor (one substrate to another). Such enzymes are valuable for industrial applications because they allow the use of water as a solvent and facilitate regiostereo- and enantioselective reactions, so they have wide applications in the industries like textile, paper and pulp, pharmaceuticals, agrochemicals, polymers, biofuels, amino acids,

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00004-3 © 2020 Elsevier Inc. All rights reserved.

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nutraceuticals and cosmetics. One group of oxidoreductase enzymes uses oxygen as a terminal electron acceptor. Oxygen is a highly reactive molecule containing lots of energy. Nature has provided complex enzymes which efficiently harness the energy present in the oxygen molecule, while maintaining the formation of reactive oxygen species using the least energy. Among the different existing oxidases, laccases have been the choice for study as a more ecofriendly option in comparison to the existing chemical processes. Laccase is one of the enzymes that have been studied since the 19th century. In 1883 Yoshida was the first to extract it from the exudates of Rhus vernicifera (Japanese lacquer tree), from which the designation laccase was derived. In 1896 Bertrand and Laborde first demonstrated the presence of laccase in fungi (Desai & Nityanand, 2011). Laccases (EC 1.10.3.2, p-diphenol: dioxygen oxidoreductase) belong to the blue multicopper oxidases which have innate properties to produce reactive radical molecules. They are able to catalyze one-electron oxidation of a wide variety of organic and inorganic substrates using oxygen as the final electron acceptor. Generally enzymes have very high substrate specificity but oxidases like laccases are rather unspecific. Laccases oxidize both phenolic and nonphenolic compounds, so they can be used for the bioremediation of environmental pollutants present in nature (Bollag, Shuttleworth, & Anderson, 1988; Singh, Sharma, & Capalash, 2009; Singh et al., 2016). Laccases have been widely used in many industries (textile, paper, pulp, petrochemical, food processing, medical and health care, design of biosensors, and nanotechnology) for various purposes. Fungal laccases are more attractive due to their ability to catalyze a wide variety of organic compounds, including polymeric lignin and humic substances (Baldrian, 2006). Generally ligninolytic fungi produce at least one laccase isoenzyme and laccases are the principal ligninolytic enzymes present in the soil. In addition, laccase-mediated delignification of agroindustrial products increases the nutritional value of animal feed and soil fertilizer (Gonzalez et al., 2013). Laccases require only molecular oxygen for catalysis so they are suitable for biotechnological applications like the transformation or immobilization of xenobiotic compounds in the environment (Couto & Herrera, 2006). Laccases have been evaluated for a large number of biotechnological applications like dye degradation, bioremediation of some toxic industrial wastes (chlorinated aromatic compounds, polycyclic aromatic hydrocarbons (PAHs), nitroaromatics, and pesticides), and development of biosensors (Couto & Herrera, 2006; Gonzalez et al., 2013; Sa´nchez et al., 2010). Laccases have been commercially used for the delignification of wood material and to produce ethanol. Research in recent years has been intense due to the wide variety of laccases available, their applications, and their attractive properties. This chapter aims to focus on the potential applications of laccase enzymes with special reference to the bioremediation of synthetic textile dyes.

Fungal laccases: versatile green catalyst for bioremediation of organopollutants 87

4.2 Distribution and physiological functions of laccases Laccases are widespread in nature and are generally found in plants, fungi, in some bacteria and insects (Minussi et al., 2007). They can be extracellular or intracellular; their physiological functions are different in the different organisms but they all have common catalytic properties, either polymerization or depolymerization (Riva, 2006). One of the earliest enzymes ever described was laccase which is characterized as a copper-containing oxidase (Bertrand, 1985). Laccases have been discovered in numerous plants but their detailed studies are difficult because the crude extracts contain a mixture of oxidative enzymes with broad substrate specificities (Ranocha et al., 1999), so the detailed information of plant laccase is limited. R. vernicifera laccase has widely been used for the investigation of the mechanism of their action (Battistuzzi, Di Rocco, Leonardi, & Sola, 2003; Johnson, Thompson, Brinkmann, Schuller, & Martin, 2003). Plant laccases are present in the xylem where they polymerize monolignols into dimers and trimers during the early stages of lignification (Gavnholt & Larsen, 2002) and are also responsible for the synthesis of radical-based lignin polymer (Hoopes & Dean, 2004). Laccases have been shown to be involved in the defense mechanism of plants, for example, healing in wounded leaves (de Marco & Roubelakis, 1997). Very few bacterial laccases have been described; the first bacterial laccase to be detected in the plant root bacterium Azospirillum lipoferum is responsible for melanin synthesis (Givaudan et al., 1993). The role of thermostable CotA laccase produced by Bacillus subtilis is to produce pigment present in the coat of the endospore (Martins et al., 2002). Streptomyces cyaneus (Arias et al., 2003) and Streptomyces lavendulae (Suzuki et al., 2003) also possess laccases but they are not generally grouped specific in prokaryotes while in eukaryotes they are specific to white rot fungi (Claus, 2003). Laccase-like proteins present in bacteria are intracellular or periplasmic proteins (Baldrian, 2006) having different physiological functions. Laccase of Bacillus licheniformis protects the strain from ultraviolet (UV) light and oxidants (Dalfard et al., 2006), as well as being involved in the dimerization of phenolic acids (Koschorreck et al., 2008). Laccases produced by Bacillus during endospore formation are involved in phenol degradation (Dwivedi, Singh, Pandey, & Kumar, 2011; Naclerio et al., 2010). More than 60 fungal strains (Basidiomycetes, Ascomycetes, and Deuteromycetes), with high redox potential, are used for the production of biotechnologically important laccases (Upadhyay, Shrivastava, & Agrawal, 2016). Laccase activity has been demonstrated, optimized, and purified in many fungal species belonging to Ascomycetes and Basidiomycetes, such as Melanocarpus albomyces (Kiiskinen, Viikari, & Kruus, 2002), Cerrena unicolour (Kim, Cho, Eom, & Shin, 2002), Magnaporthe grisea (Iyer & Chattoo, 2003), Trametes versicolour (Minussi et al., 2007), Trichoderma reesei (Levasseur et al., 2010), and

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Xylaria polymorpha (Bittner et al., 2012). Production of laccase was also demonstrated in terrestrial Ascomycete species (Aspergillus, Curvularia, and Penicillium) (Scherer & Fischer, 1998) as well as in some aquatic water Ascomycetes (Junghanns, Moeder, Krauss, Martin, & Schlosser, 2005). True laccase was also purified from the human yeast pathogen Cryptococcus (Filobasidiella) neoformans which can oxidize phenols and aminophenols but is unable to oxidize tyrosine (De Jesus, Nicola, Rodrigues, Janbon, & Casadevall, 2009). Lignin-degrading white-rot fungi (WRF) (Basidiomycetes) and litter-decomposing saprophytic fungi are the highest producers of laccase (Shekher, Sehgal, Kamthania, & Kumar, 2011) and the enzyme has been purified from many species (Hatakka, 2001). Most of the laccases characterized so far have been produced by efficient lignin-degrading WRF (Kiiskinen, Ra¨tto¨, & Kruus, 2004) like Agaricus bisporus (Wood, 1980), Botrytis cinerea (Marbach, Harel, & Mayer, 1984), Coprinus cinereus (Schneider et al., 1999), Phlebia radiata (Niku-Paavola, Karhunen, Salola, & Raunio, 1988), Pleurotus ostreatus (Sannia, Giardina, Luna, Rossi, & Buonocore, 1986), and T. versicolour (Rogalski, Lundell, Leonowicz, & Hatakka, 1991). Several laccaseencoding genes are present in fungi but their physiological roles are not well understood (Mander et al., 2006). Fungal laccases are known as ideal green catalysts because they require air (source of oxygen) and produce water as the only by-product. They have great biotechnological significance due to their few requirements and wide substrate specificity, including direct bioelectrocatalysis (Kunamneni, Ghazi, et al., 2008; Kunamneni, Plou, Ballesteros, & Alcalde, 2008). Laccases or laccase-like activities are also found in some insects, where they have been suggested for the activation of cutical sclerotization (Dittmer et al., 2004). They have been reported in termites as a candidate for lignin-modifying enzymes (lignases).

4.3 Production of laccases It is necessary to produce large amount of laccases at very low cost to evaluate their biotechnological and environment-related applications and to understand their molecular and kinetic properties. Optimization of the fermentation medium gives a higher yield of enzyme at a very low expense. Laccases are secondary metabolites. The physiological requirements of different WRF are also different, so considerable research has been carried out on the effects of various factors like carbon source, concentration and type of nitrogen source, pH, temperature, inducer, and other culture conditions for laccase production (Wesenberg, Kyriakides, & Agathos, 2003).

4.3.1 Screening of laccase-producing fungi Screening of laccase-producing fungal species is important in order to select the appropriate laccase-producing organisms. An ideal screening process should be inexpensive, rapid,

Fungal laccases: versatile green catalyst for bioremediation of organopollutants 89 and must be able to identify fungal strains and enzymes that will work under various conditions (Monteiro & Carvalho, 1998). For industrial applications, it is necessary to decipher novel laccases with broad substrate specificities and improved stabilities. Laccaseproducing fungi have been isolated on solid media with chromomeric indicators that allow the visual detection of laccase production (Nishida, 1988). Laccases can oxidize a wide variety of substrates, thus providing a broader scope for indicators used during screening. Initially tannic acid and gallic acid were used but recently they have been replaced by synthetic phenolic compounds (guaiacol and syringaldazine) (De Jong, De Vries, Field, van der Zwan, & de Bont, 1992) or by dyes like Remazol Brilliant Blue R (RBBR) and poly R478 (Raghukumar, D’Souza, Thorn, & Reddy, 1999). Laccase-producing fungi decolorized RBBR and Poly R-478 and showed a colorless halo around their growth while there was a development of a reddish brown halo with guaiacol- (Nishida, 1988) and a dark-brown colored halo in tannic acid and gallic acid-containing agar plates (Harkin & Obst, 1973) indicating a positive reaction. Kiiskinen et al. (2004) used polymeric dyes, guaiacol, and tannic acid for the screening of novel ligninolytic fungi by agar plate method.

4.3.2 Cultural and nutritional conditions for laccase production Cultural conditions and medium composition play an important role in the expression of laccase. Laccases are secondary metabolites of fungi and their production is strongly affected by time of cultivation, pH, temperature, static or submerged cultures, aeration (Dekker & Barbosa, 2001), presence of inorganic compounds, degradation or activation by protease (Palmieri et al., 2001), as well as the type and concentrations of carbon source, nitrogen source, and inducer (Palmieri, Giardina, Bianco, Fontanella, & Sannia, 2000). The physiological demands also vary among the WRF, so extensive research has been done to investigate the abovementioned factors. Dong, Zhang, Zhang, Huang, and Zhang (2005) used 12 different media under static or shaking conditions for the production of laccase by Trametes gallica and reported that in addition to static and shaking conditions, the composition of the medium largely influences the amount and pattern of laccase isoenzymes’ production. Gayazov and Rodakiewicz-Nowak (1996) reported that semicontinuous cultivation with high aeration and agitation reduced the time of laccase production as compared to static conditions. Piscitelli et al. (2011) also described that various physiological factors can influence the production of laccase in a number of WRF. An optimized medium is essential for producing higher amounts of laccases and may be used for biochemical analysis and industrial application. Extracellular laccases of WRF may be inducible or constitutive and exist as isoenzymes, for example, Ganoderma lucidum produced more than three laccase isoenzymes in liquid culture (Ko, Leem, & Choi, 2001). Pleurotus pulmonarius synthesized three laccase isoenzymes (lcc1, lcc2, and lcc3) among

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which lcc3 was found only in induced-culture filtrates while the lcc1 and lcc2 isoforms were found in noninduced cultures (De Souza et al., 2004). More than seven laccase isozymes were produced by Basidiomycete CECT 20197 (Mansur, Sua´rez, Ferna´ndezLarrea, Brizuela, & Gonzalez, 1997), similarly four induced and three constitutive laccase isozymes were produced by Marasmius querocophilus strain 17 (Farnet, Criquet, Tagger, Gil, & Petit, 2000). The pattern of isozyme production has been successfully used for the identification of different fungi such as Ectomycorrhiza (Courty et al., 2008), Deuteromycetes (Brijwani, Rigdon, & Vadlani, 2010), and Basidiomycetes (Tanesaka, 2012; Zervakis, Venturella, & Papadopoulou, 2001). Zymogram is generally used for the study of isoenzyme patterns in different sources (Micales, Bonde, & Peterson, 1992). Praveen et al. (2011) reported that the production of the laccase enzyme was independent of biomass yields but it was found to be highly dependent on the cultivation conditions of fungus. Kues and Ruhi (2011) suggested that the synthesis and activity of laccases were controlled during growth and can play a vital role in fruiting body and pigment formation. Production of laccase strongly relies on the nature and amounts of nutrients like nitrogen and trace elements in the growth medium (Buswell, Cai, & Chang, 1995). Fungi generally produce laccases in very low concentrations (Vasconcelos, Barbosa, Dekker, Scarminio, & Rezende, 2000) but the addition of various supplements in media can induce higher yields of laccase. Production of laccase in P. chrysosporium was not detected in media with varying nitrogen concentrations along with glucose as a carbon source, but it was detected when glucose was replaced by cellulose as a carbon source (Srinivasan, Dsouza, Boominathan, & Reddy, 1995). White-rot fungus G. lucidum produced higher levels of laccase in a medium having high nitrogen concentrations and glucose as a carbon source (D’Souza, Merritt, & Reddy, 1999). The ligninolytic systems in fungi were expressed during the secondary metabolic phase and were often triggered when the concentration of nitrogen (Buswell et al., 1995), carbon, or sulfur became limited (Heinzkill, Bech, Halkier, Schneider, & Anke, 1998). Production of the laccase was also influenced by addition of lignin as well as xenobiotic compounds such as xylidine and veratryl alcohol (Xavier, Evtuguin, Ferreira, & Amado, 2001). Sinegani, Emtiazia, and Hajrasuliha (2000) maximized the laccase production in liquid culture media using N-ethyl aniline, N, N-dimethyl aniline, and parabromoaniline as a laccase inducer by Aspergillus terreus, Armillaria sp., Polyporus sp., and P. chrysosporium. Periasamy and Palvannan (2010) observed an increased yield of laccase during a dye decolorization experiment, signifying an improvement in the percentage (45.0% to 84.6%) of decolorization of Reactive Blue 221 during the same time period. More than one laccase gene is present on several genomes of WRF (Collins & Dobson, 1997) but their expression levels are affected by cultivation conditions (Sethuraman, Akin, Eisele, & Eriksson, 1998), for example, high nitrogen content of the medium induces expression of laccase genes in the Basidiomycete I-62 (CECT 20197) and in Pleurotus

Fungal laccases: versatile green catalyst for bioremediation of organopollutants 91 sajor-caju (Sethuraman et al., 1998). It has been suggested that fungi produced more laccase in the presence of copper to overcome the oxidative stress caused by the free copper ions present in medium (Berka, Brown, Xu, Schneider, & Aaslyng, 1998). Many other metal ions such as Mg21, Cd21, or Hg21 can also induce laccase production (Berka et al., 1998). Aromatic compounds that are structurally related to lignin precursors (2,5, xylidine and ferulic acid) have also been found to stimulate laccase gene expression in Trametes villosa, T. versicolour, and P. sajor-caju (Sethuraman et al., 1998). Constitutively expressed laccase genes are also present in T. villosa and P. sajor-caju and they may have different physiological roles in the fungi (Sethuraman et al., 1998). Overproduction of laccase enzyme is always related to the higher expression of laccase genes in hosts, and protein engineering is also used to obtain more robust and active enzymes.

4.3.3 Heterologous production of laccases The wild-type fungi generally produce very low yields of laccases, therefore cloning of laccase genes and heterologous expression is employed to improve the production of laccase for commercial purposes. Advancement in the field of genetic engineering makes it possible to develop efficient expression vectors for the production of functional laccase. The most commonly used organisms for the cloning of the laccase gene are P. pastoris (Berka et al., 1998), Aspergillus oryzae (Record et al., 2002), A. niger (Sigoillot et al., 2004; Soden, O’callaghan, & Dobson, 2002), Aspergillus nidulans (Kurtz & Champe, 1982), T. reesei (Paloheimo et al., 2006), and Yarrowia lipolitica (Madzak et al., 2005). The production levels of laccase have often been enhanced considerably by expression in heterologous hosts, but the number of reports for industrial utilization is yet not sufficient (Table 4.1). Incorrect folding and inefficient codon usage of expression organisms are the most common problems associated with heterologous expression of fungal enzymes, which results in nonfunctional or low yields of enzyme. The incorrect replacement of carbohydrate molecules during glycosylation of laccase in the expression organism may cause an additional problem to heterologous expression. Production of heterologous laccase in yeast has been improved very often by varying the cultivation conditions like controlling the pH of the culture medium and by lowering cultivation temperatures (Liu, Chao, Liu, Bao, & Qian, 2003; Cassland & Jo¨nsson, 1999). It was proposed by Liu et al. (2003) that maintenance of the medium’s pH above 4.0 is crucial for the stability of secreted laccases and inactivation of other proteases, whereas low temperatures may improve the folding of heterologous proteins which results in better production (Aalto, Ronne, & Kera¨nen, 1993). The presence of copper in the culture medium has also proved to be essential for expressing heterologous laccase production in P. pastoris and Aspergillus sp. (Cassland & Jo¨nsson, 1999; Uldschmid, Dombi, & Marbach, 2003). Adaptation of a directed evolution strategy has aided in improving laccase production by heterologous organisms. Mutations in the

Table 4.1: Literature comparison of laccase production by various white-rot fungi (WRF) and recombinants (Kandasamy et al., 2016). Organism

Major components in the medium

Inducer

Laccase yield (U/mL)

Trametes pubescens

Glucose 40 g/L and peptone 10 g/L

2 mM CuSO4 1 mM gallic acid 1 mM 2,5, xylidine Ethanol 35 g/L ABTS 1 mM

330 (ABTS) 350 (ABTS) 275 (ABTS) 266 (ABTS) 400 (ABTS)

0.15 mM CuSO4 and 500 mM ethanol 1 mM CuSO4


100 (DMP)

Pycnoporus cinnabarinus Maltose 20 g/L Pleurotus ostreatus Glucose 10.0 g/L, peptone 0.5 g/L and 50 mg/L Vit. B1 Pycnoporus coccineus Glucose 10 g/L Tramets sp. Trichoderma sp. Cerrena unicolour Coriolopsis rigida

Xylose 15 g/Land tryptone 0.15 g% Tomato juice medium

Ganoderma sp.

Glycerol 40 g/L

Lentinus strigosus

Glucose g L

S. ochraceum

Glucose 20 g/L,

WR1

Pleurotus ostreatus

Glucose Starch 2 g% Starch 2 g% Starch 2 g% Glucose 10.5 g/Land yeast extract 5 g/L

2 mM CuSO4 and 10 mM xylidine 0.85 mM Veratryl alcohol 2 mM CuSO4 and 2,6, dimethyl phenol 2 mM CuSO4 and 2,4-dimethyl phenol

2 mM CuSO4 0.8 mM Xylidine 0.25 g% CuSO4

Lentinus tigrinus

Birch sawdust 20 g/L

1 g% Butanol

Cerrena unicolour

Ethanol production residue (40 g/L),

Cerrena sp.

4 g/L PDB and 5 g/L soytone

0.5 mM trinitrotoluene (TNT) 0.4 M CuSO4 and 2 mM 2,5 xylidine

Barley bran 50 g/L

184 (ABTS) 19 (ABTS) 40 (ABTS) 240 (ABTS) 186 (ABTS)

References Galhaup, Goller, Peterbauer, Strauss, and Haltrich (2002) Lomascolo et al. (2003) Hou, Zhou, Wang, Du, and Yan (2004) Jaouani, Guille´n, Penninckx, Martı´nez, and Martı´nez (2005) Zhang et al. (2006) Michniewicz, Ledakowicz, Ullrich, and Hofrichter (2008) ´mez, Pazos, and Alca´ntara, Go ´n (2007) Sanroma Teerapatsakul, Parra, Bucke, and Chitradon (2007) Myasoedova, Chernykh, Psurtseva, Belova, and Golovleva (2008)

33 (ABTS) 124 (ABTS) 288 (ABTS) 410 (ABTS) 692 (ABTS) 150 (DMP)

Revankar and Lele (2006)

´nchez, Loera, Tlecuitl-Beristain, Sa Robson, and Dı´az-Godı´nez (2008) Kadimaliev et al. (2008)

24 (pyrocatechin) 165 (ABTS) Elisashvili and Kachlishvili (2009) 202 (ABTS)

Chen et al. (2012)

Lentinus sp.

2.4 g/L PDB and 5 g/L soytone

58 (ABTS)

Chen et al. (2012)

Starch 20 g/Land yeast extract 4 g/L

0.4 M CuSO4 and 2 mM 2,5 xylidine 0.6 mM CuSO4

Shiraia sp. Super H2168

101 (ABTS)

Yang, Ding, Liao, and Cai (2013)

Glucose 53 g/L and tannic acid 25 g/L

0.005 g% CuSO4

770 (ADPB)

Yeast nitrogen base with ammonium sulfate 13.4 g/L and biotin 400 mg/L Yeast extract 10 g/L, peptone 20 g/L

0.1 mM CuSO4 and 0.5 g% methanol 0.2 mM CuSO4 and 0.8 g% methanol 0.3 mM CuSO4 and 0.6 g% alanines 100 mM CuSO4

140 (ABTS) 12.6 (ABTS)

Hatamoto, Sekine, Nakano, and Abe (1999) ¨nsson Hong, Meinander, and Jo (2002) Guo et al., 2008

83 (ABTS)

Hong, Zhou, Tu, Li, and Xiao (2007)

2.84 (SGZ)

Maupin-Furlow et al. (2005)

217 (ABTS) 1259 (ABTS) 1944 (ABTS) 5671.13 (ABTS)

Kandasamy et al. (2016)

Recombinant laccases Schizophyllum commune in Aspergillus sojae Trametes versicolour in Pichia pastoris Trametes versicolour in Pichia methanolica Trametes sp. 420 in Pichia pastoris Recombinant Haloferax volcanii Hexagonia hirta MSF2

Methanol Yeast
peptone
casamino acids PDB PDB Glucose 10.5 g/L and yeast extract 5 g/L Glucose 10.5 g/L and yeast extract 5 g/L

1 mM CuSO4 500 mM CuSO4 500 mM CuSO4 500 mM CuSO4 1 1 mM 2,5 xylidine

Substrate used in the assay is given in parenthesis. ABTS, 3-Ethylbenzthiazoline-6-sulfonate; ADPB, 4-amino-2,6-dibromophenol; DMP, dimethoxyphenol; PDB, Protein Data Bank.

94

Chapter 4

Myceliophthora thermophila laccase gene can produce the highest reported laccase level in S. cerevisiae (Leontievsky et al., 1997). Insufficient enzyme stocks and the cost of redox mediators (RMs) are the most important obstacles for the commercial application of laccases. Thus it is necessary to make efforts to achieve cheap overproduction of these biocatalysts.

4.3.4 Biochemical properties of laccases Fungal laccases are generally glycoproteinic in nature, having approximately 50
100 kDa MW with an acidic isoelectric point around 4.0 (Baldrian, 2006). The presence of isoenzymes of laccase was demonstrated by p-phenylenediamine staining after isoelectric focusing in all tested wood rot fungi, namely Coprinus plicatilis, Fomes fomentarius, Heterobasidion annosum, Hypholoma fasciculare, Kuehneromyces mutabilis, Leptoporus litschaueri, Panus stipticus, Phellinus igniarius, Pleurotus corticatus, P. ostreatus, Polyporus brumalis, Stereum hirsutum, Trametes gibbosa, Trametes hirsuta, and T. versicolour typically with pI in the range of pH 3.0
5.0 (Baldrian, 2006). These isoenzymes may be produced by the same or different genes encoding for laccase (Archibald, Bourbonnais, Jurasek, Paice, & Reid, 1997). The number of isoenzymes produced by fungi depends on the type of species and also within species depending on their expression (induced or uninduced). They show wide variations in their stability, optimal pH, and temperature, as well as substrate specificities (Heinzkill et al., 1998). The catalytic performance of laccases at different pH and temperature conditions has been described. They also have different roles in the physiology of different fungi or within the same species under different conditions. The white-rot fungus P. ostreatus secretes eight different isoenzymes of laccase and of these six have been isolated and characterized extensively (Farnet, Criquet, Cigna, Gil, & Ferre´, 2004; Palmieri et al., 2003). The presence of copper regulates the production of laccase isoenzymes in P. ostreatus, and two dimeric isoenzymes have only been detected when copper is present (Farnet et al., 2004). Isoenzymes of laccase detected in the litter-decomposing fungus Marasmius quercophilus had different molecular weight and pI (Farnet, Tagger, & Le Petit, 1999). The detailed study on 17 different isolates of this fungus showed that the isoenzyme pattern [one of the three laccase bands on sodium dodecyl sulfate-polyacrylamide gel electrophoresis (SDS-PAGE)] was similar within different isolates after the induction of laccase with different aromatic compounds (Xu, 1996). Molecular weights of fungal laccases are generally in the range of 50
100 kDa (SDS-PAGE) and a typical laccase gene codes for approximately 500
600 amino acid chain. Laccases are glycoproteins and the carbohydrate content of the enzyme is determined by finding the difference between the molecular weight predicted on the basis of peptide sequence and the experimentally obtained molecular weight (SDS-PAGE). It typically

Fungal laccases: versatile green catalyst for bioremediation of organopollutants 95 accounts for about 10.0
45.0% of the total MW (Hong et al., 2002). One of the major components of carbohydrate linked to laccase is mannose. The covalent linkage of carbohydrate moieties to the protein is responsible for secretion, proteolytic susceptibility, activity, copper retention, and thermal stability of the enzyme (Madhavi & Lele, 2009). The isoelectric points (pI) of microbial laccases are generally in the range of 3.0
7.0. 4.3.4.1 Kinetic properties of laccases The Michaelis
Menten constant Km and the catalytic efficiency constant Kcat are generally used to describe the catalytic action of an enzyme. Km and Kcat constants have been calculated for a large number of laccases and vast differences can be observed among them (Table 4.2). The Km values of laccases generally depend on the source of enzyme and the reducing substrate (Table 4.2). Syringaldazine showed the lowest Km values; this is a dimer having two molecules of 2,6-dimethoxyphenol linked by an azide bond. For the affinity of syringaldazine to laccases, either the azide bridge or the dimer form is required because the Km values measured for monomeric 2,6-dimethoxyphenol is higher than those obtained with syringaldazine (Table 4.2). The comparison of Km values also shows that laccases produced by different organisms have different substrate specificities (Yaver et al., 1999). The Km values (20
50 μM) for oxygen have been reported for several laccases and it suggests that their specificity for the substrate is greater than for oxygen (Xu, 1997, 2001). More than 3500-fold variations have also been observed in the catalytic efficiencies (Kcat) of different laccases with the same substrate while the Kcat values for a single laccase with different substrates do not generally differ more than 2- to 10-fold. It suggests that Kcat describes the rate of the electron transfer reactions taking place within the enzyme after the binding of substrate (Xu, 1997). The constants presented in Table 4.2 have been measured under different physical and chemical conditions and all of them have a great impact on the results. Different molar extinction coefficients have been used in spectrophotometric assays of oxidized products because the nature of the actual oxidized products is generally complex. 4.3.4.2 Effect of pH and temperature on activity of laccases The pH optima for 3-ethylbenzthiazoline-6-sulfonate (ABTS) is more acidic and the range is 3.0
5.0 (Heinzkill et al., 1998). At high pH values, the difference in redox potential between the phenolic substrate and T1 copper could increase the oxidation of the substrate. The activity of fungal laccase decreases at neutral and alkaline pH because it is affected by the increasing hydroxide anion concentration. Binding of hydroxide anion to T2/T3 copper disrupts the transfer of internal electrons between the T1 and T2/T3 centers, which results in the inhibition of laccase activity. The stability of fungal laccases is usually maximum at pH 7.0 (neutral) but the activity of laccase is at its minimum at this pH. Decreasing the pH increases the redox potential of the phenolic substrate so it becomes more prone to

96

Chapter 4 Table 4.2: Kinetic parameters of laccases at optimized conditions.

Substrate

Km (µM) Kcat (min21) pH Laccase

References

Guaiacol

1200 3100 400

150 n.ra n.r

6 6 6

5120 510

115 n.r

3.4 Trametes trogii POXL3 4.5 Gaeumannomyces graminis

10,800 6800 2700 n.r 350,000 n.r 57,000 n.r 198 n.r n.r 1090 41,400 620 n.r 74,000 790 3000 n.r 23,000 28,000 3000 n.r 180

3 6.5 5.3 5.3 3 3 3 6 3.4 3 5 5.5 3 5.5 4 3.3 6 5.3 5.3 6 6 6 6 5.5

Trametes pubescens LAP2 Pleurotus sajor-caju Lac4 Trametes villosa Lcc1 Rhizoctonia solani Lcc4 Pleurotus ostreatus POXA1 Pleurotus ostreatus POXA2 Pleurotus ostreatus POXC Chaetomium thermophilum Trametes trogii POXL3 Panaeolus sphinctrinus Coprinus friesii Coprinus cinereus Lcc1 Trametes pubescens LAP2 Trichophyton rubrum Pycnoporus cinnabarinus Pleurotus sajor-caju Lac4 Myceliophthora thermophila Trametes villosa Lcc1 Rhizoctonia solani Lcc4 Pleurotus ostreatus POXC Pleurotus ostreatus POXA1 Pleurotus ostreatus POXA2 Chaetomium thermophilum Coprinus cinereus Lcc1

6 280 1.6 100

16,800 35,000 2100 n.r

4.5 6.5 6 3.5

Trametes pubescens LAP2 Pleurotus sajor-caju Lac4 Myceliophthora thermophila Botrytis cinerea

230 2100 740 410 96 26 72 120

430 21,000 n.r 109 n.r n.r 24,000 58,000

5 5 6.5 3.4 6 4.5 3 6

Pleurotus ostreatus POXC Pleurotus ostreatus POXA1 Pleurotus ostreatus POXA2 Trametes trogii POXL3 Chaetomium thermophilum Gaeumannomyces graminis Trametes pubescens LAP2 Pleurotus sajor-caju Lac4

Palmieri et al. (1997) Palmieri et al. (1997) Chefetz, Chen, and Hadar (1998) Garzillo et al. (1998) Edens, Goins, Dooley and Henson (1999) Galhaup et al. (2002) Soden et al. (2002) Xu (1996) Xu (1996) Palmieri et al. (1997) Palmieri et al. (1997) Palmieri et al. (1997) Chefetz et al. (1998) Garzillo et al. (1998) Heinzkill et al. (1998) Heinzkill et al. (1998) Schneider et al. (1999) Galhaup et al. (2002) Jung, Xu, and Li (2002) Record et al. (2002) Soden et al. (2002) Bulter et al. (2003) Xu (1996) Xu (1996) Palmieri et al. (1997) Palmieri et al. (1997) Palmieri et al. (1997) Chefetz et al. (1998) Marbach, Harel, and Mayer (1985) Galhaup et al. (2002) Soden et al. (2002) Bulter et al. (2003) Ahmad, Othman, Yusof, and Wahab (2011) Palmieri et al. (1997) Palmieri et al. (1997) Palmieri et al. (1997) Garzillo et al. (1998) Chefetz et al. (1998) Edens et al. (1999) Galhaup et al. (2002) Soden et al. (2002)

36 66 ABTS 58 52 90 120 280 190 30 32 41 23 14 45 55 2500 290 Syringaldazine 3.9 28 20 130 140 34 26

2,6-DMP

Pleurotus ostreatus POXC Pleurotus ostreatus POXA2 Chaetomium thermophilum

a n.r, Not reported; 2,6-DMP, 2,6-dimethoxyphenol. Source Data from: Viswanath, B., Rajesh, B., Janardhan, A., Kumar, A. P., & Narasimha, G. (2014). Fungal laccases and their applications in bioremediation. Enzyme Research, 2014, 163242.

Fungal laccases: versatile green catalyst for bioremediation of organopollutants 97 oxidation by laccase (Garzillo et al., 2001). For phenolic substrates, the pH activity profiles of fungal laccases are often bell-shaped and the optimum pH can range from 3.0 to 5.0 and up to 9.0 for plant laccases (Schneider et al., 1999). Fungal laccases show different temperature optima and these vary from strain to strain. The optimum temperature range of laccase obtained from G. lucidum is 20 C to 25 C and was stable between 10 C and 50 C for 4 h (Ko et al., 2001). Temperature stability of laccases differs significantly and it also depends on the source of the organism. Generally fungal laccases are stable at 30 C
50 C and rapidly lose activity at temperatures above 60 C (Palonen, Saloheimo, Viikari, & Kruus, 2003). 4.3.4.3 Effect of inhibitors on activity of laccases Several enzyme inhibitors show similar effects on laccase activity as other enzymes. Bollag and Leonowicz (1984) reported that azide, thioglycolic acid, and diethyldithiocarbamic acid inhibited laccase activity whereas it was affected to a lesser extent by ethylenediaminetetraacetic acid. Small anions like halides, azide, cyanide, and hydroxide can bind to T2/T3 copper and disrupt the internal electron transfer reaction and thus inhibit the activity. Laccases are also inhibited by metal ions (e.g., Fe21), fatty acids, Kojic acid, sulfahydril reagents, hydroxylglycine, and cationic detergents.

4.3.5 Mode of action of laccases Laccases (benzenediol:oxygen oxidoreductases) are copper-containing polyphenol oxidases (blue oxidases) which act on diphenols, polyphenols, substituted phenols, diamines, aromatic amines, benzene thiols, inorganic compounds like iodine, and allied substances having oxygen as an electron acceptor (Fig. 4.1). In the active holoenzyme form, they are

Figure 4.1 Laccase-mediated oxidation of phenolic subunits of lignin [The structure of laccase used was obtained from Coprinopsis cinerea (1A65)]. Source: Modified from Kunamneni, A., Ballesteros, A., Plou, F. J., & Alcalde, M. (2007). Fungal laccase—a versatile enzyme for biotechnological applications. Communicating Current Research and Educational Topics and Trends in Applied Microbiology, 1, 233
245.

98

Chapter 4

monomeric, dimeric, or tetrameric glycoproteins with four copper atoms per monomer. The configuration of these four copper atoms in the protein structure has been categorized in three groups on the basis of data obtained using UV
visible and electron paramagnetic resonance (EPR) spectroscopy (Palmieri et al., 2000). At resting state, the UV
visible spectra of laccase show two maxima (280 and 600 nm) and one shoulder near to 330 nm. The absorbance ratio at 330 nm to that at 600 nm is 0.5
2.0 while the absorption ratio at 280 nm to that at 600 nm is 14
30. At 610 nm, the type I copper (T1) gives an intense blue color to the enzymes, whereas type (T2) is colorless and both are detected by EPR, while the type III copper (T3) has a pair of copper atoms which give a weak absorbance near the UV spectrum (330 nm) rendered undetected by EPR (Bourbonnais & Paice, 1995). T2 and T3 copper sites are involved in the catalytic mechanism of the enzyme because they are close together and form a trinuclear center (Bourbonnais & Paice, 1995). A hydroxyl bridge between the two T3 copper atoms is maintained by strong antiferromagnetic coupling (Claus, 2003). Laccases catalyze four one e2 oxidation of a reducing substrate with concomitant two e2 reduction of dioxygen to water. The stoichiometry includes four molecules of a reducing substrate for each molecule of oxygen, involving a total transfer of four electrons (4RH 1 O2-4R 1 2H2O). The first step of catalysis is the reduction of the reducing substrate by copper (Cu21 to Cu1) at the T1 site (Fig. 4.2), which is the primary electron acceptor. Fig. 4.3 represents the detailed arrangements of four copper atoms and the ligand distance in laccase obtained from Coriolus zonatus. The electrons extracted from the reducing substrate are transferred to the T2/T3 trinuclear site, resulting in the conversion of the resting form (fully oxidized) of the enzyme to a totally reduced state.

Figure 4.2 Mechanism of laccase action. The Cu-T1 center oxidized the substrates and transfers the electrons to the T2 and T3 center by a highly conserved motif: His-Cys-His (HCH). In the scheme, copper atoms appear in brown and were based on the 1GYC structure (laccase-2 of Trametes versicolour). Source: Modified from Baldrian, P. (2006). Fungal laccases—occurrence and properties. FEMS Microbiology Reviews, 30(2), 215
242.

Fungal laccases: versatile green catalyst for bioremediation of organopollutants 99

Figure 4.3 ˚ ). Source: With slight Arrangements of copper atoms and ligand distance (blue, distances in A ´ ˜ modifications from Rivera-Hoyos, C. M., Morales-Alvarez, E. D., Poutou-Pinales, R. A., Pedroza-Rodrı´guez, A. M., RodrI´guez-Va´zquez, R., & Delgado-Boada, J. M. (2013). Fungal laccases. Fungal Biology Reviews, 27(3
4), 67
82.

Laccase can only act on the phenolic subunits of lignin which leads to Cα oxidation, CαCβ cleavage, and aryl-alkyl cleavage. Laccases can catalyze one-electron oxidation of a wide range of aromatic compounds (Thurston, 1994) like polyphenols (Archibald et al., 1997), methoxy-substituted monophenols, and aromatic amines (Bourbonnais, Paice, Reid, Lanthier, & Yaguchi, 1995) to the corresponding quinones. Phenol-oxidizing enzymes (laccases) preferably polymerize lignin by coupling the phenoxy radicals which are produced during oxidation of lignin phenolic groups (Campos, Kandelbauer, Robra, Cavaco-Paulo, & Gu¨bitz, 2001). The majority of the laccases have copper molecules in their active center and they are classified as blue copper oxidases but some laccases do not show these typical characteristics. Laccases isolated from solid-state fungal cultures were yellow-brown and did not show typical blue laccase spectrum. The comparison of N-terminal amino acid sequences of blue laccases (P. radiata, Panus tigrinus, C. versicolour, and Phlebia tremellosus) and yellowbrown laccase showed high homology. Yellow laccases showed an altered oxidation state of copper in the active center. The remarkable property of yellow laccase is that it oxidizes nonphenolic lignin models and veratryl alcohol in the presence of oxygen (Rogalski & Leonowicz, 1992).

100 Chapter 4 Cu+ H O 4R• + 4H+ T1

Cu++

Cu++

Cu+ O2

4RH

Cu+

Cu++

H O T3

Cu+

H O

Fully reduced enzyme

Cu++

Cu++

Cu++ O2– 2

T2

Cu++

Cu++

Resting (fully oxidized) enzyme

2H+

H O

H2O 2H+

Cu++

Cu++

H2O

Cu+ Peroxide intermediate

O Cu++ OXY radical intermediate

Figure 4.4 Catalytic cycle of laccase. Source: Adapted from Wong, D. W. (2009). Structure and action mechanism of ligninolytic enzymes. Applied Biochemistry and Biotechnology, 157(2), 174
209.

A successive four e2 oxidation (from four substrate molecules) is required for the total reduction of the enzyme (Fig. 4.4). Different mediators have been used to determine the redox potential of the T1 site of many laccases and it varies from 430 mV for plant laccase to 780 mV for fungal laccase (Klonowska et al., 2002; Koroleva, Gavrilova, Yavmetdinov, Shleev, & Stepanova, 2001; Reinhammar, 1972; Reinhammar & Va¨nnga˚d, 1971; Schneider et al., 1999; Xu, 1996, 1999; Xu et al., 2000). It has been found that there is a linear relationship between the catalytic efficiency (Kcat/Km) of laccases for some reducing substrates and the redox potential of the T1 copper, indicating that the higher the potential of the T1 site, the higher is the catalytic efficiency (Xu, 1996; Xu et al., 2000). One of the laccases from P. ostreatus does not show a blue color and has been described as white colored by the author (Palmieri et al., 1997). The result of atomic adsorption of this laccase showed that it has one copper atom, two iron atoms, and one zinc atom as compared to four coppers in a typical laccase.

4.3.6 Classification of laccases according to substrate specificity Laccase is a blue copper protein and is broadly classified as a polyphenol oxidase. The polyphenol oxidases are able to oxidize aromatic compounds using molecular oxygen as the final electron acceptor (Mayer, 1987). They generally carry out three types of activities:

Fungal laccases: versatile green catalyst for bioremediation of organopollutants

101

1. o-Diphenol: oxygen oxidoreductase or catechol oxidase (EC 1.10.3.1); 2. p-Diphenol: oxygen oxidoreductase or laccase (EC 1.10.3.2); and 3. Monophenol monooxygenase or cresolase (EC 1.18.14.1). These enzymes are differentiated on the basis of their ability to use different substrates (substrate specificity). It is difficult to define laccase according to its substrate specificity because it has a similar range of substrates as tyrosinase. Laccases have paradiphenol and orthoactivity but have higher affinity toward the first group. Syringaldazine is a specific substrate for laccase because only laccases can oxidize it (Madhavi & Lele, 2009). Different substrates induce laccase production at different levels so they are nonspecific to the inducing substrates and the number of substrates oxidized varies from one laccase to another (Wood, 1980).

4.3.7 Laccase mediator system The conditions which limit the laccase catalyzed oxidation of substrates are steric hindrance (for macromolecular compounds), very low affinity between the compound and the active site of enzyme, as well as high redox potential of the putative substrates. Due to lower redox potential, laccases can oxidize only phenolic compounds. This obstacle can be overcome by using the laccase mediator system (LMS) (Srebotnik & Hammel, 2000). Laccase mediators are usually low-molecular-weight substrates, water-soluble molecules, having high redox potential ( . 900 mV), that are capable of behaving as one-electron shuttles between the enzyme and the to-be-oxidized compounds (Bourbonnais & Paice, 1990; Kawai, Umezawa, & Higuchi, 1989) (Fig. 4.5). Laccase mediator compounds are generally laccase substrates. Laccase enzymes remove one electron from the substrate and produce free radicals which are long-lived enough to diffuse away from the enzyme active site and are capable of oxidizing other substrates present in the environment, thus restoring their stable electronic configurations at the expense of other substrates which are not directly oxidized by laccase (Call & Mu¨cke, 1997). RMs are generally present in nature and many others have been synthesized. The intermediate radicals arising due to the enzymatic oxidation of RMs can behave as conventional substrates and are slowly converted to the (relatively) stable quinonoid product(s), so in a large-scale application the selected RM has to be periodically replenished. Some natural mediators are normally present in decaying wood and some others are presumably produced by white-rot fungus during wood degradation. Phenolic compounds like p-hydroxybenzyl alcohol, phydroxybenzoic acid, vanillin, acetovanillone, syringaldehyde, acetosyringone, syringic acid, p-coumaric, ferulic, and sinapic acids (Fig. 4.6) have been studied and described in detail as laccase RMs (Johannes & Majcherczyk, 2000; Maruyama, Komatsu, Michizoe, Sakai, & Goto, 2007).

102 Chapter 4

Figure 4.5 Mechanism of laccase action in the presence and absence of laccase-mediators.

Figure 4.6 Chemical structures of some natural laccase mediators.

Fungal laccases: versatile green catalyst for bioremediation of organopollutants

103

Phenolics with two o,o0 -methoxyls are good laccase substrates and they cannot be dimerized or polymerized upon oxidation by laccase because their quinonization is comparatively difficult. Therefore phenolic-type compounds are the most efficient RMs (Murugesan, Yang, Kim, Jeon, & Chang, 2009). Many artificial compounds have been assessed for their potential applications as RMs and most of them have shown good efficiency. The first synthetic molecule which has been shown to have a role as an RM is ABTS (Bourbonnais & Paice, 1990). Phenol Red (a sulfophthalein), which is a synthetic phenolic compound, has been proved to be effective as an RM (D’Acunzo, Galli, Gentili, & Sergi, 2006). For industrial applications, hydroxylamine derivatives ( . N
OH moiety) are more popular as RMs for laccase because they are readily available at reasonable prices. The main drawbacks of hydroxylamine derivatives ( . N
OH moiety) are their well-known toxicity and their tendency to act as laccase inhibitors (Can˜as & Camarero, 2010). Fig. 4.7 represents the most used N
OH type RMs. (2,2,6,6-tetramethylpiperidin-1-yl)oxidanyl and their analogues have a stable . N
O∙ radical, which is oxidized to the corresponding

Figure 4.7 Chemical structure of synthetic laccase mediators: (A) Hydroxylamine derivatives, (B) Phenothiazine and Pyrazolone-type derivatives and (C) Nitroso compounds.

104 Chapter 4

Figure 4.8 Mechanism of laccase redox mediators [Electron atom transfer (ET) or Hydrogen atom transfer (HAT)].

oxoimmonium cation by laccase (Kubala et al., 2013; Shiraishi, Sannami, Kamitakahara, & Takano, 2013). These RMs share a similar mechanism with the phenolic ones (Fig. 4.8) like hydrogen atom transfer (HAT) (Astolfi et al., 2005), while ABTS typically acts through an electron transfer (ET) mechanism (Cantarella, Galli, & Gentili, 2003) (Fig. 4.9). When we use a proper substrate, two different mechanisms produce different products. When two RMs with different mechanisms (HAT or ET) are used together, synergistic effects have been observed (Jeon, Murugesan, Kim, Kim, & Chang, 2008; Pickard, Roman, Tinoco, & Vazquez-Duhalt, 1999). LMS have been applied to numerous processes, such as pulp delignification, oxidation of organic compounds, and the development of biosensors (Khambhaty, Ananth, Sreeram, Rao, & Nair, 2015). LMS has been studied extensively and there are still unsolved problems concerned with mediator recycling, cost, and toxicity.

4.3.8 Immobilization of laccase Compared to conventional chemical catalysts, enzymatic reactions are more advantageous but have some practical problems, such as operational life time, high cost of isolation and purification, instability of their structure, and requirement of specific conditions for

Fungal laccases: versatile green catalyst for bioremediation of organopollutants

105

Figure 4.9 Production of cationic radicals of ABTS by laccase using electron transfer mechanism.

catalytic action. Immobilization of the enzyme may overcome many of these problems. The storage and operational stability of biocatalysts are frequently enhanced by the immobilization process. Moreover, the recycling of an immobilized catalyst represents an immense advantage compared to a free enzyme. Laccases have broad substrate specificity and the application of a mediator system further widens the specificity toward nonphenolic and xenobiotic pollutants. These properties make laccase an ideal biocatalyst for industrial processes, including bioremediation, biobleaching of paper pulp, chemical synthesis, textile finishing, biosensing, and wine stabilization. Different methods have been used for laccase immobilization, such as adsorption, entrapment, encapsulation, covalent binding, and self-immobilization. Recently, immobilized laccase have gained special attention for exploitation in environmental and electro-biochemistry. de Souza Bezerra, Bassan, de Oliveira Santos, Ferraz, and Monti (2015) have provided guidelines for the selection of proper solid supports and immobilization conditions. The application of covalent binding for immobilization procedures has improved the temperature and pH stability of laccase (Mazlan & Hanifah, 2014). Dura´n, Rosa, D’Annibale, and Gianfreda (2002) comprehensively reviewed the variety of supports available for immobilization, techniques of immobilization, and potential industrial applications of different immobilized laccases. To immobilize laccase, several approaches have been reported, such as the absorption method (Hou, Dong, Ye, & Chen, 2014), entrapment into polymer support (Jaiswal, Pandey, & Dwivedi, 2016; Sampaio et al., 2016), or covalent linkages (de Souza Bezerra et al., 2015; Misra, Kumar, Goel, & Varshney, 2014). Among all, covalent attachment is the most frequently used technique because it provides strong and stable enzyme attachment as well as in some cases reducing enzyme deactivation rates.

106 Chapter 4 It has been reported that the application of covalent linkage for immobilization usually increases the stability of the enzyme and prevents its leaching from the reaction system (Li et al., 2013). After immobilization, the structure, functions, and biological activities of enzyme must be maintained. Mechanical properties of the polymeric support (surface area, porosity, and functional group density) can easily be customized according to the specific requirements.

4.4 Application of laccases for bioremediation of environmental pollutants One of the major environmental problems faced by the modern world is the pollution of soil, water, and air by toxic persistent chemicals. The extensive use of pesticides in agriculture can lead to significant environmental pollution. The United States produce 80 billion pounds of hazardous organopollutants annually and only 10.0% of these are disposed of safely (Alexander, 1994). The need for the development of efficient and green oxidation technologies has increased the attention of researchers on the use of enzymes as a substitute of conventional nonbiological methods (Upadhyay et al., 2016). Fungal laccases are extracellular multicopper oxidases that use molecular oxygen to oxidize a wide variety of organic compounds, ranging from phenolic, nonphenolic ligninrelated compounds to highly recalcitrant environmental pollutants, by a radical-catalyzed reaction mechanism (Majeau, Brar, & Tyagi, 2010). Laccases can degrade lignin as well as decolorize and detoxify the industrial textile effluents, thus facilitating the bioremediation of wastewater. Besides this, the substrate range of laccase can be extended by the addition of a mediator system. Fungal laccases have the ability to oxidize a wide range of substrates and hence they are potentially important for industrial applications (Fig. 4.10). The application of a LMS to bleach indigo dye from denim can generate a desired worn appearance (Onuki, Nogucji, & Mitamura, 2000). Laccases are able to oxidize toxic organic pollutants, such as PAHs (Reddy & Mathew, 2001) and chlorophenols (Marbach et al., 1985), so they are ideal catalysts for bioremediation of environmental pollution. The most useful method for the bioremediation of contaminated soil is inoculating the soil with fungi that are efficient laccase producers because the use of isolated enzymes is not economically feasible for soil remediation at a large scale. Over the last two decades, laccases from white-rot Basidiomycetes have been investigated extensively for their enormous potential applications, viz., bioremediation in the pulp and paper, textile, dye, and food industries. Many researchers have introduced laccase as a green catalyst for potential applications in various biotechnological processes including the bioremediation of soil, water, and the development of environment-friendly processes (Table 4.3), proving it to be a multifaceted enzyme.

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107

Figure 4.10 Application of laccase in biotechnology. Source: From Gonc¸alves, I., Silva, C., & Cavaco-Paulo, A. (2015). Ultrasound enhanced laccase applications. Green Chemistry, 17(3), 1362
1374.

4.4.1 Degradation of xenobiotic compounds PAHs comprise various prime organic pollutants which are toxic, mutagenic, and/or carcinogenic in nature. Besides this, they are ubiquitous, refractory, and unmanageable. Owing to these properties, PAHs have created an alarming situation as far as public health and environmental health is concerned (Ghosal, Ghosh, Dutta, & Ahn, 2016). PAHs, pentachlorophenols (PCP), polychlorinated biphenyls, 1,1,1, trichloro 2,2, bis (4-chlorophenyl) ethane, benzene, toluene, ethylbenzene, and xylene, as well as trinitrotoluene (TNT), are persistent in the environment and are known to have carcinogenic and/or mutagenic effects. WRF are known to be ideal bioremediation candidates because they have the ability to detoxify/transform a wide variety of toxic hazardous chemicals (Bollag, Chu, Rao, & Gianfreda, 2003). An alternative method for the removal of toxic xenobiotics from the environment is the application of an enzymatic process (Marbach et al., 1985). During recent years, fungi have been studied comprehensively for the biodegradation of PAHs and reports of various fungal species degrading different PAHs are now available (Cerniglia, 1993; Cerniglia & Sutherland, 2010). Laccases show broad substrate specificity so they are able to oxidize a wide range of xenobiotic compounds, such as chlorinated phenolics (Torres, Bustos-Jaimes, & Le Borgne, 2003), pesticides (Pozdnyakova et al.,

108 Chapter 4 Table 4.3: Applications of fungal laccases for bioremediation of organopollutants. S. no. Source of laccase

Applications

References

1. 2.

Aspergillus flavus Coriolus versicolour

Decolorization of malachite green dye Degradation of textile dyes

3.

Streptomyces cyaneus Decolorization and detoxification of azo dyes

4.

Phanerochaete chrysosporium

Ali, Ahmad, and Haq (2009) Sanghi, Dixit, Verma, and Puri (2009) Moya, Herna´ndez, Garcı´aMartı´n, Ball, and Arias (2010) Heinfling, Bergbauer, and Szewzyk (1997)

5.

7.

Schizophyllum commune Pycnoporus cinnabarinus Coriolopsis gallica

8.

Pichia pastoris

9.

Pleurotus ostreatus

10.

Pleurotus eryngii

11.

Trametes pubescens

12.

F. trogii ATCC 200800 T. versicolour ATCC 200801

6.

13.

14.

L. polychrous

15. 16.

Coriolopsis gallica Trametes sp. SYBC-L4 Peniophora sp. (NFCCI-2131)

17.

Decolorization of commercially used reactive textile dyes; Reactive Orange-96, Reactive Violet-5R, Reactive Black-5 (RB-5), and Reactive Blue-38 Decolorization of wastewater released from a bagasse-pulping plant Decolorization of pigmented plant effluent Oxidation of recalcitrant polycyclic heterocycles compounds carbozole, Nethylcarbozole, fluorine, and dibenzothiophene present in coal tar and crude oil in the presence of 1hydroxybenzotriazole (HBT) and ABTS as free radical mediators Engineered to improve the efficiency of particular bioremediation processes Degradation of polycyclic aromatic hydrocarbons (PAHs) in the presence of a synthetic mediator Lignin and organopollutant degradation, as well as to improve the bioremediation potential Bioremediation of a mixture of pentachlorophenol (PCP), 2-chlorophenol (2CP), 2,4-dichlorophenol (2,4-DCP), and 2,4,6-trichlorophenol (2,4,6-TCP) Decolorization of Reactive Blue 171 and indigo carmine Decolorization of Reactive Red 198, Rem Blue RR, Dylon Navy 17, Rem Red RR, and Rem Yellow RR Decolorization of RB-5, Reactive Orange 16, Reactive Green 19, Methyl Orange, Acid Blue 80, and Water Blue Beer factory wastewater treatment Decolorization of Congo red, aniline blue, and indigo carmine Decolorization of amido black, crystal violet, brilliant green, methyl orange, and methylene blue

Belsare and Prasad (1988) Schliephake, Lonergan, Jones, and Mainwaring (1993) Dec and Bollag (2000)

Dhawan and Kuhad (2002) Pozdnyakova, Turkovskaya, Yudina, and RodakiewiczNowak (2006) ´mez-Toribio, Garcı´a-Martı´n, Go Martı´nez, Martı´nez, and Guille´n (2009) Gaitan et al. (2011)

Birhanli, Erdogan, Yesilada, and Onal (2013) Sa¸smaz et al. (2011) ¸

Ratanapongleka and Phetsom (2014) Madhavi and Lele (2009) Li, Zhang, Tang, Zhang, and Mao (2014) Shankar and Nill (2015)

(Continued)

Fungal laccases: versatile green catalyst for bioremediation of organopollutants

109

Table 4.3: (Continued) S. no. Source of laccase

Applications

References

18.

Dechlorination and decolorization of pulp and paper effluent Decolorization of simulated effluent containing Acid Red 27 or Basic Green 4 or Acid Violet 17 Decolorization of RBBR

Eaton (1980)

19.

Phaenerochaaete chrysosporium C. bulleri

20.

P. acaciicola LA 1

21

Ganoderma sp. En3

22

T. versicolour ATCC 200801 T. versicolour IBL-04 Reactive Violet 1, Reactive Blue 21, Reactive Yellow 145A, RB-5, Reactive Red 195A Coriolopsis caperata RBBR DN

23 24

Decolorization of RBBR, indigo carmine, methyl green Decolorization of Reactive Red 3

Chhabra, Mishra, and Sreekrishnan (2015) Adak, Tiwari, Singh, Sharma, and Nain (2016) Lu et al. (2016) Ilk, Demircan, Sa˘glam, Sa˘glam, and Rzayev (2016) Asgher, Noreen, and Bilal (2017) Patel et al. (2017)

RBBR, Remazol Brilliant Blue R. Source: Upadhyay, P., Shrivastava, R., & Agrawal, P. K. (2016). Bioprospecting and biotechnological applications of fungal laccase. 3 Biotech, 6(1), 15; Yesilada, O., Birhanli, E., & Geckil, H. (2018). Bioremediation and decolorization of textile dyes by white rot fungi and laccase enzymes. In: Mycoremediation and environmental sustainability (pp. 121
153). Cham: Springer.

2019), and PAHs (Pointing, 2001) (Table 4.3). Moreover, laccases can also degrade PAHs which arise from natural oil deposits and the utilization of fossil fuels (Bressler, Fedorak, & Pickard, 2000). Purified laccase from Coriolopsis gallica oxidizes carbozole, N-ethylcarbozole, fluorine, and dibenzothiophene in the presence of 1-hydroxybenzotriazole (HBT) and ABTS as laccase mediators (Dec & Bollag, 2000). Laboratory experiments have demonstrated that the application of laccase can remove phenols and aromatic amines from water (Niku-Paavola & Viikari, 2000). The basic mechanism of removal of pollutants from the environment is the enzymatic oxidation of pollutants to free radicals or quinones that undergo further polymerization and partial precipitation (Niku-Paavola & Viikari, 2000). Laccase produced by white-rot fungus T. hirsuta has been used for the oxidation of alkenes (Bollag & Myers, 1992). The oxidation involves a two-step process in which the enzyme first catalyzes the oxidation of a laccase mediator added to the reaction and then the oxidized mediator further oxidizes the alkene to the corresponding ketone or aldehyde. Laccase can also immobilize xenobiotic pollutants (phenolic compounds including chlorinated phenols and anilines such as 3, 4 dichloroaniline and 2,4,6 trinitrotoluene) on humic substances present in the soil by a coupling reaction, a process analogous to humic acid synthesis in soils (Ahn, Dec, Kim, & Bollag, 2002). Immobilization lowers the toxicity of xenobiotics to biological systems. P. pastoris was engineered by site-directed mutagenesis which improves the rate of electron transfer between the active site of laccase and an electrode (Riu, Scho¨nsee, & Barcelo, 1998). Thus laccases can be engineered to improve the efficiency of particular bioremediation processes.

110 Chapter 4

4.4.2 Decolorization of synthetic dyes WRF produce efficient oxidoreductases that degrade lignin and synthetic dyes in aerobic conditions (Ali, 2010; Khan, Bhawana, & Fulekar, 2013). WRF secrete enzyme in nutrientlimiting conditions, rather than in the presence of the pollutant. Wet processing of textiles consumes large volumes of dyestuff, chemicals, and water. The textile industry comprises two-thirds of the total dyestuff market (Banat, Nigam, Singh, & Marchant, 1996). The chemical composition of reagents used in different processes is very diverse, ranging from inorganic compounds to polymers and organic products (Zollinger, 2002). More than 100,000 structurally different dyes are available commercially and over 7 3 105 tonnes of dyestuff are produced annually in the world (Upadhyay et al., 2016). Synthetic textile dyes have complex structures so they are resistant to fading on exposure to light, water, and different chemicals and most of them are difficult to decolorize. The governments of developed countries have imposed strict legislation for the removal of dyes from the industrial effluents (McKay & Sweeney, 1980). Several dyes are synthesized from known carcinogens, such as benzidine and other aromatic compounds, which is a major health concern (Hou et al., 2004). The existing processes used to treat dye wastewater are inefficient and expensive; therefore the development of processes based on laccases seems to be an attractive solution because of their potential to degrade dyes of diverse chemical structure (Couto, 2007) including synthetic dyes currently employed in the industry (Setti, Giuliani, Spinozzi, & Pifferi, 1999). The application of laccase in the textile industry is growing rapidly since, besides decolorizing textile effluents, laccases are also used to bleach textiles and to synthesize dyes (Raghukumar, 2000). In a low nitrogen medium, Flavodon flavus decolorized several synthetic dyes, such as Azure B and Brilliant Blue R (Soares, de Amorim, & Costa-Ferreira, 2001). P. sanguineus produced laccase as the sole phenoloxidase in the culture medium and decolorized two azo dyes partially and two triphenylmethane dyes (bromophenol blue and malachite green) completely (Rodriguez, Pickard, & Vazquez-Duhalt, 1999; Yaver et al., 1996). Purified laccase of T. hirsuta was able to degrade azo, indigoid, athraquinonic, and triarylmethane dyes used in dyeing textiles (Enayatizamir et al., 2011) as well as 23 other industrial dyes (Ying, Williams, & Kookana, 2002). It was reported that the most efficient strain for decolorization of dyes is T. trogii and the mediator is 1-HBT. The rate of decolorization of dyes was increased in the presence of mediators. Decolorization of dyes and the efficiency of mediators were significantly affected by structural differences among the dyes. The redox potentials of laccases depend largely on their source. HBT could destabilize the laccase so the application of a culture filtrate instead of purified laccase could reduce the cost of decolorization. Murugesan, Dhamija, Nam, Kim, and Chang (2007) reported that laccase from G. lucidum was able to decolorize 40.7% Malachite Green (25 mg/L) after 24 h of incubation and the

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111

addition of phenolic extract of wheat bran improved the decolorization of malachite green significantly by two- to threefold more than purified laccase. Murugesan et al. (2007) used response surface methodology (RSM) for the decolorization of the azo dye Reactive Black-5 (RB-5) using laccase from P. sajor-caju and observed that HBT was required for RB-5 decolorization. He obtained 84.4% RB-5 decolorization using the concentrations of dye (62.5 mg/L), enzyme (2.5 U mL), and HBT (1.5 mM) after 36 h. Synthetic dye decolorization by the statistically optimized laccase mediator system showed that the rate of dye ¸ smaz decolorization was significantly improved and also widened the substrate specificity. Sa¸ et al. (2011) used crude laccase obtained from the submerged culture of Trametes versicolour ATCC 200801for the decolorization of Reactive Red 198, Rem Blue RR, Dylon Navy 17, Rem Red RR, and Rem Yellow RR dyes and reported that the types of mediators and dye structures have significant roles in decolorization. Zeng et al. (2011) reported that the crude laccase of T. trogii decolorized anthraquinone dyes [Reactive Blue 4 (70.0%), RBRR (85.0%), and Acid Blue 129 (46.0%)] without any mediator while HBT was required for the decolorization of azo dyes acid red 1 (90.0%) and RB-5 (65.0%). The concentration of laccase, pH, and temperature played essential roles for dye decolorization. The decolorization potential of this crude culture without a mediator may be more advantageous because of the ¨ zmen, 2014). high expense and toxicity of the mediators (Ye¸silada, Birhanli, Ercan, & O Benzina et al. (2012) reported that 150 min was required to decolorize azo dye Sirius Rose BB (99.5%). Mirzadeh et al. (2014) reported 50.0% decolorization of triazo dye (Direct Blue-71) using laccase, whereas 86% decolorization occurs in the presence of HBT (5.0 mM). Forootanfar et al. (2012) reported that laccase from P. variabile decolorized Bromophenol Blue (100.0%), Coomassie Brilliant Blue (91.0%), RBBR (47.0%), and Congo red (18.5%) in the presence of HBT (5.0 mM) after a 3 h incubation period. Ashrafi, Rezaei, Forootanfar, Mahvi, and Faramarzi (2013) applied purified laccase of Paraconiothyrium variabile for the decolorization of 13 dyes and reported that most dyes required 180 min to achieve maximum decolorization. He also reported that the use of HBT as a mediator increased the decolorization rate of Reactive Orange 16, Reactive Black-39, Direct Blue-71, Disperse Red-177, as well as Acid Yellow-36. Yang et al. (2015) successfully applied a quadratic model to decolorize malachite green with LacA and obtained maximum decolorization (91.6%) using 2.8 U/mL LacA, 109.9 mg/L dye after 172.4 min. Adak et al. (2016) reported that the laccase produced by Pseudolagarobasidium acaciicola LA 1 under solid-state fermentation on parthenium biomass was a thermostable enzyme and it functioned optimally at pH 4.5 and a temperature of 60 C. It decolorized RBBR (90.0%) and RB-5 (33.0%) without a mediator after 4 and 48 h, respectively. Lu et al. (2016) reported that crude laccase enzyme from liquid cultures of Ganoderma sp. En3 efficiently decolorized RBBR, indigo carmine, and methyl green. Studies showed that the decolorization efficiency significantly increased in the presence of RMs (ABTS, syringaldehyde, acetosyringone, and acetovanillone). Sayahi, Ladhari, Mechichi, and Sakli (2016) used

112 Chapter 4 purified laccase from T. trogii to decolorize RB-5, Reactive Violet 5 (RV5) and also the mixture of RB-5 and RV5 and RBBR. The highest decolorization of 25 mg/L dye RB-5 (93.0%) was achieved at 1.0 U mL2 1 laccase and 1.0 mM HBT concentration while the maximum decolorization of 25 mg/L dye RV5 (100%) was obtained at 0.5 U/mL enzyme and 0.5 mM HBT concentration. RBBR also acted as a mediator and increased the decolorization of these two dyes. The purified laccase also decolorized mixed dyes (RBBR, RB-5, and RV5) 55.0% without HBT after 24 h. Sayahi et al. (2016) optimized RB-5 dye decolorization by RSM and reported that the optimum concentrations of dye, enzyme, and HBT were 25.0 mg/L, 1.0 U/mL, and 1.0 mM, respectively. After statistical optimization, the rate of RBBR decolorization by C. caperata DN laccase at 1.0 U/mL is 542.0 mg/L (1000 mg/L) within 1 h without a mediator (Patel et al., 2017). Chhabra et al. (2015) immobilized laccase obtained from Cyathus bulleri in poly(vinyl alcohol)-boric acid or polyvinyl alcohol-nitrate beads and it was used for the decolorization of simulated effluent containing acid red 27 (acidic monoazo dye) or basic green 4 (malachite green) dyes. The simulated effluent containing 100 μM acid violet 17 was decolorized 90.0% in the presence of 100 μM ABTS by laccase entrapped in PVA-nitrate up to 10 cycles. It also decolorized a simulated effluent containing 100 μM Basic Green 4 (95.0%) in the presence of 100 μM ABTS up to 20 cycles under batch mode. The laccase of Trametes versicolor ATCC 200801 was immobilized by Ilk et al. (2016) on the poly(MAalt-MVE)-g-PLA/ODA-MMT nanocomposite by adsorption or covalent coupling and used for the decolorization of Reactive Red 3 (monoazo dye). They reported that under the optimum conditions, the dye decolorization potential of the immobilized laccase (65.0%) was much higher than the free laccase (33.0%) and even after 10 cycles, the activity of the immobilized laccase was retained by 77.0%. The purified laccase enzyme of Trametes versicolor IBL-04 immobilized onto chitosan microspheres showed higher catalytic efficiency and higher thermal as well as storage stability (Asgher et al., 2017). The immobilized enzyme decolorized Reactive Red 195A (100.0%), Reactive Violet 1 (99.0%), Reactive Yellow 145A (98.0%), RB-5 (97.0%), and Reactive Blue 21 (89.0%) after 4 h incubation in the presence of 1 mM ABTS as a RM and retained its 80.0% activity after 10 cycles (Asgher et al., 2017) (Table 4.3). Oxidation of the substrate by laccase alone and in the presence of a mediator results in different end products. Based on changes in the absorbance maximum, Papinutti and Forchiassin (2004) suggested that the modification of malachite green by laccase of Fomes sclerodermeus was different in the presence of a mediator (HBT) and similar results were reported by Chhabra, Mishra, and Sreekrishnan (2009). In the presence of laccase alone, the transformation of malachite green was carried out by stepwise N-demethylation while in the presence of a mediator it was hydroxylated and broken down. Yang et al. (2015) reported the presence of identical intermediates while degradation of malachite green was carried out by LacA or a LacA/ABTS system. It was also observed that after 24 h precipitates were

Fungal laccases: versatile green catalyst for bioremediation of organopollutants

113

present at the bottom of the tube suggesting polymerization of the end products. Yang et al. (2015) proposed the presence of both types of pathways responsible for rapid biocatalysts of malachite green and also revealed that LacA was a high-redox-potential laccase due to the presence of three amino acids (Ser-113, Glu-456, and Phe-459) at active sites, which is conserved among fungal laccases with high redox potentials (Eggert, La Fayette, Temp, Eriksson, & Dean, 1998; Piontek, Antorini, & Choinowski, 2002).

4.4.3 Treatment of industrial effluent Fungal laccases offer several advantages for biotechnological applications, for example, treatment of industrial effluent. Laccases exhibit broad substrate specificity, hence they can be used for bleaching of kraft pulp or to detoxify agricultural by-products including olive mill wastes and coffee pulp (Heemken et al., 2001). Laccase from F. flavus was shown to decolorize the kraft paper mill effluent produced by a bleach plant (Lyons, Newell, Buchan, & Moran, 2003). Laccase obtained from white-rot fungus T. villosa degrades bisphenol, which is an endocrine-disrupting chemical (Raghukumar, 2000). Nonylphenols have the potential to mimic the action of natural hormones in vertebrates (Lyons et al., 2003), so they have increasingly gained attention. Nonylphenols are produced during incomplete degradation of nonylphenol polyethoxylates (NPEOs), which have been widely used as nonionic surfactants in industrial processes. Both nonylphenols and NPEOs have entered into the environment because of incomplete removal from wastewater treatment facilities (Lyons et al., 2003). Nonylphenols are more resistant to biodegradation than NPEOs compounds so they are detected worldwide in wastewater treatment plant effluents and rivers (Bermek, Li, & Eriksson, 2002). Laccase produced by Clavariopsis aquatic is able to degrade xenoestrogen nonylphenol (Calvo, Copa-Patin˜o, Alonso, & Gonza´lez, 1998). Due to their potential role in such degradation processes, for example, for natural attenuation processes in freshwater environments, the laccases also offer new prospectives for environmental biotechnological applications such as wastewater treatment.

4.4.4 Potential applications in pulp and paper industry The fibers present in wood are glued together by lignin in a piece of wood. These glued fibers can be separated either by chemical or by mechanical pulping. Production of paper from wood requires the separation of wood fibers from each other and then their processing into sheets (Madhavi & Lele, 2009). The tensile strength of pulp can be increased by pretreatment of wood chips with ligninolytic fungi while the requirement of energy for mechanical pulping is also decreased. Laccase produced by C. albidus effectively reduced the lignin content of eucalyptus wood and can be used for biopulping in the pulp and paper industry (Singhal, Kumar, & Rai, 2005). More than 80.0% tensile strength can be increased by pretreatment of hardwood with Phlebia tremellosa. Four weeks pretreatment of Phlebia

114 Chapter 4 brevispora to aspen chips reduced the energy requirement by 47.0%. Current pulp bleaching processes generally use chlorine-based chemicals which results in the formation of chlorinated aliphatic and aromatic compounds that could be acutely toxic, mutagenic, and carcinogenic (Ta¸spınar & Kolankaya, 1998). In recent years, novel enzyme-based bleaching technologies that are environmentally benign have been intensively studied (Ta¸spınar & Kolankaya, 1998). The application of LMS as bioremediation agent has shown immense potential for the paper and pulp industry but the cost of the mediator and alkalinity of the effluent hindered its use at a large scale (Ta¸spınar & Kolankaya, 1998). Thus several researchers have spent considerable effort in isolating laccases that could be suitable for this type of bioremediation. The laccase of C. gallica has been used for the decolorization of alkaline effluents produced by pulp and paper industry (Milstein et al., 1988). Laccases have also been used for the direct dechlorination of wastewater generated by the pulp and paper industry (Bajpai, 1999) and for the removal of chorophenols and chlorolignins present in bleach effluents (Wong, Richardson, & Mansfield, 2000) or effluents from other industries (Ghindilis, Gavrilova, & Yaropolov, 1992). Laccases convert phenolic compounds into less toxic compounds by degradation or by polymerization reactions and/or cross-coupling of toxic phenol with naturally occurring phenols (Yaver et al., 1999). Laccases can also be used for the reduction of the kappa number of the pulp (Abadulla, Robra, Gu¨bitz, Silva, & Cavaco-Paulo, 2000) and the improvement in the papermaking properties of pulp (Kuznetsov, Shumakovich, Koroleva, & Yaropolov, 2001).

4.4.5 Applications of laccases to develop ecofriendly processes Over the last decade, laccases have been the subject of several review articles. Laccases have gained much attention because they catalyze a wide range of reactions, such as the delignification of pulp and paper, detoxification and decontamination of effluents, decolorization/degradation of dyes, degradation of PAHs, and as bioreceptors in biosensor application (Bayramo˘glu, Kaya, & Arıca, 2005; Mazlan & Hanifah, 2017). Laccases have become an ideal candidate for applications in green chemistry due to their ability to form a bond between reactants under mild and ecofriendly reaction conditions (Li, Xu, & Eriksson, 1999). Laccases also have a broad substrate range which is of interest to organic chemists as it allows their exploitation in organic syntheses, such as oxidative decoupling, formation of quinine, cross-coupling reactions, deprotection reaction, decarboxylation reaction, dimerization, polymerization, and synthesis of dye and pharmaceutical compounds (Fig. 4.11). Laccases can be applied for the removal of harmful sulfur-containing compounds which are generally emitted during the use of fossil fuel and they have been used for biosolubilization of coal to generate low-grade ores and liquid fuels (Arora & Sharma, 2010). Laccases have the potential to transform a wide range of chemical compounds under environment-friendly conditions, which is not

Fungal laccases: versatile green catalyst for bioremediation of organopollutants

O2

H2O

115

Substance

Amination Oxidation Polymerization Substance(OX) Dimerization

Product

Cyclization

Figure 4.11 Potential applications of laccases in organic synthesis. Source: Obtained from Mogharabi, M., & Faramarzi, M. A. (2014). Laccase and laccase-mediated systems in the synthesis of organic compounds. Advanced Synthesis and Catalysis, 356(5), 897
927.

possible by classical organic synthesis. It is anticipated that the increasing demand of green chemical processes will lead to the search for novel applications of laccases, which will further improve and modernize the current chemical, pharmaceutical, and industrial processes (Wellington, 2012).

4.5 Limitations and future prospects Laccases are one of the most important enzymes being used in the bioremediation of xenobiotics and pollutants being produced due to the large-scale use of chemicals in various industrial processes. Osma, Toca-Herrera, and Rodrı´guez-Couto (2011) produced laccase from the white-rot fungus Trametes pubescens and reported that the expenditure associated with the production of laccase is low but the cost of the purification of the laccase is very high. The overall cost of laccase production is increased due to the process of purification and it is the major factor that hinders its commercialization. It is necessary to develop better and more cost-effective methods for the large-scale production of laccases which would enable the development of “greener” approaches for a “clean” environment. Various limitations are associated with laccase and its utilization in large-scale industrial processes. Thus research now has to be diverted toward the discovery of novel enzymes which can work and tolerate alkaline pH, remain active in different operating condition, and have enhanced life spans. Enzymes are generally costly and are more or less quickly inactivated, so immobilization of the enzyme could help to solve these drawbacks and allow enzyme and/or mediator recovery and recycling. The activity and stability of the enzyme can be improved by protein engineering which will help to obtain a more robust and active enzyme and thus enable commercialization. It is also necessary to focus on the

116 Chapter 4 thermodynamics and physiological properties of enzymes which also affect catalytic performance (Agrawal, Chaturvedi, & Verma, 2018). Heterologous production of recombinant laccases is another novel strategy which may improve the efficiency of the laccase-based treatments of textile dyes (especially in the perspective of decreasing the economic impact of the whole process) but to obtain a fully functional enzyme the presence of posttranslation modifications (particularly, glycosylation) should be given due care. It is also necessary to search for new strains that are capable of producing industrially important laccase to be applied on a commercial scale.

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108.

CHAPTER 5

Emerging bioremediation technologies for the treatment of wastewater containing synthetic organic compounds Kunal Jain1, Jenny Johnson1, Neelam Devpura1, Rohit Rathour2, Chirayu Desai2, Onkar Tiwari3 and Datta Madamwar1 1

Environmental Genomics and Proteomics Lab, UGC-Centre of Advanced Study, Post Graduate Department of Biosciences, Sardar Patel University, Satellite Campus, Bakrol, India, 2PD Patel Institute of Applied Science, Charotar Institute of Science and Technology, Changa, India, 3 Department of Biotechnology, Ministry of Science & Technology, New Delhi, India

5.1 Introduction Ever since the evolution of human beings, tools and machinery have coevolved with them. The coevolution of human life and technology development has provided much required technical edge for their survival and sustenance during their early life on the Earth. Initial inventions, which might have been accidental discoveries, paved the way for modern industrial revolutions. With the rise in human populations the need for technologically driven life also increased proportionately and we did not realize when our environment started to lose its pristine nature. The modern industrial revolution marked a significant point in human history, since when the ease of human life on Earth has exponentially increased. The change also brought changes in the biosphere, for which humans were not prepared and had not anticipated where it would rapidly create situations where the basic entity of life is in danger. Human activities have damaged and polluted the natural terrestrial, aerial, and aquatic ecosystems at the same pace and with same intensity. In recent years environmental pollution has become a priority problem globally. It has changed drastically and been compounded with the daily transformations in the habits, lifestyles, and ever-rising living standards of humans. The major impact on the pristine environment was made by synthetic organic compounds (SOCs). The SOCs are often an intimation of natural organic compounds, that historically were derived from naturally occurring materials (petroleum, natural gas or coal). From the latter part of the 20th century natural organic compounds Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00005-5 © 2020 Elsevier Inc. All rights reserved.

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132 Chapter 5 were gradually being replaced by SOCs in both magnitude and diversity across the globe. The commercial synthesis of SOCs can be grouped primarily into six distinct production routes: (1) SOCs derived from “methane”; (2) chemicals derived from “ethylene”; (3) those derived from “propylene”; (4) SOCs derived from “C4s”; (5) those derived from aliphatics; and (6) industrially important SOCs derived from aromatics (Collins & Richey, 1992). SOCs primarily are classified on the basis of their principal uses; they are either cyclic, acyclic, aromatics, or aliphatics. The partial list includes cyclic intermediates, medicinal chemicals, dyes, organic pigments, rubber-processing chemicals, elastomers (synthetic rubber), flavor and perfume materials, plastics and resin materials, plasticizers, pesticides, surface-active agents. SOCs are a very broad category, they also include synthetic compounds, which are volatile organics carbons (VOCs), and relatively newly developed compounds consist of emerging organic contaminants (EOCs). VOCs are organic liquids, mainly composed of industrial solvents, trihalomethanes, gasoline components, refrigerants, etc., while EOCs are pharmaceuticals, veterinary medicines, endocrine-disrupting substances, antimicrobial disinfectants, hormones, personal-care products, illicit drugs, food additives, microplastics, etc. (Lapworth, Baran, Stuart, & Ward, 2012; Postigo & Barcelo, 2015). SOCs are generally found in industrial wastewaters and at wastewater treatment plants. These are point source emissions and diffuse source contamination. On entering into the open environment many SOCs undergo different phototransformations or other chemical reactions or may remain inert. Since they are xenobiotic in origin, their biotransformation is tricky. Moreover, much less information is available for SOCs’ transformation products. Together with their precursors, SOC transformation products exert ecotoxic effects on the biosphere. Some of the most significant SOCs are industrial chemicals of aromatic origin, viz., benzene, toluene, ethylbenzene, xylene, naphthalene, anthracene, etc., and their derivatives. Their persistent and recalcitrant nature in the terrestrial and aquatic environments are because of the complexity of their molecular structures [like polyaromatic hydrocarbon (PAHs)—long-chain aliphatics of petroleum origin], the type and position of substitutions of molecular structures (such as SO3 on synthetic dye molecules), the cyclic nonaromatic compounds (like pesticides, cyclohexane, etc.), and most importantly the aromatic ring/ nucleus is the thermodynamically most stable molecule. Because of their xenobiotic origins, SOCs are not a natural substrate for the enzymatic system of microorganisms or higher plants and animals, including humans. Therefore SOCs often remain recalcitrant for biological metabolism (Lapworth et al., 2012; Postigo & Barcelo, 2015). Remediation, either by physical, chemical, physicochemical, biological, or even integrated methods, is the only option for environmental pollution. It would be incorrect to proclaim the inefficiency of conventional (physicochemical and a few biological) wastewater technologies for providing pragmatic solution for chemical pollutants. The earlier

Emerging bioremediation technologies for the treatment of wastewater 133 technologies were developed to neutralize the existing pollutants, where the chemical complexity was less, the number of polluted sites were fewer, and the toxic effects of SOCs were not elucidated at a higher scale. However, with the advancements in structural chemistry, the evolving needs of human society and the rapid urbanization/industrialization induced by a steep rise in population have increased the complexity and composition of the chemical world over many times This exponential increase in xenobionts and polluted ecosystems has demanded the expansion of remediation technologies toward intelligent applications across the range of environmental pollution. The increasing knowledge and better understanding of mechanisms of operation in bioremediation, microbe pollutant and environment pollutant interactions, and biological response have paved the way for the expansion of bioremediation technologies toward the restoration of pristine environments. Since bioremediation is dependent on the metabolic potential of microorganisms and plants (phytotechnology), the accessibility and availability of organic pollutants to microorganisms is a very critical and limiting factor. A lot has been discussed about “what constitutes the bioavailable fractions of pollutants” (Alexander, 2000; Antizar-Ladislao, 2010; Megharaj, Ramakrishnan, Venkateswarlu, Sethunathan, & Naid, 2011; Vasseur, Bonnard, Palais, Eom, & Morel, 2008). Once SOCs are released into the terrestrial environment from manufacturing and applications sites, they rapidly bind to organic matter in the soil ecosystem, either through physical forces or chemical reactions or via combinations of physicochemical interactions. SOCs and soil interactions are defined by sorption, complexation, or precipitation. The effectiveness of the bioremediation process and the metabolic degradation of SOCs is largely dependent on the ability of soil aggregates to desorb the pollutants. In a terrestrial ecosystem the bioavailability of SOCs is susceptible to the transport of organic compounds into microbial cells by diffusing out from a soil aggregate to reach microbial cells present at the external surface of the aggregate (Megharaj et al., 2011). It is recognized that sorption is a critical factor for bioavailability; the precise mechanism is still poorly understood especially for bioremediation. For the biodegradation of SOCs by microorganisms in a soil ecosystem two differing concepts have been proposed: (1) aqueous solubility, that is, release of organic compounds from soil aggregates into aqueous phase (Harms & Zehnder, 1994; Shelton & Doherty, 1997); and (2) biodegradation of the organic compounds directly into a pollutants soil aggregate composite (i.e., without being desorbed) by enzymes or other microbial metabolites (Megharaj et al., 2011; Singh et al., 2003). PAHs have a strong tendency to get absorbed/adsorbed into soil aggregates, whose biodegradation has been critically enhanced by microbes producing biosurfactants and the development of a vertical gradient between soil aggregates and interfacial pollutants (Megharaj et al., 2011; Tang, White, & Alexander, 1998). Thus bioremediation may be termed as species specific.

134 Chapter 5 The molecular characteristics of SOCs determine its aqueous solubility, volatility, or chemical reactivity which may directly influence their bioavailability in aqueous and soil media. SOCs in their dissolved form (i.e., in pore water) are often considered as bioavailable (and exert biotoxic effect), as compared to the bound organic compound, which has a less toxic or in certain cases no toxic effect. The degree of bioavailability changes with the type of organic compounds, soil types, water content, temperature, and pH of the ecosystem (Megharaj et al., 2011). However, Cornelissen, Rigterink, Ferdinandy, and Van Noort (1998) are of the opinion that for the persistence of (rapidly) desorbing fractions of PAHs nonbiodegraded by bioremediation, microbial parameters were essentially responsible, but not the bioavailability. Another important aspect which increases the persistence of organic pollutants into a soil ecosystem is the sequestration of SOCs. The aging of pollutants in soil may increase the sequestration process due to prolonging the contact period between organic compounds and soil aggregates. Various factors like soil texture and surface area, other organic content, cationic exchange capacity between pollutants and soil organic matter, micropore volume, etc. affect the sequestration of SOCs (Chung & Alexander, 2002; Megharaj et al., 2011). According to Megharaj et al. (2011), for effective bioremediation of SOCs a few significant issues need to be considered, such as (1) the bioavailability and toxic effect of parent compounds, as well as their intermediates and residues in soil ecosystem; (2) the evaluation of remobilized SOCs during the postremediation stage; and (3) the determination of organic compounds which are nontoxic to the biosphere and their end points in the reclaimed soils. Looking at the sophistication required for effective bioremediation, in the present chapter we discuss a few of the emerging biological technologies, describe observations made during their applications, and note challenges for the further improvement and enhancement of the technologies

5.2 Electrobioremediation Electrobioremediation is a hybrid technology, which combines the principle of bioremediation and electrokinetics. The method is widely used for the treatment of hydrophobic organic compounds (Megharaj et al., 2011). The method involves the passing of a direct current between the electrodes in polluted soils, and the placement of the electrodes is critical. The supply of current orients the transport of pollutants and accelerates the degradative mechanism of indigenously present microorganisms of polluted soil (Chilingar, Loo, Khilyuk, & Katz, 1997; Li, Guo, Wu, Li, & Niu, 2010). In electrokinetics a weak electric field of about 0.2 2.0 V/cm is applied (Saichek & Reddy, 2005). The basic phenomena of electrokinetic remediation are electrolysis, diffusion, electrophoresis, electromigration, and electroosmosis.

Emerging bioremediation technologies for the treatment of wastewater 135 Since there is a supply of a direct electric current, the adverse effect on microorganisms is a subject of concern. Moreover, the primary function of the electric current is to extract the pollutants and to transport them over large distances for the metabolism of microorganisms. The impact of a direct electric current on microorganisms compounds and microorganisms soil interactions has been not studied in detail (Wick, Shi, & Harms, 2007). But the study on Sphingomonas sp. LB126—a soil bacterium capable of degrading PAH—showed that there was no negative effect on its degradation efficiency when applying direct current of X 5 1 V/cm; J 5 10.2 mA/cm (Shi, Muller, Harms, & Wicks, 2008). The supply of a current in fact increased the intracellular level of adenosine triphosphate by 60% and boosted other cellular mechanisms, while cellular integrity was not affected. Wick et al. (2007) observed that, due to electroosmosis, the bacteria Mycobacterium frederiksbergense LB501 and Sphingomonas sp. L138 moved from surface to subsurface in PAH degradation experiments used in electrobioremediation. In another study, Niqui-Arroyo and Ortega-Calvo (2007) found a twofold increase in the biodegradation of PAHs in electrokinetically pretreated soil slurries as compared to untreated soils. Like other emerging technologies, electrobioremediation does have a few limitations. The foremost and most significant is the toxic effect of the electrodes used during the treatment procedure. Although studies have demonstrated no negative effect of electrons on the microbial cell, the change in physicochemical surface properties or the breakdown of a dielectric cell membrane and a change in cell metabolism under the influence of the electrodes (and electric current) could not be ruled out. Moreover the solubility of pollutants to be removed, their desorption properties from the soil matrix, and the requirement of conducting pore fluids (such as water) for the mobilization of pollutants into the soil matrix are of great concern. If the contaminated sites contains large amounts of big rocks or gravels, the use of electrobioremediation is highly limited. The large concentrations of particular nonrequired ions at sites also decrease the efficiency. Another significant restriction for the use of electrobioremediation is the availability of suitable microorganisms of contaminated sites (Megharaj et al., 2011; Sogorka, Gabert, & Sogorka, 1998; Velizarov, 1999; Virkutyte, Sillanpaa, & Latostenmaa, 2002)

5.3 Bioelectrochemical systems/technology Bioelectrochemical systems (BESs) are unique technological setups capable of translating chemical energy into bioelectrical energy (and possibly vice versa), employing microorganisms as catalysts (Bajracharya et al., 2016). Technologies like microbial fuel cells (MFCs) are capable of generating bioelectricity from SOCs and other organic wastes, providing a dual advantage of chemical waste removal and greener power generation, without further polluting the environment. Microbial electrolysis cells (MECs) are used to

136 Chapter 5 produce value-added products, such as hydrogen, acetates, and the recovery of nutrients (and metals), etc. from wastewaters (Fatima, Farooqa, Lindstrom, & Saeed, 2017; Maktabifard, Zaborowska, & Makinia, 2018). Therefore BESs are highly regarded as an emerging and promising technology for future wastewater treatment. Different variants of BESs have been developed or are in their development stage and according to Bajracharya et al. (2016) they are broadly classified into electrohydrogenesis systems, electrogenesis systems, microbial electrosynthesis (MES) systems, microbial desalination systems, or bioelectrochemical treatment systems. They are MFCs, MES, microbial electrolysis cells (MECs), enzymatic fuel cells, microbial solar cells (MSCs), plant microbial fuel cells, and microbial desalination cells. These technologies are regarded as energy neutral or energy positive since they simultaneously produce energy (bioelectricity) from organic compounds of wastewater. BESs use electrochemically active microorganisms to catalyze the oxidation and reduction reactions at the cathode and anode in specially designed electrochemical cells. Primarily, two different configurations of MFCs have been developed: (1) dual-chamber MFCs—these have an anaerobic anodic chamber for the oxidation reaction and an aerobic cathodic chamber for the reduction reaction and both chambers are separated by a membrane with ion exchange potential. The anodic chamber consists of anaerobic sludge (wastewater to be treated) along with easily metabolized carbon substrates like glucose (as electron donor), while the cathode consists of aerobic sludge usually with certain electron accepters (Chou, Whiteley, & Lee, 2014). (2) Single chamber MFC(widely used MFCs), consisting of a single compartment, normally an anodic chamber with a microfiltration membrane, an air cathode at the outer side opens to the air and to water on the inner side (Fatima et al., 2017). When a membrane is used in a single chamber MFC, it is placed directly onto the water-exposed side of the cathode. Wastewater, along with suitable carbon source like glucose (as electron donor), is kept at the anode. In MECs electrochemically active bacteria oxidize organic compounds to generate CO2, protons, and electrons. The generated electrons are transferred to the anode and the protons are released into the solution. The electrons travel across a wire mesh toward the cathode and combine with free protons in solution. MFCs in a short span of time have been used for the treatment of a wide range of polluted wastewater, such as textile, pharmaceutical, petrochemical, municipal and sewage, food and agroindustries (dairy, brewery and winery, rice milling industry, molasses and palm-based industrial wastewater animal wastes, etc.), metal (chromium, copper, silver, vanadium, zinc, and cadmium)-containing wastewaters, as well as acid mine drainage. Different studies at the lab scale and pilot scale are being conducted for the treatment of dye manufacturing and textile wastewaters using MFCs. Both these wastewaters contain various simple and

Emerging bioremediation technologies for the treatment of wastewater 137 complex organic compounds, acids, alkalis, metals, and of course unused aromatic dyes in large concentrations. Using MFCs two mechanisms for dye degradation are proposed: abiotic, where dyes act as an electron acceptor in the cathodic chamber; and biotic, where microbial metabolism occurs in the anodic chamber. A dual-chamber MFC was used for abiotic degradation of methyl orange, orange I, and orange II dyes using Klebsiella pneumoniae strain L17 (Liu, Li, Feng, & Li, 2009). The bacterium was inoculated as a biocatalyst in the anodic chamber along with glucose, while dyes containing wastewater were used as the catholyte. Dye degradation was observed in the cathodic chamber, while the maximum power generated was 34.77 mW/m2. In another study, actual dye wastewater was treated using a granular activated carbon-based MFC, that is, granular activated carbon-microbial fuel cell (GACB-MFC), mimicking a dual-chamber MFC (Kalathil, Lee, & Cho, 2011). The chemical oxygen demand (COD) removal during treatment was 71% and 76% at the anode and cathode, respectively, with a maximum electricity generation of 1.7 W/m3. An interesting observation was made that in the anodic chamber the treated dye wastewater was found to be toxic, while at the cathodic chamber the treated wastewater was nontoxic. Besides the textile wastewater, hydrocarbon-containing wastewater from refineries, petrochemicals, hydrocarbon-manufacturing industries, coking, etc. were shown to be remediated using MFCs. For the treatment of coking wastewater, Liu et al. (2009) developed a single-chambered anaerobic fluidized MFC. The system degraded aldehydes, alcohols, and phenols, along with other nitrogenous compounds. The COD removal rate was B84% with maximum power generated at 2.13 mW/m2. For the treatment of petroleum refinery wastewater, Srikanth, Kumar, Singh, Singh, and Das (2016) have developed a mediator-less single-chamber MFC. The system was operated in continuous and batch modes. In continuous mode the treatment efficiency was 84% COD removal and power generation was about 225 1.4 mW/m2, while in batch mode power generation was 54 1.4 mW/m2 and COD removal was 82%. Dual-chamber MFC was used for the treatment of PAHs contamination from diesel. The developed system removed 82% of PAHs with power generation of about 31 mW/m2. In another study, tubular dual- and single-chamber MFCs were used to demonstrate ex situ and in situ treatment of groundwater or refinery wastewater containing a mixture of PHAs, including phenanthrene and benzene (Adelaja, Keshavarz, & Kyazze, 2017). A dualchamber MFC during in situ bioremediation showed 90% COD removal and power generation of 6.75 mW/m2, whereas a single-chamber MFC during ex situ treatment showed 77% COD removal and a power output of 6.75 mW/m2. Various organochloride and substituted phenols, like trichloroethane, 4-chlorophenol, 1-2-dichloroethane, and 2-chlorophenol, used in industrial chlorinated solvents, preservative agents, chlorinecontaining pesticides, or bleaching, etc. were treated using BESs (Bajracharya et al., 2016).

138 Chapter 5

5.4 Phytotechnology (phytoremediation) Like other organisms, plants are affected by environmental pollution and terrestrial pollution directly affects indigenous plants. The effect of pollutants on plants is found to be dependent on their concentration and the response of plants toward the pollutants mainly depends on the type of plants. Calabrese and Blain (2009) suggests that at a low concentration of pollutants a plant gains weight, whereas at higher concentration plant growth is inhibited. The mechanism of the removal of organic pollutants by plants is either by immobilization (and removal) or by promoting microbial degradation (Megharaj et al., 2011). The various mechanisms of the direct involvement of plants are: (1) phytovolatilization, where some SOCs are actively or passively transported across the plant cell membranes and are released through evapotranspiration; (2) phytoextraction, where SOCs are extracted, transported, and accumulate in plant tissues; (3) phytodegradation, SOCs are degraded through enzymatic processes; and (4) phytostabilization, a few volatile SOCs are sequestered into plants rendering them less bioavailable (Megharaj et al., 2011). Phytoremediation has been widely used for the treatment of sewage and municipal wastewaters, tannery effluents, and heavy metal dumping sites and mine fields. Phytoextraction of various heavy metals is a thoroughly studied, understood, and hugely successful method (Khandare & Govindwar, 2015). The phytotechnology-based approach for the treatment of dyes and dye-containing effluents (textile wastewater) is a developing technology. A variety of plants from various habitats, such as garden or ornamental plants, aquatic macrophytes, halophytes, even those from arid ecosystems, have been successfully used for the treatment of textile effluents. Garden and ornamental plants, like Petunia grandiflora, Zinnia angustifolia, Aster amellus, Grindelia grandiflora, Glandularia pulchella, and many others, have been shown to possess an excellent abilities for the degradation of aromatic dye compounds (Kabra, Khandare, Kurade, & Govindwar, 2011; Kabra, Khandare, Waghmode, & Govindwar, 2011; Khandare, Rane, Waghmode, & Govindwar, 2012; Watharkar, Rane, Patil, Khandare, & Jadhav, 2013). Aquatic macrophytes, like Alternanthera philoxeroides, Phragmites australis, Typha angustifolia, Typha domingensis, Typhonium flagelliforme, etc., were used for the treatment of dye-containing wastewaters (Carias, Novais, Martins-Dias, & Novais, 2008; Nilratnisakorn, Thiravetyan, & Nakbanpote, 2007; Rane et al., 2015; Shehzadi et al., 2014). Small trees such as R. rhabarbarum were reported to degrade sulfonated anthraquinones (a precursors for dye compounds), while Sesuvium portulacastrum, a halophyte, was found to decolorize azo dye Green HE4B under high salinity (Khandare & Govindwar, 2015; Patil, Lokhande, Suprasanna, Bapat, & Jadhav, 2012).

5.4.1 Phytoreactors and constructed wetlands An advanced form of phytoremediation includes the development of constructed wetlands (CWs) for in situ or ex situ treatment of dye-containing wastewaters. CWs are large-sized

Emerging bioremediation technologies for the treatment of wastewater 139 indigenously developed open reactor systems, designed to mimic actual conditions prevailing at wastewater treatment plants or in natural polluted ecosystems (Khandare & Govindwar, 2015). Different types of CWs have been widely used at lab scale and pilot scale for the treatment of dyes and textile effluents. For the treatment of azo dye Acid Orange 7, a vertical flow CW was designed with a capacity of 1 m3, along with feeding reservoir, using P. australis plant. The phytoreactor was batch-fed with 700 L of dye wastewater at a concentration of 130 700 mg/L. The system consisted of a basal gravel layer of sandy clay soil to support the plant growth with a drainage facility. A submersible centrifuge pump and flowmeter was used for mixing and feeding and to control the inlet flow, respectively. This phytoreactor system was found to remove 69% and 67% of COD and total organic carbon (TOC), respectively (Davies, Carias, Novais, & Martins-Dias, 2005; Davies, Ferreira, Carias, & Novais, 2009). Bulc and Ojstrsek (2008) developed vertical and horizontal flow triple bedded interconnected pilot-scale phytoreactors for the treatment of textile effluent. P. australis plants were grown on a basal sand and gravel layer. The system was found to remove textile dyes and decreased COD and biochemical oxygen demand (BOD) by 84% and 66%, respectively. Total nitrogen, organic nitrogen, ammonia, sulfate ions, surfactants, total suspended solids (TSS), and TOC were also reduced by 52%, 97%, 331%, 88%, 80%, 80%, and 89% respectively.

5.4.2 Plant microbe phytoremediation Plant microbe phytoremediation, including rhizoremediation, is one of the most powerful tools of phytotechnology. Rhizoremediation can occur naturally or be induced or triggered by augmenting specific pollutant-degrading microbes or even plant growth-promoting microorganisms (Gerhardt, Huang, Glick, & Greenberg, 2009). In a vertical subsurface flow CW, the arid plants G. pulchella and Po. grandiflora were used in two independent systems for the treatment of textile wastewater. The system was designed such that half of the grown roots were retained in soil and the other half in wastewater, with a continuous circulatory flow to increase the retention time and exposure time of the wastewater to plants. The system was augmented with potential dye-degrading bacteria; the efficiency of the phytoreactor was enhanced and 93%, 66%, 70%, and 74% of ADMI, COD, BOD, and TOC, respectively, were removed from textile wastewater (Kabra, Khandare, & Govindwar, 2013). The study claimed to achieved better efficacy in a plant microbe system than the individual plants and bacterial studies. In another study, in a lab-scale batch rector, Paspalum crinitum was grown on a thermocol base, where roots are exposed to textile wastewater at the bottom of the reactor. Bacterial cells of Bacillus pumilus were immobilized and augmented directly into wastewater near to the root system of the plants. Post-augmentation, the treatment potential of P. crinitum was found to be enhanced and 93%, 78%, and 70% of ADMI, COD, and BOD/TSS,

140 Chapter 5 respectively, were removed from real textile wastewater after 12 d of treatment (Watharkar et al., 2015). The augmentation of Pseudomonas putida in a Po. grandiflora-based phytoreactor showed a considerable reduction in various chemical parameters of textile wastewater after 48 h of treatment. While the treatment time increased to 72 d for individual plant and bacterial systems in order to achieve a similar treatment efficiency (Khandare, Kabra, Kadam, & Govindwar, 2013). Rane et al. (2015) developed a similar system, but they have used a rhizofiltration approach in a lab-scale semipilot phytoreactor of 340 L (1.2 m 3 0.61 m 3 0.61 m). A. philoxeroides plants were grown on an iron grid in a steel rector, where roots were allow to float and grow in the textile wastewater directly. In a 96 h treatment wastewater with wide range of pH (3.5 10.5) was treated and significant reductions of COD, BOD, and other chemical parameters were observed. In rhizoremediation the root length is an important consideration. Many of the plants used in the phytoremediation are herbaceous, and thus their root length and depth vary from plant to plant, depending on the type of soils, and seasonally. If the pollution is deep in the soil matrix, natural phytoremediation is difficult to achieve. In such conditions, it requires the excavation of the pollutants, along with polluted soils or the specific selection of trees with deeper roots system. Dendroremediation, another form of phytoremediation, uses trees (with deep root systems) instead of herbaceous plants and the treatment of SOCs like trichloroethylene and 2,4,6-trinitrotoluene from groundwater and soils were reported (Susarla, Medina, & McCutcheon, 2002).

5.4.3 Plant enzymes and metabolites Plant-dependent phytoremediation is mainly mediated by enzymes. Dye decolorization and degradation using plants relies on a number of plant-originated enzymes. Peroxidases from Ipomoea palmate were found to decolorize various dye compounds, such as Brilliant Green, Methyl Orange, Crystal Violet, etc. Likewise, Saccharum spontaneum-produced peroxidases showed excellent decolorization and degradation of Supranol Green (Khandare & Govindwar, 2015). Peroxidases obtained from Brasicca rapa have been shown to degrade various textile dyes (Khandare & Govindwar, 2015). Plant-based lignin peroxidases, laccases, polyphenol oxidases, and other plant-based enzymes have been shown to degrade aromatic dyes either in their free form or in an immobilized state (Khandare & Govindwar, 2015). Besides enzymes, plants produce “secondary plant metabolites” (SPMEs), such as root exudates, allelopathic chemicals, phytosiderophores, phytohormones, phytoalexins, and phytoanicipins, which are derived from alkaloid, fatty acid/polyketide, isoprenoid, and phenylpropanoid pathways (Hadacek, 2002; Megharaj et al., 2011). These metabolites are found to be analogues of organic xenobiotic compounds within the network of

Emerging bioremediation technologies for the treatment of wastewater 141 suprametabolism, and may aid in the metabolism of xenobiotics by plants producing them (Singer, Thompson, & Bailey, 2004). Kuiper, Kravchenko, Bloemberg, and Lugtenberg (2002) isolated P. putida PCL1444 from rhizosphereic soils of Lolium multiflorum cv. Barmultra, when the plant was grown in soils polluted with PAH. The bacterium found to utilize root exudates of plants for its growth and protects plants from the toxic effect of PAHs by the higher transcription of naphthalene catabolic genes. SPMEs like cymene, pinene, carvone, and liminene showed enhanced degradation of polychlorinated biphenyls (PCBs) (Gilbert & Crowley, 1998; Kim, Oh, So, Ahn, & Koh, 2003). In another study, because of the exudation of SPMEs (like phenylpropanoids), gfp-tagged P. putida PML2 was shown to increase the growth of bacterium and enhanced the PCB degradation (Megharaj et al., 2011). Siciliano et al. (2001) suggested that the number of pollutant-degrading endophytes and their activities are plant and pollutant dependent. Methylobacterium sp. strain BJ001, a phytosymbiont, was isolated from tissue culture plantlets of Populus deltoids 3 nigra DN34, and showed the degradation of 2,4,6-trinitrotoluene, as well as the complete mineralization of octahydro-1,3,5,7-tetranitro-1,3,5-tetrazocine and hexahydro-1,3,5-trinitro-1,3,5-triazine into CO2 (Van Aken, Yoon, & Schnoor, 2004).

5.4.4 Hydroponic systems Hydroponic systems are another variant of phytoremediation, where plants are grown in a soil-less environment directly into the wastewater. Hydroponic systems have been used for the treatment of wastewater containing textiles dyes. Dyes such as Acid Blue 92, Basic Red 46, and Reactive Blur 19 were degraded using hydroponically grown duckweed Lemna minor, Nasturtium officinale, and T. angustifolia, respectively (Khataee, Movafeghi, Vafaei, Lisar, & Zarei, 2013; Mahmood et al., 2014; Torbati, Khataee, & Movafeghi, 2014). The treatment of organic compounds, such as sulfonated anthraquinones (widely used in synthetic dye synthesis), was demonstrated using hydroponic systems. Four hydroponic plants, Rumex acetosa, Apium graveolens, Rheum rhabarbarum, and Rumex hydrolapatum, were screened for their ability to degrade sulfonated anthraquinones (Aubert & Schwitzguebel, 2004). Paquin, Sun, Tang, and Li (2006), during their study on the degradation of dye compounds Poly R-478, found that 61%, 59%, 88%, and 66% of dye was decolorized by Eleocharis calva, Cyperus javanicus, Pennisetum purpureum, and Fimbristylis cymosa, respectively, at an initial dye concentration of 20 mg/L after 1 4 weeks of treatment. Herbaceous plants like Hibiscus furcellatus and Rumex crispus decolorize Poly R-478 by 55% and 77%, respectively.

5.4.5 Plant tissue culturing After years of lab-scale studies, the actual mechanism of wastewater treatment using phytoremediation is still unclear. Tissue culture offers an added advantage in studying the

142 Chapter 5 basic phytoremediation mechanism, metabolic capabilities, and toxicity tolerance by selecting plants or model plants in culturing. Various types of plant cultures, for example, hairy roots, cell suspensions, whole plants, and calli, are commonly used in phytoremediation studies at the lab scale (Khandare & Govindwar, 2015). A drawback of phytoremediation by wild-type species is their seasonal availability, but this can be circumvented using the plant tissue culture approach. Tissue cultured plants can be grown under controlled conditions and they can be controlled for indefinite propagation and to increase their natural life span. Moreover, in order to understand the actual potential, response, and mechanisms of pollutants degradation, in vitro culturing is very useful. Since the method required to maintain the sterile environment (i.e., free of microbes), it can distinguish the plant response from microbial action (Khandare & Govindwar, 2015). In a typical phytoremediation process, usually roots are the first plant organs to come into contact with the pollutants. Root culturing (hairy roots) was demonstrated to degrade phenolics, PCBs, trinitrotoluene, and pharmaceuticals, and to detoxify heavy metals (Guillon, Tremouillaux-Guiller, Pati, Rideau, & Gantet, 2006). Hairy root culturing was shown to decolorize Reactive Red 198 by Tagetes patula, Solanum indicum, Nicotiana tabacum, and Solanum xanthocarpum (Patil, Desai, Govindwar, Jadhav, & Bapat, 2009). Callus and suspension culturing methods are also used in phytoremediation. The use of suspension cell culturing and callus would enable ideal systems for understanding the biochemical basis of organic compound degradation. As early as 1999, Duc, Vanek, Soudek, and Schwitzguebel (1999) showed the potential of the cell suspension of Rheum palmatum for the accumulation of sulfonated anthraquinones. Cell suspension of B. malcolmii has been shown to degrade triphenylmethane dye Malachite Green (Kagalkar, Jadhav, Bapat, & Govindwar, 2011). The use of in vitro grown whole plantlets could also provide a basic understanding of pollutant degradation and the response of plants using phytoremediation. For the standardization of various parameters, the use of whole plantlets is more advantageous, and using whole plants under in vitro conditions for initial optimization steps would provide important results for pilot-scale applications. T. flagelliforme, Pe. grandiflora, Nopalea cochenillifera, G. pulchella, and Ipomoea hederifolia were widely used in aromatic-substituted dye degradation (Khandare & Govindwar, 2015).

5.5 Electron beam irradiation Electron beam (EB) irradiation has been widely used in the consumer goods industries and even food-packing units for surface sterilization. In recent years the International Atomic Energy Agency has been collaborating with different wastewater treatment plants in a few countries, such as China and South Korea, to develop an advanced EB-based technology. EB irradiation is a process that involves the use of electrons with high internal energy

Emerging bioremediation technologies for the treatment of wastewater 143 which causes the degradation of complex high-molecular-weight organic molecules (Capodaglio, 2018). The EB device is embedded in a sealed container which is kept under high vacuum, whilst a heated emitter, that is, a cathode, releases a beam of high-energy electrons that are accelerated in a narrow gapped grid under a high-voltage direct current power supply or radiofrequency (Capodaglio, 2018). Magnetic or electrostatic fields are used for controlling (by focusing and deflecting) the high-energy beams and they are focused toward the exit window. The energy and intensity of the electrons emitted from the window is dependent on the voltage applied at the anode and their number is dependent on the current applied at cathode. By regulating both these parameters beam penetration and applied dose (absorbed) rate can also be controlled. During wastewater treatment EB has the potential to generate strong oxidants as well as reducers from H2O molecules (superoxides, i.e., hydroxyl radicals •HO2, hydrogen atoms 2 H, hyperoxides O2 2 , solvated electrons e(aq) ), which aid in the degradation of organic pollutants. These species are very short-lived, extremely reactive, with a half-life in the order of 10 µs at 1024 M concentration and oxidation potential even stronger than O3 (Capodaglio, 2018; von Gunten, 2003). By controlling the irradiation dose and the density of the reactive species, the degree of organic compound degradation can also be controlled to obtain partial or complete degradation (mineralization of pollutants). The EB technology in wastewater treatment is dose dependent, which would be in the range of 1 4 kGy and an average beam current of 20 mA. In one of the field-scale applications of EB, 10,000 m3/day of textile wastewater was treated in a treatment plant at Daegu, South Korea (Kuk et al., 2011). EB irradiation was applied as a primary or pretreatment for the partial degradation of wastewater to break/ degrade the high-molecular-weight organic compounds, mainly unused dyes and starch. The irradiated wastewater was then treated with a biological method, which increased the efficiency of the overall treatment by 40%. The treatment time, usage of chemical reagents, and retention time drastically decreased with EB-treated textile wastewater. In another study, dyes and pesticides were efficiently removed by radiation. It has been demonstrated that with the addition of TiO2 prior to EB irradiation the degradation efficacy of pesticides was improved (Emmi, De Paoli, Takacs, Caminati, & Palfi, 2008; Solpan, 2008a, 2008b; Takacs, Wojnarovits, Palfi, & Emmi, 2008; Trojanowicz et al., 2008). The advantage of EB irradiation technology is that it does not generate any secondary contaminants, unlike conventional physicochemical or even some biological methods. The initial success with few large-scale demonstrations suggests that it does not require any additional chemicals, it is relatively energy efficient, and can easily be implemented at conventional wastewater plants. The integration with existing technologies at treatment plants, either as a pre-treatment or post-treatment, can easily increase the efficiency of the treatment to achieve the best possible treatment goals at the lowest possible cost. It is

144 Chapter 5 claimed that post-treatment with EB, no residual radiation remains in the treated medium and no residual radioactivity remains in the electron accelerator or its components (Capodaglio, 2018). The generated radicals in fact would revert back to the original water molecules, within a few milliseconds, if they do not react with any organic compounds or pollutants, thereby no radical residues or radioactivity are left in the water. Although EB irradiation is a non-biological method, it is highly imperative to mention and discuss it in this chapter, since EB greatly reduces the complexity of wastewater, can be easily treated further by biological technologies. Theoretically, the integration of EB irradiation and the bioremediation method could be a very promising and highly futuristic technology which may provide the correct solution for wastewater treatment.

5.6 Conclusion: unresolved challenges and future perspectives During the early stages of the industrial revolution, the environmental effects of many of the SOCs were not anticipated. Soon, the recalcitrant nature and ecotoxic impacts of organic pollutants on the biosphere were recognized and understood. Moreover, the universal problems of water pollution, water scarcity, and water recycling are posing serious challenges to both developing and developed countries. In particular, the presence of SOCs is a matter of growing concern. Many technologies for the removal of pollutants have been developed and depending upon the nature of pollution and characteristics of pollutants the best available technologies have been applied. These technologies, including physical, chemical, physico-chemical, and biological methods, have provided much needed initial success, but with the ever-emerging newer SOCs, the need for the constant evolution of remediation technologies still remains. The increasing cost and inherent limitations of non-biological methods made bioremediation more attractive. Bioremediation is a very promising technology. It directly uses organisms (autochthonous or allochthonous) or their products (viz. enzymes) for either in situ or ex situ bioremediation. Even though the efficacy of bioremediation made the technology versatile and complete, there is a constant battle between the technology applied and the emerging pollutants. The effectiveness of bioremediation methods relies on numerous factors; the challenges for existing technologies are still revolving around these factors. Although bioavailability is a bottleneck in bioremediation, other significant parameters like oxygen concentration (many SOCs required anaerobic conditions, while PAHs required an oxygenrich environment for their degradation), temperature, nutrient availability, cosubstrates requirements, the presence of potential toxic compounds, physiological state, the metabolic potential of indigenous microorganisms, and most importantly the molecular structure of SOCs have to be considered for the effective degradation of organic compounds at polluted sites (Romantschuk et al., 2000).

Emerging bioremediation technologies for the treatment of wastewater 145 Currently a large number of technologies are still at the prototype stage, while a few have been applied at large-scale or field level at polluted sites. There are a few successful and commercial options for bioremediation. But the need for new and emerging technologies is because each industrially polluted site is unique, the nature of pollutants are different, and, as mentioned above, geographical and environmental conditions also vary. Therefore it requires the assessment of all relevant factors before new tools are developed and widely applied worldwide, however the lack of data, poor knowledge of newly developed processes, uncertainty in prediction of SOCs emissions, and over parameterization are still a challenge. Although phytotechnology is a promising method, it is still under developed and in its nascent phase. The knowledge gap and technical acquaintance between lab-scale results and field level applications is considerable, especially for the treatment of textile wastewaters and PAH-contaminated sites. Since plants are the main driver of the technology, the availability of plants across the year is one of the major limitations. Many plants species reported for phytotechnology are seasonal, therefore extensive screening of plants which can be available across the year is still required. While the plant tissue culture option can be used for continuous propagation, the actual load of pollutants and their practical handling of hydraulics would be entirely different, and at polluted sites it is constantly changing. Due to the rate and the intensity at which environments are being polluted, the need of the hour is not only to find suitable technology but also the ‘time’ consideration is highly important. Since the phytotechnology is dependent on plants, the time required for treatment in a batch or continuous mode is very high. To improve the time considerations, the technology either needs to be integrated with other physicochemical and biological methods or could be augmented with acclimatized microorganisms. Most importantly, phytotechnology, unlike other conventional methods, cannot be applied as a primary or secondary treatment technology. It is still being used as a tertiary or polishing step (at few of the wastewater treatment centres) before the disposal of treated wastewater into an open environment. In BESs, functioning and application is a multidisciplinary approach and requires experts from many disciplines like microbiology, biotechnology, environmental science, bioelectrochemistry, electrochemistry, material science, engineering, etc. However, with the involvement of multidisciplinary aspects, BESs are still in their infancy. Conventional bioremediation involving microorganisms or plants is dependent on the conversion of substrates (electron donor and organic pollutants), the biocatalytic and metabolic potentials of the biological source, the specificity of the enzymes involved in the cascade of reactions and entire pathways in the conversion process, and redox conditions (Bajracharya et al., 2016). However, BESs are also dependent on the conductive potential of anolytes and

146 Chapter 5 catholytes (electrolytes), cathode and anode potentials, voltage losses, the mediators involved, charge and mass transfer, and many other factors. BESs are a combinations of both biological and electrochemical processes, and thus many limitations and challenges are complex and difficult to understand. Another major technological concern and matter of further study is voltage loss (i.e., the reduction of overpotentials) and high coulombic efficiencies. Like other emerging technologies, translating the lab-scale results into the industrial level, as well as the upscaling of bioelectricity production at larger level, is a major challenge in BESs technologies. Even after existence of various bioremediation methods the biological response to synthetic pollutants largely remains unknown and various studies have suggested that it changes within a microbial guild (Ramakrishnan, Megharaj, Venkateswarlu, Naidu, & Sethunathan, 2010). If bioremediation is to take center stage and micro-bioremediation along with other emerging methods (viz., phytoremediation) is to improve its competence, the understanding of metabolic cooperation amongst the microbial communities and plant microbes interactions needs to be enhanced. The development of next-generation sequencing and fingerprinting technologies are widening the scope of bioremediation. The structure and functions of microbial communities at polluted sites are now being studied on a real-time basis at actual micro-community level along with differential temporal and spatial scales. With numerous studies on bioremediation or even with other non-biological methods, it has been hard-earned lesson that, it is nearly impractical to restore all the contaminated sites and to restore all natural functions of industrially polluted sites with the current state of technologies. The ever-changing dynamics of pollutants and wastewater content, developing a universal treatment protocol is unrealistic. However, the application of the principal of the function-directed remediation approach might minimize the risks of persistence and the diffusion of recalcitrant synthetic compounds. We need to properly address the microdynamics of polluted sites to develop eco-innovations in remediation technology.

Acknowledgments The authors are grateful to the Department of Biotechnology, Ministry of Science and Technology, New Delhi for financial support (BT/Env/BC/01/2014).

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Emerging bioremediation technologies for the treatment of wastewater 149 Niqui-Arroyo, J. L., & Ortego-Calvo, J. J. (2007). Integrating biodegradation and electroosmosis for the enhanced removal of polycyclic aromatic hydrocarbons from creosote-polluted soils. Journal of Environmental Quality, 36, 1444 1451. Paquin, D., Sun, W., Tang, C., & Li, Q. (2006). A phytoremediation study: Selection of tropical and other vascular plants for decolorization of Poly R-478 dye. Remediattion Journal, 16, 97 107. Patil, A., Lokhande, V., Suprasanna, P., Bapat, V., & Jadhav, J. (2012). Sesuvium portulacastrum (L.) L.: A potential halophyte for the degradation of toxic textile dye, Green HE4B. Planta, 235, 1051 1063. Patil, P., Desai, N., Govindwar, S., Jadhav, J., & Bapat, V. (2009). Degradation analysis of Reactive Red 198 by hairy roots of Tagetes patula L. (Marigold). Planta, 230, 725 735. Postigo, C., & Barcelo, D. (2015). Synthetic organic compounds and their transformation products in groundwater: Occurrence, fate and mitigation. Science of the Total Environment, 503-504, 32 47. Ramakrishnan, B., Megharaj, M., Venkateswarlu, K., Naidu, R., & Sethunathan, N. (2010). The impacts of environmental pollutants on microalgae and cyanobacteria. Critical Reviews in Environmental Science and Technology, 40, 699 821. Rane, N., Chandanshive, V., Watharkar, A., Khandare, R., Patil, T., Pawar, P., & Govindwar, S. (2015). Phytoremediation of sulfonated Remazol Red dye and textile effluents by Alternanthera philoxeroides: An anatomical, enzymatic and pilot scale study. Water Research, 83, 271 281. Romantschuk, M., Sarand, I., Petanen, T., Peltola, R., Jonsson-Vihanne, M., Koivula, T., et al. (2000). Means to improve the effect of in situ bioremediation of contaminated soil: An overview of novel approaches. Environmental Pollution, 107, 179 185. Saichek, R. E., & Reddy, K. R. (2005). Electrokinetically enhanced remediation of hydrophobic organic compounds in soil: A review. Critical Reviews in Environmental Science and Technology, 35, 115 192. Shehzadi, M., Afzal, M., Islam, E., Mobin, A., Anwar, S., & Khan, Q. (2014). Enhanced degradation of textile effluent in constructed wetland system using Typha domingensis and textile effluent-degrading endophytic bacteria. Water Research, 58, 152 159. Shelton, D. R., & Doherty, M. A. (1997). A model describing pesticide bioavailability and biodegradation in soil. Soil Science Society of America Journal, 61, 1078 1084. Shi, L., Muller, S., Harms, H., & Wicks, L. Y. (2008). Effect of electrokinetic transport on the vulnerability of PAH-degrading bacteria in a model aquifer. Environmental Geochemistry and Health, 30, 177 182. Siciliano, S., Fortin, N., Mihoc, A., Wisse, G., Labelle, S., Beaumier, D., et al. (2001). Selection of specific endophytic bacterial genotypes by plants in response to soil contamination. Applied and Environmental Microbiology, 67, 2469 2475. Singer, A. C., Thompson, I. P., & Bailey, M. J. (2004). The tritrophic trinity: A source of pollutant degrading enzymes and its implications for phytoremediation. Current Opinion in Microbiology, 7, 239 244. Singh, N., Megharaj, M., Gates, W. P., Churchman, G. J., Anderson, J. A., Kookana, R. S., et al. (2003). Bioavailability of an organophosphorus pesticide, fenamiphos, sorbed on an organo-clay. Journal of Agricultural and Food Chemistry, 51, 2653 2658. Sogorka, D. B., Gabert, H., & Sogorka, B. J. (1998). Emerging technologies for soils contaminated with metals—electrokinetic remediation. Hazardous Industrial Waste, 30, 673 685. Solpan, D. (2008a). Decoloration and degradation of some textile dyes by gamma-irradiation. In Radiation treatment of polluted water and wastewater. Report IAEA-TECDOC-1598. International Atomic Energy Agency, Industrial Application in Chemistry Section. Vienna: IAEA. Ş olpan, D. (2008b). The degradation of some pesticides in aqueous solutions by gamma radiation. In Radiation treatment of polluted water and wastewater. Report IAEA-TECDOC-1598. International Atomic Energy Agency, Industrial Application in Chemistry Section. Vienna: IAEA. Srikanth, S., Kumar, M., Singh, D., Singh, M. P., & Das, B. P. (2016). Electro-biocatalytic treatment of petroleum refinery wastewater using microbial fuel cell (MFC) in continuous mode operation. Bioresource Technology, 221, 70 77. Susarla, S., Medina, V. F., & McCutcheon, S. C. (2002). Phytoremediation: An ecological solution to organic chemical contamination. Ecological Engineering, 18, 647 658.

150 Chapter 5 Takacs, E., Wojnarovits, L., Palfi, T., & Emmi, S. S. (2008). Irradiation treatment of textile dye containing wastewater. In Radiation treatment of polluted water and wastewater. Report IAEA-TECDOC-1598. International Atomic Energy Agency, Industrial Application in Chemistry Section. Vienna: IAEA. Tang, W. C., White, J. C., & Alexander, M. (1998). Utilization of sorbed compounds by microorganisms specially isolated for that purpose. Applied Microbiology and Biotechnology, 49, 117 121. Torbati, S., Khataee, A., & Movafeghi, A. (2014). Application of watercress (Nasturtium officinale R. Br.) for biotreatment of a textile dye: Investigation of some physiological responses and effects of operational parameters. Chemical Engineering Research and Design, 92, 1934 1941. Trojanowicz, M., Drzewicz, P., Bojanowska-Czajka, A., Nałec̨ z-Jawecki, G., Gryz, M., Sawicki, J., . . . Zimek, Z. (2008). Application of ionizing radiation for removal of pesticides from groundwaters and wastes. In Radiation treatment of polluted water and wastewater. Report IAEA-TECDOC-1598. International Atomic Energy Agency, Industrial Application in Chemistry Section. Vienna: IAEA. Van Aken, B., Yoon, J. M., & Schnoor, J. L. (2004). Biodegradation of nitro-substituted explosives 2,4,6trinitrotoluene, hexahydro-1,3,5-trinitro-1,3,5-triazine, and octahydro-13,5,7-tetranitro- 1,3,5-tetrazocine by a phytosymbiotic Methylobacterium sp. associated with poplar tissues (Populus deltoides 3 nigra DN34). Applied and Environmental Microbiology, 70, 508 517. Vasseur, P., Bonnard, M., Palais, F., Eom, I. C., & Morel, J. L. (2008). Bioavailability of chemical pollutants in contaminated soils and pitfalls of chemical analyses in hazard assessment. Environmental Toxicology, 23, 652 656. Velizarov, S. (1999). Electric and magnetic fields in microbial biotechnology: Possibilities, limitations and perspectives. Electro-Magnetobiology, 18, 185 212. Virkutyte, J., Sillanpaa, M., & Latostenmaa, P. (2002). Electrokinetic soil remediation—critical review. Science of the Total Environment, 289, 97 121. von Gunten, U. (2003). Ozonation of drinking water: Part I. Oxidation kinetics and product formation. Water Research, 37, 1443 1467. Watharkar, A., Khandare, R., Waghmare, P., Jagadale, A., Govindwar, S., & Jadhav, J. (2015). Treatment of textile effluent in a developed phytoreactor with immobilized bacterial augmentation and subsequent toxicity studies on Etheostoma olmstedi fish. Journal of Hazardous Materials, 283, 698 704. Watharkar, A., Rane, N., Patil, S., Khandare, R., & Jadhav, J. (2013). Enhanced phytotransformation of Navy Blue RX dye by Petunia grandiflora Juss. with augmentation of rhizospheric Bacillus pumilus strain PgJ and subsequent toxicity analysis. Bioresource Technology, 142, 246 254. Wick, L. Y., Shi, L., & Harms, H. (2007). Electro-bioremediation of hydrophobic organic soil contaminants: A review of fundamental interactions. Electrochimica Acta, 52, 3441 3448.

CHAPTER 6

Bacterial quorum sensing in environmental biotechnology: a new approach for the detection and remediation of emerging pollutants Debapriya Sarkar, Kasturi Poddar, Nishchay Verma, Sayantani Biswas and Angana Sarkar Department of Biotechnology and Medical Engineering, National Institute of Technology Rourkela, Odisha, India

6.1 Introduction Quorum sensing (QS) of bacteria or microbes can be defined as a stimuli or response which can be visible as a phenotypic change, correlated to microbial population. It is the cell to cell communication phenomenon of prokaryotes. This prokaryotic cell to cell signaling is mediated by autoinducers (AI) which are formed and secreted by the microbes (Williams, 2007). QS is the mechanism of communication between the microorganisms depending upon their cell density which can have an influence on behaviors like virulence factor secretion, biofilm formation, bioluminescence competence, and others in bacteria (Miller & Bassler, 2001). The first QS in bacteria was observed in Streptococcus pneumoniae on genetic competence and in marine Vibrio sp. on bioluminescence. This communication depends on specialized AI (signaling) molecules (Albuquerque & Casadevall, 2012). QS can be well explained by the process of releasing toxic materials by the pathogenic microbes in the host cell. Toxic chemicals or host disrupting materials are not abruptly released by pathogenic microbes. When the number of microbes is low, their virulent genes do not get expressed. But when they increase in number, the virulent genes get activated. This helps microbes to ensure the maximum impact of virulent genes (Williams, 2007). This cell to cell coordination is an example of QS. Similarly, biofilm formation is also directly related to the size of the bacterial cell population. Single cells cannot form biofilm individually. When there is a significant population size, the genes for extracellular matrix get activated and a biofilm can form. These kinds of coordinated functions by single-celled Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00006-7 © 2020 Elsevier Inc. All rights reserved.

151

152 Chapter 6 prokaryotes are achieved by QS (Niu, Clemmer, Bonomo, & Rather, 2008). The activated genes, dependent upon the size of population, are actually activated by the AI, secreted by the microbes themselves. When the size of the population is very low, the concentration of the AI is also very low; hence it has no effect on the corresponding genes. But as population size increases, AI concentrations also increase and activate those genes. QS has also been observed in the soil microbial system where they maintain a symbiotic environment or growth competition (Fray, 2002). A similar kind of effect has also been reported for insects and plants. Social insects, for example, ants and honeybees, show colonization with the help of QS. It is also know as a de-centralized system since no individual is responsible for decision-making in the colonies. It depends on the population or group of individuals (Franks et al., 2015). Prokaryote eukaryote interactions (especially with plants) also seem to be carried out by QS molecules, such as N-acyl-homoserine-lactone. Arabidopsis thaliana has been taken as a model organism to understand the effect of the QS molecules on plants. It has been found that the short chain of AHL has an influence on cellular signaling in plants. Therefore QS molecules impact on the physiological health of plants along with ecological effects (Mohan, Benton, Dangelmaier, Fu, & Sekhar, 2018). This chapter will detail the microbe-mediated quorum-sensing phenomenon, particularly regarding bacterial QS activity. This chapter will emphasize the basic working mechanisms of bacterial QS methods and their application in the area of environmental biotechnology. This chapter will be particularly dedicated to the quorum sensor-based detection and bioremediation of inorganic, organic, and pathogenic substances.

6.2 Mechanisms of bacterial quorum sensing In order to perform QS, bacteria should possess certain characteristics: (1) be able to secrete signaling molecule, an autoinducer molecule which identifies the concentration of signaling molecules; and (2) be able to control the gene expression of cells in response. The principle of QS is based on the diffusion of signaling molecules. At a low concentration of signaling molecules they diffuse away, but as the population and cell density increases the concentration of signaling molecules increases and once the concentration reaches the threshold level it triggers or controls gene expression (Williams & Ca´mara, 2009). The basic mechanism of QS in bacteria follows three steps; 1. A specific type of signaling molecule is secreted in the outer space of the cell. 2. Molecule accumulation takes place with the increase in the cell population. 3. At the final stage, after reaching the threshold concentration the signaling molecule is sensed by the bacteria and triggers the series of regulatory activities (Podbielski & Kreikemeyer, 2004).

Bacterial quorum sensing in environmental biotechnology 153 Bacteria have the ability to control their gene expression at the population level using the principle of QS. Oligopeptides act as signaling molecules in Gram-positive bacteria. These signaling molecules are detected either by surface receptors using two-component systems or by internalization of oligopeptides inside cell using oligopeptide transport. These signaling molecules then interact with the transcriptional regulators and thus help in controlling the gene expression (Chen et al., 2002). These regulators help in controlling various functions such as virulence, competence, conjugation, etc. Bacilli, Streptococci, and Enterococci are some of the Gram-positive bacterial genera in which such a mechanism of gene expression control is observed (Monnet & Gardan, 2015).

6.2.1 Two-component system in Gram-positive bacteria QS in Gram-positive bacteria is carried out by oligopeptides; these peptides are impermeable to the membrane, and therefore the peptides secretion for QS is carried out by special transporters. In Gram-positive bacteria, the peptide synthesis initially undergoes certain processes such as modification and cyclation followed by secretion. The sensor of peptides in Gram-positive bacteria is membrane bound. These are known as two-component systems which include a cognate receptor and a histidine kinase receptor. These twocomponent systems transfer information via a series of phosphorylation events. Cognate receptors are the cytoplasmic response receptors which also act as the transcriptional regulators (Hawver, Jung, & Ng, 2016). The precursor oligopeptide is internally synthesized by bacteria and is modified and transported to the extracellular environment via special transporter molecules. The oligopeptides concentration increases with the increase in the cell density or population, and when the concentration increases to the threshold concentration it is recognized by the membrane-bound receptors. This membrane-bound receptor is a histidine kinase receptor to which the oligopeptide binds and autophosphorylation activity is activated. The phosphate group is transferred to the cytoplasmic response receptor called a cognate receptor which is activated, and the response also controls the expression of the targeted gene (Fig. 6.1) (Monedero, Revilla-Guarinos, & Zuniga, 2017).

6.2.2 Acyl homoserine lactone in Gram-negative bacteria In Gram-negative bacteria, acyl homoserine lactone (AHL) is used as the autoinducer molecule for QS. LuxI and LuxR are the two important components in the Gram-negative bacteria. LuxI is an enzyme that synthesizes the N-3-(oxyhexanoyl)-homoserine lactone. LuxI catalyzes the acylation and lactonation reaction between two molecules, adenosylmethionine and hexanoyl acyl carrier protein, to form 3OC6HSL. After the synthesis of this acyl homoserine lactone, it is transferred to the extracellular environment and its concentration increases with the increase in the cell density. When it reaches the

154 Chapter 6

Extracellular environment Modified oligopeptide

Transporter molecule

Cell surface H

H

Histidine kinase

PO4

Cognate receptor

Cytosol

Precursor oligopeptides

Precursor gene

Target gene

Figure 6.1 The schematic representation of the mechanism of quorum sensing (QS) in Gram-positive bacteria. The figure shows the precursors formed inside the cell get modified and transported outside the cell. When their concentration increases they bind to the membrane-bound histidine kinase which undergoes phosphorylation and transfers phosphate to a cognate receptor which gets activated and hence controls the gene expression.

threshold concentration it is recognized by the LuxR which in response controls the expression of targeted gene (Fig. 6.2) (Papenfort & Bassler, 2016).

6.3 Quorum sensing in environmental biotechnology In recent years QS has been commonly used in environmental biotechnology, particularly for the detection of different toxic inorganic heavy metals including arsenic, lead, zinc, copper. It can also be used for the detection of the pathogenicity of microbes before their application in environmental aspects. Microbial quorum-sensing techniques can easily be exploited for the removal of several organic as well as inorganic compounds.

6.3.1 Heavy metal detection QS is a synergistic approach for communication between microorganisms to conduct an activity in a group. This interaction between the communicating molecules of the complex helps in the development of various biosensors. Biosensing activity in microbial biosensors is

Bacterial quorum sensing in environmental biotechnology 155

Acyl homoserine lactone Cell surface

Luxl

LuxR

LuxR

Target gene

Figure 6.2 The figure shows the mechanism of quorum sensing (QS) in Gram-negative bacteria. Acyl homoserine lactone (AHL) is secreted and synthesized by LuxI and it is released to the extracellular environment. When it reaches the threshold concentration it is recognized by LuxR which then controls the gene expression.

mainly initiated by a wide range of microorganisms capable of detecting the bioavailability of different toxic chemical species. The basic principle of such biosensors relies on the expression of the analyte-induced reporter gene, which may express fluorescence, bioluminescence, or simple colorimetric indications. When the concentration of any toxic chemical species increases beyond a certain measure, the AI gene gets expressed producing visible signals (Wang, Barahona, & Buck, 2013). The bioelectrochemical systems, which generate electric power utilizing the chemical energy possessed in organic wastes, have been effectively used for producing signals against the unfavorable state of environmental factors, such as biological oxygen demand, limited presence of carbon, etc. For example, the bioelectrochemical system-based biosensing principle for the detection of heavy metals like arsenic involves the generation of a direct electric signal output by the reduction of the anode of the electrochemical cell (Friedman, TerAvest, Venkataraman, & Angenent, 2012). The biosensors function based on the metal reduction pathway of Shewanella oneidensis which includes a plasmid-encoded copy of its mutant type mtrB along with a promoter gene which is induced by arsenic. In the presence of arsenic inoculated with S. oneidensis, the mtrB reduces Fe(III) to Fe(II) in the negative electrode thus producing an electric pulse (Webster et al., 2014). Some of the well-known QS systems that have been reported for the detection of heavy metals have been listed in Table 6.1.

156 Chapter 6 Table 6.1: Different microbial quorum-sensing (QS) systems for heavy metal detection. Sl. No. QS system 1.

2.

3.

Metal detected Basic principle

Arsenic specific Metal reduction pathway of S. oneidensis is under the control of an arsenicsensitive promoter Zinc and Bacterial regulatory system is modified copper ZntA-ZntR and CueO, which encodes zinc and copper efflux and the proper quorum system is interrupted Vibrio harveyi Lead This strain is modified to make genetically modified defective AI1 and AI2, which results in lead precipitation Genetically engineered Shewanella oneidensis Escherichia coli strain XL1-Blue

4.

Actinobacillus Iron actinomycetemcomitans

5.

Acinetobacter junii BB1A

6.

Burkholderia multivorans

7.

Whole cell Clorella vulgaris

8.

Pseudomonas aeruginosa

9.

Pseudomonas aeruginosa

10. Sinorhizobium meliloti

11. Genetically engineered Deinococcus radiodurans

Nickel, arsenic dioxide, cadmium, mercury Nickel and cadmium

References Webster et al. (2014) Ravikumar, Ganesh, Yoo, and Hong (2012)

Mire, Tourjee, O’Brien, Ramanujachary, and Hecht (2004) In iron-limited conditions the growth of Fong, Gao, and cell density is high when there is a Demuth (2003) plasmid containing a functional copy of luxS It is a novel metal-tolerant bacterium Sarkar and which produces a biofilm in the Chakraborty (2008) presence of metal Sub millimolar concentration of Nickel and cadmium inhibits biofilm formation by inhibiting acylhomoserine lactone quorum sensing (QS) Conductometric biosensors with gold planar interdigitated electrodes

Vega et al. (2014)

This has a lasR/lasI QS system which is activated by transcriptional regulator PA4778 which is copper resistant Iron, vanadium Vandium has antibacterial activity against P. aeruginosa in conditions of iron limitation Mercury, EPS production helps to pump out arsenic metal from cell back into the culture medium, so they are resistant, and results in formation of biofilm Cadmium The sensor plasmid (PRADI-P0659-1) containing crtl as a reporter gene. under the control of P0659-1 which is regulated by cadmium which is introduced into a crtl deleted mutant strain of D. radiodurans

Thaden, Lory, and Gardner (2010)

Cadmium, nickel, cobalt, lead, and zinc Copper

Berezhetskyy et al. (2007)

Aendekerk, Ghysels, Cornelis, and Baysse (2002) Nocelli, Bogino, Banchio, and Giordano (2016) Joe et al. (2012)

(Continued)

Bacterial quorum sensing in environmental biotechnology 157 Table 6.1: (Continued) Sl. No. QS system

Metal detected Basic principle

References

12. Acidithio bacillus

Copper

Wenbin et al. (2011)

13. Pseudomonas fluorescens

Cadmium

14. Cupriavidus metallidarans

Lead

When (5Z)-4-bromo-5(bromomethylene)-2(5H)-furanone (FUR) compound is absent in the culture, the microbe is resistant to copper but when the concentration of FUR compound is more than 0.05 µg/ mL then the microbe shows sensitivity to copper Shows differential gene expression in the presence of the toxic metal cadmium Main mechanisms of lead resistance are extracellular polysaccharides adsorption and ion efflux into the cell exterior

Rossbach et al. (2000) Jarosławiecka and Piotrowska-Seget (2014)

Bacterial biosensors have also been found to be successful for the detection of metals like zinc and copper. The detection system is based on the activation of gene regulators like zntA-zntR and CueO that are present in the bacterial cell cytoplasm. The efflux proteins thus produced in the limited presence of zinc and copper detoxify them. However, in the presence of higher ionic concentration of the metals, the efflux proteins are not generated, following which the presence of an exogenous toxic metal is not recognized (Ravikumar, Yoo, Lee, & Hong, 2011). The bacterial two-component system including ZraSR and CusSR, on the other hand, can recognize exogenous metals, along with an initiated response in a stimuli-response coupled system to changes even at higher concentrations in the neighborhood of the bacteria (Ravikumar et al., 2012). Vibrio harveyi, on the other hand, has the AI1 AI2 system that can block lead precipitation in its surrounding environment. Three pleiotropic quorum-sensing defects can be introduced in the strain through mutagenic modifications due to which the strain acquires the ability to convert soluble lead matter into insoluble lead precipitate in the form of lead phosphate salt which can then be quantified to obtain the amount of lead present in the surrounding environment (Mire et al., 2004). The LasR/LasI QS system in P. aeruginosa exhibits gene expression and pathogenesis. The QS system is activated by transcription regulator PA4778 which is a copper-resistant system. By genetically modifying the PA4778 regulator, P. aeruginosa can be converted to a copper-sensitive microbe and by the increase in the concentration of copper the growth rate of P. aeruginosa will decrease quantitatively (Thaden et al., 2010).

158 Chapter 6

6.3.2 Pathogen detection Pathogenicity or virulence can be dependent on QS in two opposite ways. In some cases, for example, Vibrio cholera, QS helps in the disposal of bacteria within the host body by suppressing biofilm formation (Bridges & Bassler, 2019). However, in Pseudomonas sp. QS supports the pathogenicity by forming a biofilm with an increase in cell density (Bhushan et al., 2013; Kalia, 2013). In other cases, with an increase of cell density the virulence gene is activated due to the increasing concentration of autoinducers (Defoirdt, 2018). In both cases, QS can be targeted to restrict the spreading of the pathogen, restricting its protective barriers or inhibiting their virulent character at the gene expression level. For example, niclosamide is an FDA-approved drug to treat tapeworm infection. It was observed that it suppresses the biofilm formation of Xanthomonas in rice, and hence restricts its pathogenicity (Sahu, Zheng, & Yao, 2018). A similar kind of compound thymol was observed in which virulence gene expression can be restricted and also biofilm formation occurred in different species (Singh, Gupta, Tandon, & Pandey, 2017). These approaches of inactivating QS or quorum quenching can be very useful to enable a wide range of bacterial strains with phenomenal biodegrading or bioremediation applications which are now not in use as they have objectionable pathogenic activity toward the animal or plant kingdom. Examples of such kinds of bacterial strains include different strains of Pseudomonas, Mycobacterium, and Xanthonomas. Apart from these, another most favorable application of this aspect can be found in food preservation, where food spoilage and the spreading of human pathogens can be restricted using quorum quenching (Morris & Monier, 2003).

6.3.3 Bioremediation To eliminate recalcitrant pollutants, bioremediation is considered to be the best of the existing remediation strategies, as it is the most ecofriendly, economic, and socially acceptable. A wide range of bacteria has been identified along with their bioremediation abilities for particular pollutants and most of them have a QS influence in their degradation pathway. Burkholderia sp. DW2-1, which has the ability to degrade pesticides present in soil as well as wastes present in water, has a QS influence on their degradation metabolism. CepI and CepR are two QS molecules involved in biosurfactant production in glucose nitrate supplemented media, enhancing their degradation metabolism (Matsumiya, Wakita, Kimura, Sanpa, & Kubo, 2007). Similar reports have been made regarding B. cenocepacia BSPW and its quorum sensing-dependent degradation ability (Wattanaphon, Kerdsin, Thammacharoen, Sangvanich, & Vangnai, 2008). Similarly, microbes collected from aerobic granular sludge (AGS) with applications in wastewater treatment were reported to exhibit QS-dependent bioremediation. In this case, AI2 was found to control

Bacterial quorum sensing in environmental biotechnology 159 their bioremediation ability in a boron-supplemented medium which resulted in the rapid formation of exopolysaccharides and faster AGS formation (Zhang, Yu, Guo, & Wu, 2011). P. aeruginosa has been reported with a wide range of degradation ability including anthranilate, pharmaceutical, cosmetic, food industry, and oil contamination. Each degradation capability has a direct relation with QS depending on different inducer molecules. For example, anthranilate biodegradation depends on the autoinducer named N-decanoyl-L-HSL (C10HSL) in another case Pseudomonas quinolone signal controls the QS-mediated bioremediation by transcription-dependent iron chelation in P. aeruginosa (Chugani & Greenberg, 2010; Mu¨ller et al., 2012; Schertzer, Boulette, & Whiteley, 2009). Enhancing or providing a supplement of these quorum sensor molecules in the operation site, along with the respective degrading bacteria, will result in better, enhanced, faster, and more effective bioremediation.

6.3.4 Biofilm formation Biofilm is a sophisticated multicellular structure formed by the autoaggregation of the same species or multispecies bacteria under different environmental stimuli which includes nutrient availability or stressed conditions like osmotic stress, desiccation, ultra-violet radiation, or even pH change (Braeken, Daniels, Ndayizeye, Vanderleyden, & Michiels, 2008; Branda, Vik, Friedman, & Kolter, 2005; Morris & Monier, 2003). The major component of biofilm is formed by exopolysaccharide and the minor component includes protein, lipid, and even nucleic acid. In general, biofilm is negatively charged and contains acidic moieties which include uronic acid. Its negative charge is contributed by the abundance of sulfate and phosphate (Bogino, Oliva Mde, Sorroche, & Giordano, 2013). Biofilm formation is often considered as an expression of communication between individual bacterial strains of the same species or different species via QS. In most cases biofilm formation was observed to be population density-dependent signifying the contribution of QS molecules in its formation (Niu et al., 2008). Biofilm formation can be useful for the bioremediation of different effluents using a bioreactor. The principle hurdle in bioremediation using bioreactor digestion is maintaining considerable bacterial growth inside the bioreactor as well as minimizing the bacterial discharge from the bioreactor (Kalia, Prakash, Ray, & Koul, 2018). Among the different entrapment methods, biofilm can be considered as the best as it contains dominant discrete channels inside the three-dimensional structure, enabling proper nutrient water and metabolite passage (Mohan et al., 2018). It also helps to improve a low food to cell concentration ratio which supports degradation. QS has great implications for this approach (Kang & Park, 2010a; Sharma & Lal, 2017). In an alternative mode of the bioremediation of wastewater, a membrane bioreactor faces a bottleneck problem of plugging of the membrane (membrane biofouling) which is mainly caused by biofilm formation on the membrane (Monclu´s Sales et al., 2015). To overcome

160 Chapter 6 this problem, an approach that has been suggested is called “quantum quenching” (Meng et al., 2017). It is a process in which the quantum sensing is blocked by supplementing extrinsic inhibitor molecules to block AI, hence inhibiting biofilm formation, and thus remediating biofouling on the membrane bioreactor (Lee, Yu, Zhang, & Choo, 2018). To block AI in QS for biofilm formation, different quorum quenching agents are used. Among these enzymes, lactonase enzyme, acylase, and oxidoreductase are well proposed in different literature (Dong, Xu, Li, & Zhang, 2000; Lin et al., 2003; Uroz et al., 2005). Due to the instability and high cost of these quorum-quenching enzymes, alternatives of quantum-quenching bacterial strains have also been found which possess the ability to block quorum-sensing AI. Among these, Pseudomonas sp. 1A1, Rhodococcus sp. BH4, and Bacillus methylotrophicus sp. WY are well reported to remediate biofouling on membranes by inhibiting biofilm formation (Cheong et al., 2013; Oh et al., 2012).

6.3.5 Hydrocarbon remediation Hydrocarbon contamination is a current challenge as these pollutants have a wide range of environmental effects, such as decreasing soluble nutrient, nitrogen, and salt concentrations in soil, as well as decreasing DO in water, quite apart from their cytotoxic activity. Hydrocarbon contaminants is a collective term used for a wide range of long-chain organic chemicals which include phenolic compounds, alkanes, anthranilate, naphthalene, and benzene derivatives. Different microbes have been identified with the degradation capability of different hydrocarbons. The common feature they share is the ability to form biosurfactant and biofilms (Bhargava, Sharma, & Capalash, 2012; Skeels & Whitby, 2018). Microbial consortia made up of Enterobacter sp., Klebsiella spp., and Pantoea sp. have been reported with mixed hydrocarbon degradation ability as well as biofilm formation. Since biofilm formation is directly dependent on intercellular communication via QS, the biodegradation also may have been dependent on QS (Poddar, Sarkar, & Sarkar, 2019). Apart from these, a negative QS has been reported for Pseudomonas sp. As1 in which an oxidative AI is generated during the degradation process restricting cell growth and the degradation process as well. Quorum quenching of this oxidative AI helps to enhance the degradation process (Kang et al., 2007). P. aeruginosa CGMCC 1.860 has been reported to have an aromatic degradation capability along with AHL production. It was reported that the extrinsic addition of AHL extract enhanced the pollutant biodegradation (Yong & Zhong, 2013). In contrast, anthranilate biodegradation by the same strain is indirectly regulated by AHL (Chugani & Greenberg, 2010). In Acinetobacter sp. DR1, the QSdependent hydrocarbon degradation is evident. A mutant strain that has lost the ability to produce AHL has shown defective mineralization and growth in the presence of hexadecane. On the other hand, the wild strain supplemented with AHL shown improved hexadecane mineralization and growth as well, stating the role of QS and biofilm formation in hydrocarbon degradation (Kang & Park, 2010a, 2010b).

Bacterial quorum sensing in environmental biotechnology 161

6.4 Limitations of microbial quorum sensing As the fundamental principle of QS depends on the bacterial cellular density, QS has some notable limitations in practical real-life use. For example, to enable the QS effect requires a threshold concentration of AI elements whose concentration is directly related to the biomass concentration. If the real-life environment is not supportive for cell growth, then QS effects will not be evident as cellular growth requires significant time and in the practical environment it may be quite slow. Therefore the time required to realize the effect of QS is quite long. Hence the time required for bioremediation implemented by QS must require a larger time period. Similarly, if the environment contains any AI blocker in its vicinity, the QS may be ineffective due to the blocking effect of that contaminant. Hence the location of the implication of the QS-mediated bioremediation has to be well analyzed and characterized. Thus the practical application of QS is best realized in a controlled environment and for its practical application in real time the success is heavily dependent on the environment of the application.

6.5 Conclusion QS, the communication medium between single-celled bacterial communities, plays a vital role in the well-being and survival of bacteria. As most of the extraordinary functions of different bacterial strains, such as biofilm formation, biodegradation, cell disposals, survival in stressed environment, mutualism, commensalism, etc., depend on QS, it can be exploited for human benefits which include pollutant detection, bioremediation, biofilm formation or restriction, and biocontrol of pathogenicity. In spite of extraordinary reports regarding QS, the practical application of QS has significant limitations which make it best for controlled use only. However, there are remarkable opportunities for its use as a biosensor and in bioremediation.

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164 Chapter 6 Ravikumar, S., Yoo, I. K., Lee, S. ,Y., & Hong, S. H. (2011). A study on the dynamics of the zraP gene expression profile and its application to the construction of zinc adsorption bacteria. Bioprocess and Biosystems Engineering, 34, 1119 1126. Rossbach, S., Kukuk, M. L., Wilson, T. L., Feng, S. F., Pearson, M. M., & Fisher, M. A. (2000). Cadmiumregulated gene fusions in Pseudomonas fluorescens. Environmental Microbiology, 2, 373 382. Sahu, S. K., Zheng, P., & Yao, N. (2018). Niclosamide blocks rice leaf blight by inhibiting biofilm formation of Xanthomonas oryzae. Frontiers in Plant Science, 9, 408. Sarkar, S., & Chakraborty, R. (2008). Quorum sensing in metal tolerance of Acinetobacter junii BB1A is associated with biofilm production. FEMS Microbiology Letters, 282(2), 160 165. Schertzer, J. W., Boulette, M. L., & Whiteley, M. (2009). More than a signal: Non-signaling properties of quorum sensing molecules. Trends in Microbiology, 17, 189 195. Sharma, A., & Lal, R. (2017). Survey of (Meta)genomic approaches for understanding microbial community dynamics. Indian Journal of Microbiology, 57, 23 38. Singh, A., Gupta, R., Tandon, S., & Pandey, R. (2017). Thyme oil reduces biofilm formation and impairs virulence of Xanthomonas oryzae. Frontiers in Microbiology, 8, 1074. Skeels, K., & Whitby, C. (2018). Microbial ecology of naphthenic acid (NA) degradation. Microbial communities utilizing hydrocarbons and lipids: Members, metagenomics and ecophysiology (pp. 1 22). Cham: Springer. Thaden, J. T., Lory, S., & Gardner, T. S. (2010). Quorum-sensing regulation of a copper toxicity system in Pseudomonas aeruginosa. Journal of Bacteriology, 192, 2557 2568. Uroz, S., Chhabra, S. R., Camara, M., Williams, P., Oger, P., & Dessaux, Y. (2005). N-Acylhomoserine lactone quorum-sensing molecules are modified and degraded by Rhodococcus erythropolis W2 by both amidolytic and novel oxidoreductase activities. Microbiology, 151, 3313 3322. Vega, L. M., Mathieu, J., Yang, Y., Pyle, B. H., McLean, R. J., & Alvarez, P. J. (2014). Nickel and cadmium ions inhibit quorum sensing and biofilm formation without affecting viability in Burkholderia multivorans. International Biodeterioration and Biodegradation, 91, 82 87. Wang, B., Barahona, M., & Buck, M. (2013). A modular cell-based biosensor using engineered genetic logic circuits to detect and integrate multiple environmental signals. Biosensors and Bioelectronics, 40, 368 376. Wattanaphon, H. T., Kerdsin, A., Thammacharoen, C., Sangvanich, P., & Vangnai, A. S. (2008). A biosurfactant from Burkholderia cenocepacia BSP3 and its enhancement of pesticide solubilization. Journal of Applied Microbiology. Available from https://doi.org/10.1111/j.1365-2672.2008.03755.x. Webster, D. P., TerAvest, M. A., Doud, D. F., Chakravorty, A., Holmes, E. C., Radens, C. M., . . . Angenent, L. T. (2014). An arsenic-specific biosensor with genetically engineered Shewanella oneidensis in a bioelectrochemical system. Biosensors and Bioelectronics, 62, 320 324. Wenbin, N., Dejuan, Z., Feifan, L., Lei, Y., Peng, C., Xiaoxuan, Y., & Hongyu, L. (2011). Quorum-sensing system in Acidithiobacillus ferrooxidans involved in its resistance to Cu2 1 . Letters in Applied Microbiology, 53, 84 91. Williams, P. (2007). Quorum sensing, communication and cross-kingdom signalling in the bacterial world. Microbiology, 153, 3923 3938. Williams, P., & Ca´mara, M. (2009). Quorum sensing and environmental adaptation in Pseudomonas aeruginosa: A tale of regulatory networks and multifunctional signal molecules. Current Opinion in Microbiology, 12, 182 191. Yong, Y. C., & Zhong, J. J. (2013). Regulation of aromatics biodegradation by rhl quorum sensing system through induction of catechol meta-cleavage pathway. Bioresource Technology, 136, 761 765. Zhang, S. H., Yu, X., Guo, F., & Wu, Z. Y. (2011). Effect of interspecies quorum sensing on the formation of aerobic granular sludge. Water Science and Technology, 64, 1284 1290.

CHAPTER 7

Bioremediation: an effective technology toward a sustainable environment via the remediation of emerging environmental pollutants Komal Agrawal1, Ankita Bhatt1, Venkatesh Chaturvedi2 and Pradeep Verma1 1

Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Bandarsindri, Kishangarh, Ajmer, India, 2SMW College, MG Kashi Vidyapeeth, Varanasi, India

7.1 Introduction There is an emerging group of pollutants that have not been studied and pose a threat to the environment. These pollutants are not considered under the “water-quality regulations” (La Farre, Pe´rez, Kantiani, & Barcelo´, 2008) but may be potential candidates for future regulations. The emerging pollutants are released from various industries and pose great risk to the existing life on land and in aquatic bodies and these wastes include bisphenol A (BPA), polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), pharmaceutical wastes, hospital effluents, and disinfectant by-products. The emerging pollutants are new compounds which have been generated from various sources and the effect/mechanism of action of these toxic compounds varies depending on the chemical composition and source of production. The emerging pollutants are unregulated compounds and depending upon their occurrence, along with potential health effects, they can be segregated and the mechanism of treatment can be determined (Verlicchi, Galletti, Petrovic, & Barcelo´, 2010). These pollutants can be treated by various physical chemical and biological methods, for example, microbial, fungal, mix culture, phyto- and phycoremediation, enzymatic treatment, zoo- and vermiremediation. However, biological methods are preferred over the other methods as they are more cost-effective and ecofriendly. These treatments can be either used alone or in combination with various technologies for the effective removal of the pollutants from the environment. Apart from these techniques there are numerous other recent technologies, for example, immobilization, electrokinetic Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00007-9 © 2020 Elsevier Inc. All rights reserved.

165

166 Chapter 7

Figure 7.1 The types of emerging pollutants existing in the environment.

remediation, metagenomics, protein engineering, bioinformatics, which have been effectively used for the bioremediation of various pollutants. Therefore to considering the above-mentioned points the present chapter deals with various kinds of emerging pollutants that are present in nature, along with various existing bioremediation techniques and the latest technologies which have gained interest amongst researchers for treating these pollutants.

7.2 Emerging pollutants The various types of emerging pollutants (Fig. 7.1), their sources, and their effects are represented in Table 7.1 and are discussed below.

7.2.1 Bisphenol A BPA belongs to the class of endocrine disrupting chemicals (EDC) and is frequently detected in wastewaters (Flint, Markle, Thompson, & Wallace, 2012; Rathnayake, Xi, Frost, & Ayoko, 2016). The EDCs have become the prime cause of concern because they have various harmful effects on mankind and the ecosystem. Owing to the structural similarity between BPA and the hormone estradiol, BPA can bind and activate the estrogen receptor just like the natural hormone at lower concentrations (Rochester, 2013). These compounds affect the growth, development, behavior, and reproduction of animals and humans. Two phenolic groups constitute the structure of this organic compound. BPA is

Bioremediation: an effective technology 167 Table 7.1: Sources and effects of various emerging pollutants. Pollutant

Sources

Effects

Bisphenol A (BPA)

Cans of packed food items, bottle tops, water supply pipes, baby bottles, household electronics, and plastic lenses Wood burning, wildfires, tobacco smoking, biofuel combustion, vehicular emissions, and oil spills

Endocrine disruptor, reproductive Huang et al. damage, and poor development in (2012) aquatic organisms

Polycyclic aromatic hydrocarbons (PAHs) Polychlorinated biphenyls (PCBs)

Landfills containing capacitors, building materials, and incineration of municipal wastes

Cardiovascular disease, cancer of vital organs, and poor fetal development

Acute systemic poisoning, reduced thyroid hormone, mutagenic, liver and stomach injuries, immune system effects, and behavioral changes Cancers, birth defects, bladder Disinfection by- Formed as a result of reactions products with naturally present amino acids cancer, and miscarriages and other organic matter Pharmaceutical Animal husbandry, sewage Reproductive damage, behavioral wastes effluent, aquaculture, landfill alterations, bioaccumulation, and leachate, and horticulture biomagnification

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Abdel-Shafy and Mansour (2016) Alcock, Behnisch, Jones, and Hagenmaier (1998) Krasner (2009)

Gaw, Thomas, and Hutchinson (2014)

produced at a large scale for its widespread use in the production of epoxy resins and polycarbonate plastics (Ballesteros-Go´mez, Rubio, & Pe´rez-Bendito, 2009). BPA has a complex aromatic structure and low biodegradability which makes it a recalcitrant water pollutant (Kang, Kondo, & Katayama, 2006), leading to its detection in the wastewater treatment plant effluents (Park, Sun, Ayoko, & Frost, 2014). There are various methods available for the removal of BPA from water such as biodegradation (Mohapatra, Brar, Tyagi, & Surampalli, 2010), adsorption (Zhao, Fu, & Ma, 2014), photodegradation (Chen, Zhang, & Zuo, 2013), nanofiltration (Braeken & Van der Bruggen, 2009), chemical oxidation (Xi, Sun, Hreid, Ayoko, & Frost, 2014), and ozonation (Umar, Roddick, Fan, & Aziz, 2013). The other methods which have been used for the BPA removals are as follows: 7.2.1.1 Inorganic organic clays BPA removal has achieved efficient results with the utilization of organoclays (Zheng, Sun, Park, Ayoko, & Frost, 2013). This involves the replacement of interlayer cations with cationic surfactants, resulting in a transition from a hydrophilic to hydrophobic surface, thereby changing the surface properties of the clays. Inorganic organic clays (IOCs) are modified clay minerals obtained by intercalating both organic and inorganic modifiers like hydroxy aluminum cation and quaternary alkyl ammonium cations (Srinivasan & Fogler, 1990). A variety of organic and inorganic contaminants can be simultaneously absorbed by the IOCs. They are effective in the removal of phenol, and its nitro- and chloroderivatives.

168 Chapter 7 7.2.1.2 Photodegradation The photocatalytic method of BPA removal is gaining more importance within the other classical methods for BPA degradation. The ecofriendly method relies on the utilization of cheap sources of energy and has only one limiting factor, namely the absorbing capacity of catalysts (Cojocaru et al., 2017). Various photocatalysts are capable of efficient BPA removal including Ti/TiO2 (Frontistis, Daskalaki, Katsaounis, Poulios, & Mantzavinos, 2011), photo-Fenton systems (Yang et al., 2013), N TiO2 (Yang, Dai, & Li, 2011), Ag TiO2 (Rengaraj & Li, 2006), and Pt TiO2 (Chiang, Lim, Tsen, & Lee, 2004). The use of gold particles as catalysts has also achieved promising results (Iliev, Tomova, Bilyarska, & Tyuliev, 2007).

7.2.2 Polycyclic aromatic hydrocarbons PAHs are colorless, pale yellow or white colored organic solids which comprise only carbon and hydrogen atoms. PAHs comprise single or fused aromatic rings sharing a pair of carbon atoms between the rings (Abdel-Shafy & Mansour, 2016). They are present in a mixture of two or more of these compounds and are released into the environment by a variety of routes. PAHs are ubiquitous and are found mostly in air, water, and soil (Baklanov et al., 2007). They are produced during biological processes or from natural (forest fires) and anthropogenic combustion sources (vehicular emissions, cigarette smoking). The incomplete combustion of inorganic material like coal, wood, and oil is the major source of PAHs. They are highly toxic, potent immune-suppressants, carcinogenic, and disrupt the functions of cellular membranes and associated enzymes systems (Armstrong, Hutchinson, Unwin, & Fletcher, 2004). Depending upon the number of aromatic rings present, PAHs are classified into small or large PAHs. The molecules with six or fewer fused aromatic rings are called small PAHs and those with more than six rings fall into the category of large PAHs. The commercial uses of PAHs include intermediates in pharmaceuticals, photographic products, agricultural products, thermosetting plastics, and other chemical industries. Biodegradation, chemical oxidation, and photooxidation are the various mechanisms for the degradation of PAHs in the environment (Nadarajah, Van Hamme, Pannu, Singh, & Ward, 2002). 7.2.2.1 Bacterial catabolism of polycyclic aromatic hydrocarbons Bacteria are mostly known to degrade certain PAHs like phenanthrene and naphthalene. Depending upon the presence or absence of oxygen, there are two main mechanisms for the degradation of PAHs, namely aerobic and anaerobic. The bacterial genus Rhodococcus is very unique and possesses an enormous ability to catabolize versatile PAHs. This grampositive bacterium comprises only three structural genes responsible for the degradation of naphthalene (Kulakov, Chen, Allen, & Larkin, 2005; Larkin, Kulakov, & Allen, 2005).

Bioremediation: an effective technology 169 7.2.2.2 Halophilic/halotolerant bacteria and archaea for the degradation of polycyclic aromatic hydrocarbons PAHs are the prime cause of pollution in hypersaline habitats; they are released into the environment in the effluents of industrial and municipal sectors, with the petroleum industry being the major source of aromatic hydrocarbons (Fathepure, 2014). There are six major bacterial/archaeal groups that are known to possess PAHs-degrading ability, namely Alphaproteobacteria, Betaproteobacteria, Gammaproteobacteria, Actinomycetes, Firmicutes, and Archaea (Halophiles) (Ghosal, Ghosh, Dutta, & Ahn, 2016). 7.2.2.3 Fungal degradation The degradation of PAHs by fungi is solely dependent on monooxygenase enzymes (Cerniglia & Sutherland, 2010). The fungi involved are categorized into two main groups, namely ligninolytic fungi and nonligninolytic fungi. The former are also called white-rot fungi and their action involves the activity of enzymes such as lignin peroxidase, laccases, and manganese peroxidase for lignin degradation, whereas nonligninolytic fungi employ cytochrome P450 monooxygenase-like enzymes (Hofrichter, 2002; Tortella, Diez, & Dura´n, 2005). Another type of ligninolytic fungi, that is, brown-rot fungi, depends upon the use of hydrogen peroxide for the degradation of cellulose and hemicelluloses (Ghosal et al., 2016). 7.2.2.4 Chemical oxidation and photodegradation In this method, the PAHs oxidation rate depends upon the structure of the compound, the physical state, the molecular weight of compound, the temperature, and the nature of oxidizing agent. This method is mostly employed for the efficient removal of PAHs from surface water (Moursy & Abdel-Shafy, 1983). Photodegradation consists of the degradation of target pollutants with the help of reactions initiated by light absorption. The absorption of light by PAHs results in electronic activation within the molecules that further leads to the production of an unstable structural arrangement, allowing further various physical and chemical processes to act on the compound (Schwarzenbach & Gschwend, 2016).

7.2.3 Polychlorinated biphenyls PCBs are halogenated aromatics and are present worldwide in different environments. Certain properties like inflammability, lipophilicity, and chemical stability are responsible for their industrial applications and also contribute to their environmental problems. After entering into the environment, PCBs become very stable, degrade slowly, and undergo further cycling within the ecosystem (Safe, 1994). They also have the ability to bioaccumulate and biomagnify in higher trophic levels of the food chain (McFarland & Clarke, 1989). The composition of environmental and commercial PCBs differs significantly due to various physical and biological processes (metabolism, degradation).

170 Chapter 7 The individual components of these mixtures act in a synergistic or antagonistic method among themselves and along with other classes of pollutants, further impacting the environment and ecosystem at a larger scale. There are several possible routes for human exposure to these PCBs mixtures; namely the exposure of workers who are involved in PCBs production or utilize PCBs-containing products, accidental and environmental exposure via contaminated air, food, and water. These recalcitrant environmental pollutants have hazardous impacts, such as dermal toxicity, carcinogenicity, immunotoxicity, effects on reproduction or endocrine function, and other toxic effects (Safe, 1994). 7.2.3.1 Biological degradation There are three processes for the reduction of PCBs toxicity, namely physical, chemical, and biological degradation (Duda´sˇova´ et al., 2016). Under specific conditions, PCBs undergo partial degradation by the aforementioned three mechanisms (Perelo, 2010). Various biological agents are now widely employed for PCBs remediation to achieve the transformation of persistent organic pollutants (POPs) to simpler and less toxic compounds (Duda´sˇova´ et al., 2016). Many microbial strains are isolated and characterized for their ability to remediate PCBs by aerobic or anaerobic processes of degradation (Koubek, Mackova, Macek, & Uhlik, 2013). 7.2.3.2 Halogenated organic compounds technology for polychlorinated biphenyls degradation The employment of zero-valent iron particles for the remediation of halogenated organic compounds (HOCs) has emerged as an innovative technology for PCBs degradation (Aral & Guan, 1996). Many HOCs, like chlorinated aromatics, chlorinated aliphatics, and PCBs, have been efficiently degraded by granular iron particles (Chuang, Larson, & Wessman, 1995). This technology can also be easily and efficiently applied in the field (Wang & Zhang, 1997). Various other metals, like zinc (Zn), palladium (Pd), and tin (Sn), also serve as efficient agents for the remediation of HOCs, and of these the best results have been obtained with the use of palladium (Boronina, Klabunde, & Sergeev, 1995).

7.2.4 Pharmaceutical wastes The word “pharmaceutical” includes a variety of compounds with different structures, behaviors, functions, and activity (Jones, Voulvoulis, & Lester, 2005) and large amounts of toxic contaminants are released into the environment by various pharma industries (Patneedi & Prasadu, 2015). These pharmaceutical compounds are employed to fight infections or cure diseases in humans and animals. They are incompletely metabolized in the body and thus enter the sewer system which feeds into the aquatic environment (Tixier, Singer, Oellers, & Mu¨ller, 2003; Zuccato, Calamari, Natangelo, & Fanelli, 2000). They are also released as runoff from animal wastes, seepage from sewer lines (Glassmeyer et al., 2005), hospital effluents, and biosolids (Daughton & Ternes, 1999). Pharmaceuticals also

Bioremediation: an effective technology 171 consist of nonsteroidal antiinflammatory drugs, such as naproxen, diclofenac, and ibuprofen, that are most frequently present in sewage and groundwater systems, whereas the surface water system contains drugs like ibuprofen, acetylsalicylic acid, ketoprofen, diclofenac, naproxen, and indomethacin. Diclofenac has a high toxicity toward cattle and vultures and long-time exposure to trace amounts of diclofenac leads to drastic impacts on aquatic life and human health (Patneedi & Prasadu, 2015). The long-term exposure of complex pharmaceuticals also affects the stream biota resulting in various acute and chronic damage (Quinn, Gagne´, & Blaise, 2008), such as inhibition of cell proliferation (Pomati et al., 2006), damage of the reproductive system (Nentwig, 2007), behavioral changes (Gaworecki & Klaine, 2008), and bioaccumulation in the tissues (Brooks et al., 2003). Two hundred of the most widely employed pharmaceuticals products were studied in 2006 and six of them were found to belong to the category of antidepressants (La Farre et al., 2008). Since there are few methods are available to detect the presence of antidepressants, little information about their occurrence and fate in environmental matrices is available. Thus various treatment methods have been employed for the removal of pharmaceutical wastes, which are as follows: 7.2.4.1 Photodegradation Some of the drugs are sensitive to sunlight and are effectively degraded, for example, diclofenac is degraded effectively in water bodies due to the action of ultraviolet (UV) light. Other drugs like naftifine, sulbentine, triclosan, cloxiquin, tolnaftate, and chlorphenesin are also sensitive to light. Thus exposure to light can be implemented for the effective removal of various light-sensitive drugs (Derksen, Rijs, & Jongbloed, 2004; Thoma, Kubler, & Reimann, 1997). 7.2.4.2 Sludge treatment Pharmaceutical waste is sensitive to heat and might be degraded in the sewage treatment plants during the composting due to heat, for example, probenecid and methaqualone. Certain waste is degraded during the chemical treatment step as well as being biodegraded (Guerin, 2001; Ternes, 2001).

7.2.5 Hospital effluents as source of emerging pollutants Hospital effluents are a source of various pollutants like pharmaceuticals, surfactants, personal care products (PCPs), illicit drugs, endocrine disruptors, radionuclides, solvents, disinfectants, and gasoline additives. Hospitals are an important source of metabolites and medicaments which are excreted unchanged, mostly in urine and partially in feces (Lienert, Bu¨rki, & Escher, 2007). Similarly, large amounts of contaminants, for example, chemicals, disinfectants, heavy metals, detergents for endoscopes, radioactive markers, drugs and their

172 Chapter 7 metabolites, are also released from laboratory, diagnostic, and research activities. The most widely employed substances in hospitals are disinfectants which are defined as complex products and mixtures of alcohol or aldehydes or chlorine-containing compounds (recalcitrant chlorophenol). They are used for the disinfection of instruments and skin, surface disinfection, in food processing, and in glue and size production. Hospital wastewaters are also a source of various heavy metals. Platinum is found in wastewater as a result of excretion by oncological patients who are treated with carboplatinum or cisplatinum. Diagnostic agents usually contain mercury and gadolinium is also found as a consequence of its use in magnetic resonance imaging where it is employed owing to its high magnetic moment (Verlicchi et al., 2010). These compounds are excreted unchanged quickly after their administration in patients, for example, iodinated X-ray contrast media— derived from X-ray examinations and other radiological processes—exhibits high biochemical stability and more than 90% is thus excreted in unmetabolized forms (Ternes & Hirsch, 2000). 7.2.5.1 Biological treatment Various studies have been conducted on the potential to utilize ultrafiltration membrane biological reactors and conventional activated sludge processes for the treatment of emerging pollutants (Radjenovi´c, Petrovi´c, & Barcelo´, 2009). The nitrifying bacteria and sludge retention time (SRT) play an important role in the biological treatment technique. It was found that higher removal efficiencies were achieved at longer SRTs as this allowed for the successful establishment of even the slowest growing bacteria. In addition, it enhanced the metabolic and cometabolic processes, thereby affecting recalcitrant compounds and finally leading to complete mineralization (Oppenheimer, Stephenson, Burbano, & Liu, 2007). Also the removal efficiency was found to be independent of the structure of the target pollutant. Further, Batt, Kim, and Aga (2006) and Marttinen, Kettunen, and Rintala (2003) demonstrated that nitrifying bacteria play a prime role in pharmaceutical biodegradation processes that employ higher SRTs (Verlicchi et al., 2010).

7.2.6 Other emerging pollutants Emerging pollutants are a diverse group of compounds that includes drugs, PCPs, steroids, hormones, perfluorinated compounds (PFCs), surfactants and pharmaceuticals, flame retardant, gasoline additives, industrial additives, as well as their transformation products (La Farre et al., 2008). Further, nanomaterials, swimming pool disinfection by-products, and 1,4-dioxane have also been added to the category of emerging pollutants (Richardson, Plewa, Wagner, Schoeny, & DeMarini, 2007). These pollutants are transported and distributed in the environment via different routes. Their behavior in the environment is determined by various physicochemical properties like vapor pressure, water solubility, and polarity. A variety of veterinary drugs used in the treatment and prevention of diseases in

Bioremediation: an effective technology 173 farming get introduced into the environment through the spraying of liquid manure onto agricultural fields. These drugs and their metabolites then contaminate soil and reach the groundwater by leaching or runoff from livestock slurries (La Farre et al., 2008). Various nonionic surfactants, like alkylphenol ethoxylates (AEOs), are also present in wastewater treatment plants. More than 90% of these AEOs are nonylphenol ethoxylates, which are extensively utilized in domestic and industrial sectors. PFCs are another group of emerging pollutants and include perfluorooctanoic acid, perfluorooctane sulfonate, and other structurally related compounds (Ju, Jin, Sasaki, & Saito, 2008; Strynar & Lindstrom, 2008). UV filters are widely employed in cosmetics, sunscreens, and other PCPs. They are found to be present in sufficient concentrations in environmental waters (Cuderman & Heath, 2007) and cause toxic effects on animals, including estrogenicity (Kunz & Fent, 2006) and endocrine and developmental toxicity (Schmitt, Oetken, Dittberner, Wagner, & Oehlmann, 2008). Benzotriazole and its derivatives have various applications, the most significant ones being their use as a corrosion inhibitor, antifoggant in photography, and UV-light stabilizer for plastics (La Farre et al., 2008). Benzotriazoles and totyltriazoles are water soluble, resist biodegradation, and undergo partial degradation in the wastewater treatment processes (Richardson & Ternes, 2005). Naphthalenic acids are known to cause toxicity to aquatic organisms including daphnia, fish, phytoplankton, and mammals; they are also potent endocrine disruptors (La Farre et al., 2008). From time immemorial, chlorination has been employed as a method to control waterborne infectious diseases. An exposure to chlorination by-products is found to be associated with incidents of cancer in human beings (Gopal, Tripathy, Bersillon, & Dubey, 2007), whereas disinfection by-products, namely iodo acids, iodine-containing trihalomethanes, bromonitromethanes, haloamides, and nnitrosodimethylamine, have potential negative impacts on human health and act as probable carcinogens (Richardson, 2005).

7.3 Types of bioremediation To date numerous types of bioremediation have been developed and employed for the treatment of emerging pollutants (Fig. 7.2). These are as follows.

7.3.1 Microbial bioremediation Microorganisms are abundant and are ubiquitous in nature (Curtis, Sloan, & Scannell, 2002) and can survive in environments of extreme temperature and pressure (Kumar, Bisht, Joshi, & Dhewa, 2011). The microorganisms possess the ability to transform complex lipophilic organics to simple water-soluble compounds (Pandey & Fulekar, 2012). They possess a highly efficient metabolic machinery which facilitates the degradation and further utilization of various toxic compounds (Paul, Pandey, Pandey, & Jain, 2005). The metabolic processes are classified as aerobic or anaerobic (Rayu, Karpouzas, & Singh, 2012). A

174 Chapter 7

Figure 7.2 Types of bioremediation employed for the treatment of the various pollutants.

complete mineralization is involved in an ideal bioremediation process yielding water and carbon dioxide as end products without the accumulation of intermediates (Sharma & Fulekar, 2009). In addition, microorganisms employ cometabolism for the degradation of some relatively resistant compounds like dibenzodioxins, carbon tetrachloride, PCBs, and trichloroethylene (Guimara˜es et al., 2010). 7.3.1.1 Bacterial bioremediation The ability of these microorganisms to utilize the pollutants as a substrate relies on their capability to generate certain enzymatic and metabolic responses in a polluted environment (Agrawal, Shrivastava, & Verma, 2019). The capability of pesticide degradation is prevalent in the members of the genera Flavobacterium, Alcaligenes, Rhodococcus, and Pseudomonas (Aislabie & Lloyd-Jones, 1995). As shown by Schrijver and Mot (1999), a significant potential for pesticide remediation is also present in Actinomycetes, especially members of the genera Rhodococcus, Arthrobacter, Nocardia, Clavibacter, Streptomyces, and Nocardioides. These microorganisms can degrade the PAHs in both aerobic and anaerobic conditions, the latter being represented by bacteria isolated from petroleumcontaminated soils, namely Brevibacillus sp. and Pseudomonas sp. (Grishchenkov et al., 2000). Also, Kafilzadeh, Sahragard, Jamali, and Tahery (2011) found ten bacterial genera, namely Bacillus, Staphylococcus, Corynebacterium, Streptococcus, Alcaligenes, Shigella, Escherichia, Acinetobacter, Enterobacter, and Klebsiella, out of 80 bacterial strains that

Bioremediation: an effective technology 175 have hydrocarbon-degrading activity and concluded that Bacillus was the most efficient genus. The major processes of hazardous waste cleanup systems also employ the degradative abilities of bacterial groups (Levinson, Stormo, Tao, & Crawford, 1994). 7.3.1.2 Mycoremediation Fungi degrade substrates to smaller molecules by the extracellular fungal enzymes. These molecules are further absorbed and metabolized in the cells (Sardrood, Goltapeh, & Varma, 2013). The exogastric nature of fungi paves the way for the degradation and utilization of complex substrates like toxic substances that are nonpolar, nonsoluble, and nondocile to the intracellular processes (Levin, Viale, & Forchiassin, 2003). There are many other advantages of mycoremediation, namely their high growth rate, ability to penetrate the pollutants, and relatively better survival under stressed conditions (low pH, nutrient limitation) (Davies & Westlake, 1979). Further, the condition of nutrient stress stimulates the fungal enzymatic machinery, thereby offering a significant benefit for bioremediation (Aust, Swaner, & Stahl, 2004; Mansur, Arias, Copa-Patino, Fla¨rdh, & Gonza´lez, 2003). In addition to that, fungi can rapidly increase their population as most have short life cycles and high sporulation rates. Many fungal genera have the ability to metabolize hydrocarbons and are thus employed in the bioremediation of oil-polluted regions. Some of these genera include Aspergillus, Saccharomyces, Candida hansenula (Bartha & Atlas, 1977), and Cladosporium (Walker, Cofone, & Cooney, 1973).

7.3.2 Phycoremediation Microalgae possess high photosynthetic abilities and are efficient fixers of carbon dioxide (Agrawal et al., 2019). These biochemical factories serve as vital members in both the aquatic and terrestrial ecosystems (Das & Chandran, 2011). Many algal species play a crucial role in bioremediation and have the ability to degrade various organic and inorganic pollutants such as PAHs. An algae, Prototheca zopfii, isolated by Walker, Colwell, Vaituzis, and Meyer (1975) showed significant degradation potential and utilized crude oil or a mixed hydrocarbon substrate. This microalga also had the ability to degrade n-alkane, isoalkanes, and other organic hydrocarbons. Also, the naphthalene degrading ability of cyanobacteria, red, brown, and green algae, and diatoms was demonstrated by (Cerniglia & Gibson, 1977). Wang & Zhao, 2007 showed that certain freshwater algae like Chlorella vulgaris, Scenedesmus quadricauda, S. capricornutum, and S. platydiscus can efficiently utilize PAHs as their substrate. Some microalgae can also increase the degradation of pyrene and fluoranthene when present along with certain bacterial groups. The mixotrophic strains of cyanobacteria and microalgae also hold a great potential for bioremediation (Chekroun, Sa´nchez, & Baghour, 2014). This mode of nutrition provides many advantages over the other microbial groups like fungi and bacteria for the degradation of organic pollutants. Such efficient mixotrophic strains can be identified and selected by various

176 Chapter 7 molecular biology methods and metabolic studies (Subashchandrabose, Ramakrishnan, Megharaj, Venkateswarlu, & Naidu, 2013). Also microalgae are photosynthetically more competent and possess high lipid production potential. These features can be further utilized for the production of alternative sources of energy like biodiesel. The biofuel production process could be achieved with or without the simultaneous treatment of wastewater and this would enhance the economic feasibility of biofuel production (Sivakumar et al., 2012).

7.3.3 Mixed cell culture system The current techniques of environmental remediation rely on consortia of bacteria, bacteria microalgae, and bacteria fungi (Kuppusamy et al., 2017). The algae bacterial consortia are more beneficial than the others for the purpose of bioremediation (Subashchandrabose, Ramakrishnan, Megharaj, Venkateswarlu, & Naidu, 2011) because many lightweight compounds and extrapolymeric substances consisting of lipids, proteins, and nucleic acids are released by the microalgae or cyanobacteria and these substances serve as substrates for the microbial growth. This further enhances the degradation of pollutants by the bacterial partners (Kirkwood, Nalewajko, & Fulthorpe, 2006). The aerobic degradation of pollutants is also increased by the oxygen supplied by microalgae (De Llasera, de Jesu´s Olmos-Espejel, Dı´az-Flores, & Montan˜o-Montiel, 2016). Thus such microalgae bacterial consortia can be further explored and utilized for the successful and efficient bioremediation of emerging environmental pollutants.

7.3.4 Phytoremediation Phytoremediation, also known as plant-assisted remediation relies on using plants with or without the aid of microbes for bioremediation (Gerhardt, Huang, Glick, & Greenberg, 2009). A low level of POP can be accumulated, immobilized, and transformed by this type of bioremediation technique (Pulford & Watson, 2003). The remediation of the polluted soil or groundwater is accomplished by this vegetation-based remediation process through a variety of mechanisms, namely phytoextraction, rhizofiltration, phytostabilization, phytodegradation, phytovolatilization, and phytostimulation (Rayu et al., 2012), as represented in Table 7.2. 7.3.4.1 Phytoextraction In the process of phytoextraction or phytoaccumulation, pollutants are removed from sediments, soils, or water into harvestable plant biomass by the activity of plants or algae. This process can occur naturally and continuously with the help of hyperaccumulators, or chelates can be added to enhance bioavailability and induce phytoaccumulation (Utmazian & Wenzel, 2006). The pollutants are initially absorbed by the plants via the root system and further they are transported upward to stems and leaves or stored in the root biomass.

Bioremediation: an effective technology 177 Table 7.2: Types of phytoremediation processes and their descriptions. Phytoremediation methods

Process description

References

Rhizofiltration

Plant is removed after the uptake of metals and their accumulation into plant tissue Metal uptake into roots of plant

Phytostabilization

Migration and mobility of contaminated soil is reduced

Kumar, Dushenkov, Motto, and Raskin (1995) Dushenkov, Kumar, Motto, and Raskin (1995) Bolan, Park, Robinson, Naidu, and Huh (2011) Newman and Reynolds (2004) Limmer and Burken (2016)

Phytoextraction

Phytodegradation

The organic contaminants undergo enzyme-catalyzed metabolism by rhizospheric microorganisms Phytovolatilization Evapotranspiration processes is employed to remove mercury, selenium, and volatile hydrocarbons Phytostimulation Polychlorinated biphenyls (PCBs) are removed by the plant soil microbe symbiotic interaction

Nwoko (2010)

The pollutants are absorbed by the plant until the harvest period following which a lesser amount of pollutant will persist in the soil. Thus to ensure a substantial cleanup several crops must be employed to repeat the growth harvest cycles. The cleaned soil can then support other kinds of vegetation (Mukherjee & Kumar, 2012). 7.3.4.2 Rhizofiltration This is employed mostly for the remediation of pollutants present in groundwater. It involves the adsorption of pollutants on the root surface or absorption by the plant roots. The plants in this technique are first adapted to the pollutant and not fixed directly in in situ conditions (Pletsch, de Araujo, & Charlwood, 1999). A dense root system is initially established by growing the plant hydroponically and later the plant is exposed to polluted water for acclimatization. After the familiarization of the plants, they are sown in the target polluted area. The pollutants along with the polluted water are then up taken by the plants until the roots reach the point of saturation. These roots are then harvested and disposed of in a systematic and safe manner (Kumar et al., 1995). 7.3.4.3 Phytostabilization This technique comprises the immobilization of toxic pollutants in the soil and involves the use of vegetation to hold the polluted soils and sediments in place. Herbaceous species, grasses, wetland species, and other fibrous root system-containing plants are employed for the purpose of phytostabilization (Salt et al., 1995). The windblown dust is also prevented by the rooted vegetation, thereby reducing the human health-related risks. The movement of leachate toward groundwater or the receiving waters is also prevented by the transpiration of large volumes of water via plants, thereby paving the way for hydraulic control. The subsurface pollutants can also be stabilized by the prevention of the

178 Chapter 7 interaction between waste and water and this is applied mostly in the case of landfill covers (Niti, Sunita, Kamlesh, & Rakesh, 2013). 7.3.4.4 Phytodegradation Phytodegradation, also known as phytotransformation, refers to the process of change in the chemical structure of a compound without its complete breakdown into molecules like carbon dioxide and water (Chaudhry, 1998). It consists of the degradation of the compounds that are complex and recalcitrant in nature into smaller and simpler forms by the action of plant enzymes (Newman & Reynolds, 2004). The released plant exudates can also contribute to remediation through the process of cometabolism in the rhizosphere. The phytotransformation of xenobiotics is known as the “Green Liver Model” as the plants show analogous behavior to the human liver. It comprises two steps, namely, Phase I and Phase II metabolism. The former involves the addition of functional groups like hydroxyl groups by the plant enzymes to the xenobiotic compounds, thereby increasing their polarity. This is like the enhancement of polarity of drugs by the human liver. The Phase II metabolism of phytotransformation comprises the addition of glucose and amino acids to the polarized xenobiotics. This further enhances the polarity of the compounds and is called conjugation. The toxicity is thus reduced, and pollutants are treated by the plants (Niti et al., 2013). 7.3.4.5 Phytovolatilization This mechanism of phytoremediation refers to the process of volatilization of the pollutants from the leaf stomata or stem of the plant. Radial diffusion can occur through the tissues of the plant stem as reported by Narayanan, Russell, Davis, and Erickson (1999) and Zhang, Davis, and Erickson (2001). An example includes the escape of methyl tert-butyl ether (MTBE) into the atmosphere via the leaves, stems, and bark (Hong et al., 2001). Plants employ radial diffusion as the prime mechanism of dissipation rather than transpiration through the leaves (Ma & Burken, 2003). 7.3.4.6 Rhizoremediation This method relies on the use of rhizospheric microorganisms. The microorganisms selected by the enrichment method can also be employed for the purpose of the bioremediation of target pollutants. The root exudate serves as the prime source of food for these microbes (Niti et al., 2013). Owing to the favorable conditions of moisture, nutrients, oxygen in soil, and the support of the microbial population by the root exudate, the rhizosphere holds a great potential for bioremediation (Davis, Erickson, Narayanan, & Zhang, 2003). The activity of soil microorganisms is stimulated by nutrients provided by root exudates. This further enhances the plant growth and decreases its metal toxicity. Many factors like soil condition, plant species, climate conditions, and microorganisms associated with rhizosphere determine the success of this technique (Prasad, Freitas, Fraenzle,

Bioremediation: an effective technology 179 Wuenschmann, & Markert, 2010). The microorganisms present in close association with roots are the plant growth promoting rhizobacteria (PGPR). These microbes also play a prime role in nutrient recycling, detoxification of harmful substances, plant pest control, and soil structure (Rajkumar, Ae, Prasad, & Freitas, 2010; Rajkumar, Vara Prasad, Freitas, & Ae, 2009). The PGPR and arbuscular mycorrhizal fungi are most widely employed for the bioremediation of the polluted soils (Ma, Prasad, Rajkumar, & Freitas, 2011).

7.3.5 Enzymatic bioremediation Various kinds of organic pollutants can be degraded by the action of microbial enzymes (Wainwright, 1999). The reactions of biodegradation are mediated by those enzymes that are responsible for activating certain energy yielding metabolic pathways, for example, this includes the species that have the potential to thrive in the absence or presence of oxygen. Such species have the capability to utilize molecular oxygen (in the case of aerobic) or nitrate (in the case of anaerobic) as the final electron acceptor in the electron transport chain. The fermentative microbes, however, do not require either of the electron acceptors and they obtain the desired energy by a fermentative process (Pandey & Fulekar, 2012). In addition, some microorganisms secrete enzymes that are functional only in the extracellular environment. A few of the genes encoding these crucial biodegradative enzymes are also present on plasmids that can be exchanged among different microbes in a population. This could lead to the genesis of new microbial strains exhibiting a wide array of degradative capabilities with applications in many fields (Scott et al., 2011; Singh, 2009).

7.3.6 Zooremediation As the word implies, zooremediation relies on the activities of animals for the removal of environmental pollutants. Various animals like fishes, arthropods, aquatic filter feeders, and earthworms can be employed for the purpose of remediation (Gupta, Yunus, & Pandey, 2003). The use of this technique is limited due to the ethical, human, and ecological safety concerns associated with the use of animals (Dada, Njoku, Osuntoki, & Akinola, 2015). However, invertebrates like clams, oysters, mussels, fishes, sponges, earthworms, and polychaetes can be used for bioremediation since the word “animal” in most jurisdictions is considered to include “all living forms that are nonhuman vertebrates,” thus allowing the utilization of various invertebrates (Gifford, Dunstan, O’Connor, Koller, & MacFarlane, 2007).

7.3.7 Vermiremediation This technique employs earthworms to degrade pollutants present in the soil environment. It relies on the ability of earthworms to survive successfully in the presence of a wide variety

180 Chapter 7 of organic and inorganic pollutants, such as PAHs, crude oil, heavy metals, and pesticides (Pattnaik & Reddy, 2011). These organisms have the potential for the remediation of target pollutants through various processes like biotransformation, biodegradation, physical actions, or bioaccumulation (Dada et al., 2015).

7.4 Emerging techniques The advancement of technologies has led to the development of techniques which are efficient for the effective removal of contaminants, for example, the use of biosurfactants, immobilization, adsorption and electrostatic binding (Fig. 7.3).

7.4.1 Application of biosurfactants Biosurfactants are employed for enhancing the bioavailability of hydrophobic pollutants (Juwarkar, Misra, & Sharma, 2014). The chemical surfactants can overcome these problems, however, certain amphiphilic biological molecules capable of partitioning at the interfaces are now widely utilized as a green and economically feasible alternative to the use of chemical surfactants (Banat, Makkar, & Cameotra, 2000). Biosurfactants are defined as microbial substances that possess high surface and emulsifying activity. Living cells synthesize surface-active substances which are ecofriendly in nature, highly selective, and exhibit diversity (Rahman, Rahman, McClean, Marchant, & Banat, 2002). Their diversity is observed due to the different microbial genes involved in their production (Bodour et al., 2004). Thus the hydrophobicity of the target pollutants is reduced and their bioavailability is enhanced using biosurfactants (Ron & Rosenberg, 2002).

Figure 7.3 Various emerging bioremediation techniques used effectively for the removal of contaminants.

Bioremediation: an effective technology 181

7.4.2 Immobilization techniques Immobilization limits the mobility of microbes or their enzymes while preserving their catalytic functions and viability (Dzionek, Wojcieszy´nska, & Guzik, 2016). Recent bioremediation techniques employ various methods of immobilization. The economic feasibility and efficiency of the bioremediation process is significantly enhanced by various immobilization techniques (Guzik, Hupert-Kocurek, Krysiak, & Wojcieszy´nska, 2014; Guzik, Hupert-Kocurek, & Wojcieszy´nska, 2014). Other advantages are higher efficiency, cost reduction, reduced risk of genetic mutation, increased survival rate of biocatalyst during storage, establishment of a stable microenvironment for cells/enzymes, and increased tolerance to higher concentration of pollutants (Bayat, Hassanshahian, & Cappello, 2015). Adsorption, entrapment, encapsulation, and covalent/electrostatic binding on a surface are the main techniques of immobilization (Kourkoutas, Bekatorou, Banat, Marchant, & Koutinas, 2004).

7.4.3 Adsorption and electrostatic binding The physical interaction of microbial cells/enzymes with the surface of water insoluble carriers is responsible for the immobilization process (Dzionek et al., 2016). This is the most widely used method in bioremediation as it is rapid, cost-effective, simple, and ecofriendly. The formation of weak bonds is responsible for adsorption on the carrier surface. Due to this, the probability of cells leaking from the carrier is very high. Hence, this method is not utilized for the immobilization of genetically modified (GM) microorganisms (Bayat et al., 2015). Electrostatic binding is similar to physical adsorption; however, there is lower probability of microorganisms leaking into the surrounding environment (Dzionek et al., 2016). The negatively charged cells/enzymes are attracted toward the hydrophilic surface of the carrier, which in turn is developed by washing the carrier surface with the buffer solution (Hudson, Magner, Cooney, & Hodnett, 2005). The immobilization mechanism is different in cases where covalent binding is involved. This is because a binding agent is required in covalent binding. Chemically activated carriers that have abundant ether, amide, or carbamate bonds are responsible for the immobilization process. Owing to the toxicity of the binding agents to the microbial cells, this method is mainly employed for enzyme immobilization (Cabana, Alexandre, Agathos, & Jones, 2009; Guzik, Hupert-Kocurek, Krysiak, et al., 2014; Guzik, Hupert-Kocurek, & Wojcieszy´nska, 2014).

7.4.4 Entrapment in porous matrix and encapsulation This is a well-known rapid, versatile, nontoxic, and inexpensive method which is widely employed in bioremediation (Wojcieszy´nska, Hupert-Kocurek, & Guzik, 2013).

182 Chapter 7 The entrapment further provides the advantage of protecting the microorganisms (Wojcieszy´nska, Hupert-Kocurek, Jankowska, & Guzik, 2012). The microbial cells can move within a carrier after the entrapment which further prevents the leakage of cells in the surrounding environment. However, the exchange of metabolites and nutrients may be limited by this process. The heterogeneous carrier entraps physiologically diverse microorganisms with those possessing high metabolic activity present near the surface and the starved cells in the interior of carrier (Bleve et al., 2011). The ratio of the pore size of the carrier to the size of microbial cell is an aspect of prime importance because if the pore size is larger than the microbial cells will leak into the environment (Kourkoutas et al., 2004). Encapsulation is similar to the entrapment technique, but here a semipermeable membrane is involved which separates the immobilized particles from the external environment. It provides the advantage that the biological material is significantly protected from adverse environmental conditions. The probability of membrane damage by growing cells and the limited permeability inhibits the use of the encapsulation technique in ex situ bioremediation processes (Klein et al., 2012).

7.4.5 Electrokinetic remediation Electrokinetic remediation involves the application of direct current between appropriately distributed electrodes in the soil (Pazos, Rosales, Alca´ntara, Go´mez, & Sanroma´n, 2010) and is used in low hydraulic permeability soils. This leads to the transportation of ionic pollutants by electromigration to the oppositely charged electrodes. The soluble pollutants are mobilized by the electroosmotic flow (Reddy, Ala, Sharma, & Kumar, 2006). The use of this technique for PAHs removal is limited by the hydrophobic nature of PAHs and their slow desorption rates. This has led to the emergence of in situ electrokinetic techniques employing surfactants, cyclodextrins, and cosolvents for PAHs removal (Kuppusamy et al., 2017).

7.4.6 Metagenomics Metagenomics allows for the increased and more efficient utilization of transgenic plants and microbes in bioremediation sectors. This technique involves the harvest of total genetic material from environmental samples, thereby negating the intermediate step of cultivation. Further, the clone libraries are generated by transferring entire genetic material into a surrogate host, generally Escherichia coli (Singh, 2010). The beneficial biotechnological products are then searched from the metagenomic library by either obtaining the sequence of the entire metagenome or functional screening of the obtained libraries. The use of this technique is limited by the logistics to generate, maintain, and screen the clone libraries (Rayu et al., 2012). Also mass production of novel degrading enzymes from bacteria that are uncultivable in nature can be achieved by metagenomics. This contributes to increase

Bioremediation: an effective technology 183 the efficiency of the enzymatic remediation process. Fan, Liu, Huang, and Liu (2012) isolated a pyrethroid hydrolyzing enzyme. This novel thermostable enzyme could be used in pyrethroid detoxification. Similarly, a novel gene having the ability to degrade a toxic metabolite, that is, 3,5,6-trichloro-2-pyridinol of the insecticide chlorpyrifos, was isolated from the cow rumen (Math et al., 2010).

7.4.7 Protein engineering Proteins can be fine-tuned to achieve the desired stereoselectivity and substrate-specificity. The kinetic properties and stability of the protein/enzymes can be improved by various engineering methods, such as random priming (Shao, Zhao, Giver, & Arnold, 1998), DNA shuffling (Kuchner & Arnold, 1997), and the staggered extension process (Zhao, Giver, Shao, Affholter, & Arnold, 1998). The technique of site-directed mutagenesis (SDM) can also be employed for rational protein design (Ju & Parales, 2006). For example, the topology and volume of the active site of cytochrome P450 was engineered by SDM. This enhanced the catalytic activity of the enzyme (Holloway, Knoke, Trevors, & Lee, 1998). Chimeric enzymes can also be produced by various engineering methods. These enzymes are functionally superior to their parent enzymes (Beil, Mason, Timmis, & Pieper, 1998). They are produced by combining the best properties of related enzymes and exchanging various sequences or subunits. The catalytic power of the enzymes can be further increased by the incorporation of multiple binding sites within a single peptide which facilitates the attachment of various cofactors and other molecules to the enzyme (Rayu et al., 2012). This strategy can be utilized for the removal of metal wastes (Pazirandeh, Wells, & Ryan, 1998). However, these modifications require great caution as many alterations can interfere with the catalytic activity and three-dimensional folding of the target protein/enzyme.

7.4.8 Bioinformatics Preinformation of various cellular processes associated with bioremediation is acquired by employing bioinformatics. Various features like protein, gene functions, metabolic and regulatory pathways, molecular mechanisms, and other cellular processes can be identified and analyzed using various tools of bioinformatics (Juwarkar et al., 2014). It allows for efficient control over the microbial cells that serve as factories. Bioinformatics is also widely employed in the field of bioremediation to determine the xenobiotic degradation pathways and structure function relationships (Fulekar & Sharma, 2008).

7.4.9 Nanotechnology “Nanobioremediation” is defined as the utilization of nanoparticles to enhance the activity of microbes for pollutant degradation. The process time as well as the associated costs is

184 Chapter 7 reduced with this nano-based technology. The biofabrication of bifunctional macromolecules, which are used as tools for the manipulation or construction of nanoobjects, is defined as “nanotechnology through biotechnology” or “bionanotechnology.” The microbial cells serve as ideal producers of nanostructures due to their small size, controlled culture conditions, wide physiological diversity, and genetic manipulability (Dixit et al., 2015). These nanostructures range from natural products like magnetosomes and polymers to engineered proteinaceous constructs like tailored metal particles or virus-like proteins (Sarikaya, Tamerler, Jen, Schulten, & Baneyx, 2003). Nanoremediation can be also be employed for PAHs removal by the development of nanooxidizers and nanofertilizers. It can also be integrated with already established PAHs removal techniques to achieve increased efficiencies and higher removal rates (Kuppusamy et al., 2017). Nanobioremediation is thus a promising technique to solve the problem of emerging pollutants such as heavy metals and other organic contaminants in the environment (Kuppusamy, Thavamani, Megharaj, & Naidu, 2015).

7.4.10 Genetic engineering Genetically engineered microorganisms have been successfully employed for the degradation of various emerging pollutants and are known to possess a high degradative capacity under defined conditions. The metabolic versatility and gene diversity of microorganisms have been explored using the method of genetic approach (Boricha & Fulekar, 2009). The chromosomal and extrachromosomal DNA of such microbes contain the blueprints of genes which are responsible for encoding various biodegradable enzymes. Such degradative genes can be detected and transformed into an appropriate host employing suitable vectors and promoters by recombinant DNA techniques, thereby allowing for the development of the xenobiotic degrading microorganisms. Various processes, such as polymerase chain reaction, SDM, particle bombardment, electroporation, and antisense RNA techniques, constitute the recombinant DNA technology which is employed for the production of pollutant-degrading microbes through genetic modification and strain improvement (Pandey & Fulekar, 2012).

7.4.11 Designer microbe and plant approach Genetically engineered microorganisms are defined as those organisms which consist of genetic material that has been modified by recombinant DNA technology techniques to attain more efficient strains for the bioremediation of emerging pollutants (Sayler & Ripp, 2000). These strains can tolerate adverse environmental conditions and thus serve as potential bioremediators. Various microbial biosensors have been developed by genetic engineering for the quick and accurate estimation of the level of contamination in polluted sites (Dixit et al., 2015). The genetic modification of endophytes and rhizospheric bacteria

Bioremediation: an effective technology 185 for plant-associated pollutant remediation in soil has emerged as a promising technique for the treatment of metal contamination sites (Divya & Kumar, 2011). The engineered strains of Moreaxella sp. and E. coli have demonstrated higher metal accumulation capacity relative to their wild-type strains (Bae, Mehra, Mulchandani, & Chen, 2001; Bae, Wu, Kostal, Mulchandani, & Chen, 2003). The use of the phytoremediation technique is limited by the accumulation of pollutants in plant tissue which further shortens their life span and releases pollutants into the atmosphere by volatilization. Various approaches, such as the manipulation of metal tolerance, accumulation, and the degradation ability of plants against various pollutants, can help overcome the aforementioned constraint of phytoremediation (Dixit et al., 2015). This approach involves the introduction of metal-degrading bacterial genes into the plant tissue, facilitating metal degradation within the plants (Van Aken, Tehrani, & Schnoor, 2011). Furthermore, certain fast-growing plants and high biomass producing plants, like Jatropa, Poplar, and Willow, can be used for phytoremediation and simulation energy production (Dixit et al., 2015).

7.4.12 Rhizosphere engineering This bioremediation strategy involves the addition of nutrients to the polluted soil or the addition of growth stimulants to the rhizosphere. The former enhances the microbial growth and bioremediation abilities of microbes or GM plants, while the latter provides electron donors/acceptors for the reduction of heavy metals (Lovley, 2003). The GM plants have demonstrated the production of specific compounds which facilitate the transformation of heavy metals in the rhizosphere. Many GM organisms with the ability to reduce heavy metals exhibit increased expression of enzymes like uranyl and chromate reductase (Dixit et al., 2015).

7.4.13 Manipulation of plant microbe symbiosis The storage and accumulation of the pollutants in plant materials is the main drawback of phytoremediation technology. The remediation process slows down and becomes inadequate when multiple pollutants are present (Ma et al., 2011). The combination of microbe plant symbiosis within the plant rhizosphere (Wu, Wood, Mulchandani, & Chen, 2006) or the introduction of microbes as endophytes to facilitate pollutant degradation within the plant tissues serve as appropriate solutions to this problem. The major strategies adopted for implementing bioremediation processes include bioaugmentation and biostimulation approaches, which are guided by specific microbes in combination with plants.

186 Chapter 7

7.4.14 Cometabolic bioremediation The indigenous microbial populations with the capability for simultaneous degradation of the target pollutant and cosubstrate (like methane, toluene, and propane) are stimulated by cometabolism strategies (Hazen, 2018). The amendment costs are reduced by ensuring that the only microbes targeted are those that have the potential to degrade pollutants. This technique is employed for recalcitrant pollutants, such as dioxane (Mahendra, Petzold, Baidoo, Keasling, & Alvarez-Cohen, 2007), perchloroethylene (Enzien, Picardal, Hazen, Arnold, & Fliermans, 1994), MTBE (Chen, Kao, Chen, Weng, & Tsai, 2006), trichloroethylene (Ensley, 1991), trinitrotoluene (Yasin, Shah, Hameed, Ahmed, & Hasan, 2008), and atrazine (Ghosh and Philip, 2004). For example, methanotrophs can produce methane monooxygenase which can degrade over one thousand different types of compounds. Recently, cometabolic bioremediation has achieved promising results for the degradation of various emerging trace organic pollutants (Liu, Binning, & Smets, 2015), such as pharmaceuticals (Gauthier, Yargeau, & Cooper, 2010), lincomycin (Li, Zhou, Gong, Wang, & He, 2016), textile dyes decolorization (Karim, Dhar, & Hossain, 2017), 1,1,2,2-tetrachloroehane (Cappelletti, Pinelli, Fedi, Zannoni, & Frascari, 2018), carbazole (Shi, Qu, Zhou, Ma, & Ma, 2015), tetrabromobisphenol (Gu et al., 2016), and dibenzofuran (Shi et al., 2013).

7.5 Conclusion The emerging pollutants are pollutants which cannot be placed within any other particular group as they can only be categorized after their detection. As the industrialization increases the detection of emerging pollutants is also increasing, leading to adverse effects on the environment and life forms. On the other hand, this problem can be addressed by the various treatment methodologies which can help in the effective removal of the pollutants. The existing bioremediation methods and the new technical advancements like metagenomics, genetic and protein engineering, nanotechnology, and bioinformatics, can further enhance the application of these technologies at a larger scale with minimal side effects on the environment, thus enabling a slow and gradual approach toward a “cleaner and healthier surrounding” for all life forms.

Acknowledgment PV is thankful to DBT (Grant No. BT/304/NE/TBP/2012; Grant No. BT/PR7333/PBD/26/373/2012), KA is thankful to Central University of Rajasthan, Ajmer, India for providing the financial support.

Competing interests All the authors declare that they have no competing interests.

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CHAPTER 8

Application of metagenomics in remediation of contaminated sites and environmental restoration Vineet Kumar1,2, Indu Shekhar Thakur1, Ajay Kumar Singh2 and Maulin P. Shah3 1

Environmental Microbiology and Biotechnology Laboratory, School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India, 2Department of Environmental Microbiology, School of Environmental Sciences, Babasaheb Bhimrao Ambedkar (A Central) University, Lucknow, India, 3 Industrial Waste Water Research Laboratory, Division of Applied & Environmental Microbiology, Enviro Technology Limited, Ankleshwar, India

8.1 Introduction Rapid industrialization and anthropogenic activities are associated with a continuous buildup of environmental contaminants and pollution, posing considerable risk to human and environmental health (Chandra, Dubey, & Kumar, 2018; Chandra & Kumar, 2017a, 2017b). A huge amount of toxic waste containing a mixture of numerous organic and inorganic pollutants, such as petroleum hydrocarbons (i.e., benzene, pyrene, toluene, and xylene); polycyclic aromatic hydrocarbons (PAHs); polychlorinated biphenyls (PCBs); phenols; pesticides; herbicides; insecticides; chlorinated solvents [i.e., trichloroethylene (TCE)]; explosives, such as 2,4,6-trinitrotoluene, hexahydro-1,3,5-trinitro-1,3,5-triazene, and octahydro-1,3,5,7-tetranitro-1,3,5-tetrazocine; and heavy metals, that is, iron (Fe), lead (Pb), zinc (Zn), cadmium (Cd), chromium (Cr), cobalt (Co), copper (Cu), nickel (Ni), and mercury (Hg), has been dispersed in the environment; this threat is likely to worsen unless action is taken to remediate the several million contaminated sites occurring globally (Chandra & Kumar, 2015; Chandra, Kumar, & Yadav, 2015). Hence the challenge is to develop innovative cost-effective sustainable solutions to decontaminate polluted environments, to make them safe for human habitation and consumption, and to protect the functions of the life-supporting ecosystems (Fig. 8.1). To safeguard both humans and the environment from the adverse consequences of toxic and hazardous pollutants novel approaches must be designed, and bioremediation is one such approach (Kumar, AlMomin, et al., 2018; Kumar, Shahi, & Singh, 2018). Bioremediation is a commonly used method to

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00008-0 © 2020 Elsevier Inc. All rights reserved.

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198 Chapter 8

Figure 8.1 The various sites showing the soil and water contamination due to disposal of partially treated or untreated effluent from industries. (A) Tannery waste-contaminated site; (B) pulp and paper mill waste-contaminated site; and (C) distillery waste-contaminated site.

restore the natural and useful values of contaminated sites by microorganisms able to degrade, oxidize/reduce, bind, immobilize, volatile, chelate, or transform various toxic contaminants (Chakraborty, Wu, & Hazen, 2012; Chandra & Kumar, 2015). In the current scenario, microbial-mediated bioremediation is of great significance because it promises an inexpensive, ecofriendly, and simpler method when compared to the more commonly employed chemical-based nonbiological remedies (Das & Dash, 2014; Kumar & Chandra, 2018a, 2018b, 2020a, 2020b). Though the word “bioremediation” was first used in 1987 in peer-reviewed scientific literature (Hazen, 1997), the concept had a link to the early days of human civilization (about 6000 BCE) and the relatively “modern” use of bioremediation was apparent in 1891, with the opening of the first biological sewage treatment plant in Sussex, United Kingdom (Tabak, Lens, Hullebusch, & Dejonghe, 2005). If the bioremediation process occurs at the site affected by pollutants, wherein there is no possibility to transfer polluted soil and/or water, then it is called in situ bioremediation. In contrast, deliberate relocation of the contaminated material (soil and water) to a different place to accelerate biocatalysis is referred to as ex situ bioremediation (USEPA, 2012).

Metagenomics in remediation of contaminated sites 199 As microbes are the primary pollutant degraders in contaminated habitats, understanding microbial processes in individual sites is essential for predicting the best strategy for bioremediation. There are three basic methods of in situ bioremediation: (1) natural attenuation; (2) biostimulation; and (3) bioaugmentation (Tyagi, da Fonseca, & de Carvalho, 2011; USEPA, 1999, 2012). Natural attenuation, also known as passive bioremediation, is an appealing, invasive approach, whereby native (indigenous) microorganisms are used to detoxify contaminants using natural processes without utilizing any artificial steps (Margesin & Schinner, 2001). This method avoids damaging the habitat, allows the ecosystem to revert to its original condition and enables the detoxification of toxic pollutants (NRC, 2000). Intrinsic bioremediation depends on the underlying metabolic activities of indigenous microorganisms to degrade and/or transform toxic pollutants and accelerate the degradation process (Kumar, AlMomin, et al., 2018; Kumar, Shahi, et al., 2018; USEPA, 2012). However, in many systems, the rates of natural attenuation may be prohibitively slow and not responsive to environmental and health risks. On the other hand, a bioaugmentation method is applied when the indigenous microorganisms are unable to break down pollutants, or when the population of microorganisms capable of degrading contaminants is not sufficiently large (Gough & Nielsen, 2016; Nzila, Razzak, & Zhu, 2016). To make the bioaugmentation process successful, microorganisms introduced into a polluted environment as a free or immobilized inoculum should be able to degrade specific contamination and survive in a foreign and unfriendly habitat, be viable and genetically stable, and move through the contaminated media. However, the result of bioaugmentation depends on the interaction between exogenous and indigenous populations of microorganisms because of the competition, mainly for nutrients (Major et al., 2002; Mrozik & Piotrowska-Seget, 2010). Biostimulation consists of boosting the degrading indigenous microflora by the addition of nutrients or chemicals in a contaminated environment (Roy et al., 2018). Thus the remediation of contaminated sites using the metabolic activities of microbes and plants is a permanent solution that can end with the degradation or transformation of environmental contaminants into harmless or less toxic forms (Simpanen et al., 2016). However, bioremediation technology has been hampered by the lack of a holistic systemwide understanding of the complex interactions between degrading organisms and genes, the wider metabolic network of the microbial community, and the environmental variability in each specific habitat (Dzionek, Wojcieszy´nska, & Guzik, 2016). Thus there is an urgent need to explore the relationship between microbial community composition and pollutants degradation potential and to elucidate which chemical variables and microbial diversity metrics can predict chemical degradation. The use of molecular biology and metagenomics has also greatly expanded our understanding of the biological systems present in the contaminated environments and in many cases has greatly enhanced our understanding of the microbial world (Handelsman, 2004; Hugenholtz & Tyson, 2008). Metagenomics is a

200 Chapter 8 strategic approach of analyzing complex microbial communities growing in the contaminated environment using DNA directly extracted from environmental samples (Hugenholtz & Tyson, 2008; Kachienga, Jitendra, & Momba, 2018). Metagenomics studies reveal the extent of microbial diversity in natural habitats, including uncultured microbes, and help to improve our understanding of the biological functions within microbial communities. This provides a glimpse of the microbial community view of “Uncultured Microbiota.” Recent studies suggest that microbial communities are the potential alternatives to eliminate toxic contaminants from our environment (Huaidong, Waichin, Riqing, & Zhihong, 2017; Jung, Philippot, & Par, 2016; Kumar, AlMomin, et al., 2018; Kumar, Shahi, et al., 2018; Maphosa et al., 2012; Sahu, Deori, & Ghosh, 2018). In this chapter we seek to provide a key background on metagenomics-based approaches and summarize how these tools have been employed to understand contaminated environments in an effort to inform the best practices for cleanup of contaminated environment. We also discussed recent approaches of metagenomics employed in the remediation of water, wastewater, and soil contaminants with the help of multiple case studies.

8.2 Mechanism of bioremediation Microbial bioremediation is increasingly considered a reliable, cost-effective, and sustainable waste management technique for remediation of environmental contaminants and eco-restoration of polluted sites (Kumar, AlMomin, et al., 2018; Kumar & Chandra, 2020a, 2020b; Kumar, Shahi, et al., 2018). Microorganisms have the ability to interact, both physically and chemically, with substances, leading to structural changes or complete mineralization of the target pollutants. A huge number of bacteria, fungi, and actinomycetes genera possess the capability to degrade and/or detoxify organic and inorganic pollutants in the environment. The biodegradation of organic pollutants through microorganism is based on two processes (1) primary metabolism and (2) cometabolism (Luo et al., 2014). Primary metabolism of an organic compound has been defined as the use of the substrate as a source of carbon and energy. This substrate serves as an electron donor resulting in microbial growth. This process results in complete degradation of organic pollutants. Cometabolism, also called gratuitous metabolism, cooxidation, accidental or free metabolism, often happens in the natural environment where those degradation processes are accompanied by the transformations of other compounds, other xenobiotics (Nzila, 2013; Smith, O’Reilly, & Hyman, 2003). The term cometabolism has been defined as the metabolism of an organic compound that does not serve as a source of carbon and energy or as an essential nutrient which can be achieved only in the presence of a primary substrate (Dalton & Stirling, 1982). The application of cometabolism to remediation of a xenobiotics-contaminated site is required when the compound cannot serve as a source of carbon and energy by nature of the molecular structure, which does not induce the required catabolic enzymes. It is a common phenomenon where microbial activity occurs at the contaminated site.

Metagenomics in remediation of contaminated sites 201 The prerequisites of metabolic transformation are the enzymes of the growing cells and the synthesis of cofactors necessary for the enzymatic reaction, for example, hydrogen donors (nicotinamide adenine dinucleotide, NADH) for oxygenase (Nzila, 2013). Cometabolism of pollutants was mainly reported for PAHs with more than five aromatic rings, chlorinated biphenyls, chlorinated monoaromatics, and chlorinated aliphatics. Depending on the contaminant of concern and the media, a technology may exploit aerobic or anaerobic metabolism of heterotrophic microorganisms. Aerobic metabolism, also called aerobic respiration, involves oxygen as a reactant to oxidize part of the carbon in the contaminant to carbon dioxide (CO2), with the rest of the carbon used to produce new cell mass. Aerobic metabolism is more commonly exploited and can be effective for hydrocarbons and other organic compounds. In some cases the contaminants are aerobically degraded to CO2, water, and microbial biomass, but in other cases the microbes do not completely degrade contaminants (Chandra & Kumar, 2015). Dioxygenases and monooxygenases are two of the primary enzymes employed by aerobic organisms during the transformation and mineralization of xenobiotics (Camacho-Pe´rez, Rı´os-Leal, Rinderknecht-Seijas, & PoggiVaraldo, 2012). On the other hand, anaerobic metabolism involves microbial reactions occurring in the absence of oxygen and encompasses many processes including fermentation, methanogenesis, reductive dechlorination, sulfate-reducing activities, and denitrification (Zhang & Bennett, 2005). In anaerobic metabolism, nitrate, sulfate, CO2, oxidized metals, or organic compounds may replace oxygen as the electron acceptor (Chandra & Kumar, 2015; Heider, Spormann, Beller, & Widdel, 1998; Zhang & Bennett, 2005). Anaerobic metabolism has been used as the most common method in bioremediation. However, anaerobic metabolism has some drawbacks. For instance, Trichloroethylene (TCE) degradation from microbial reductive dechlorination produces dichloroethylene (DCE) (cis-DCE and trans-DCE), vinyl chloride (VC), and ethylene. Particularly, DCE and VC are more toxic to the environment than TCE itself and can also accumulate in the environment (Griffin, Tiedje, & Lo¨ffler, 2004; Zhang & Bennett, 2005). The indigenous microbial communities grown in polluted environments have a high tolerance toward toxic pollutants and utilize these pollutants as a carbon or nitrogen source and/or energy source and degrade them into simpler intermediates up to complete mineralization (Feng et al., 2018; Guo et al., 2019; Siciliano, Germida, Banks, & Greer, 2003) However, contaminated environments usually have a mixture of several organic and inorganic compounds. Therefore complete remediation of such pollutants requires synergistic interactions between different degrading bacteria, bacteria fungi, bacteria plants, fungi plants, or bacteria fungi plants (Gremion, Chatzinotas, Kaufmann, Sigler, & Harms, 2004; Hou et al., 2017; Kumar & Chandra, 2018a, 2018b, 2020a, 2020b; Palmroth et al., 2007). Readily biodegradable contaminants may remain undegraded or biodegrade very slowly if their concentrations in the matrices are too low. The uptake of organic pollutants by plants and microorganisms depends on a number of physicochemical characteristics, such as octanol water partition coefficient (log Kow), acidity constant

202 Chapter 8 (pKa), aqueous solubility (Sw), octanol solubility, and the concentration of the pollutant. Octanol water partition coefficient is the determining factor for the entry and translocation of organic pollutants inside the cells of plants and microorganisms. Some of the microorganisms are unable to mineralize these pollutants because they may not be recognized as a substrate by the existing enzymes used for degradation. Sometimes, they might be chemically and biologically very stable, containing substitution groups like amino, carbamyl, halogens, methoxy, nitro, and sulfonate. Moreover, in some cases, the compounds might be insoluble in water and could remain adsorbed to the external matrices of soil. Furthermore, the large molecular size of persistent organic pollutants (POPs) and absence of permease in the microbial cell might reduce their transport into microbial cells. Therefore there may be numerous reasons for unproductive biodegradation in contaminated sites. A key concept in evaluating all bioremediation technologies is microbial bioavailability. Bioavailability is defined as the fractions of a pollutant in soil that can be taken up or altered by living organisms. Simply stated, if the contaminant of concern is so tightly bound up in the solid matrix that microorganisms cannot access it, then it cannot be remediated. There are two important factors that determine the amount of bioavailable pollutant: mass transfer and intrinsic activity of the cell. Bioavailability differs between species and organisms and the in situ microbial degradation of organic pollutant is a function of the bioavailability of contaminant and catabolic activity of microbes. Removal rates and extent vary based on the contaminant of concern and site-specific characteristics. Removal rates also are affected by variables such as contaminant distribution and concentration; cocontaminant concentrations; indigenous microbial populations and reaction kinetics; and parameters such as pH, moisture content, nutrient supply, and temperature. Many of these factors are a function of the site and the indigenous microbial community and thus are difficult to manipulate.

8.3 Approaches used to study microbial communities involved in in situ and ex situ bioremediation Microorganisms have been integral to the history and function of life on Earth. They are ubiquitous in the environment and play a vital role in ecological functions, particularly those associated with nutrients cycling processes and the maintenance of ecosystem health in soil. The total number of prokaryotic cells on Earth has been estimated at 4 6 3 1030 (Whitman, Coleman, & Wiebe, 1998), thought to comprise between 106 and 108 separate geno species (distinct taxonomic groups based on gene sequence analysis) (Amann, Ludwig, & Schleifer, 1995). This diversity presents an enormous (and largely untapped) genetic and biological pool that can be exploited for the recovery of novel genes, entire metabolic pathways and their products (Cowan, 2000). Methods for investigating microbiota from a given ecosystem can be either culture-dependent or culture-independent.

Metagenomics in remediation of contaminated sites 203

8.3.1 Culture-based techniques The diversity of soil microorganisms that are related to bioremediation processes has been exploited for many years based on the cultivation and isolation of microbial species (Stefani et al., 2015). Generally, the majority of bioremediation studies rely on the “treatability study,” in which samples from contaminated sites are typically incubated under laboratory conditions, and the rates at which the contaminants are immobilized or degraded are recorded (Ellis, Morgan, Weightman, & Fry, 2003). These studies provide an estimate regarding the potential metabolic activities of the microbial consortia but give little insight into the microbes responsible for bioremediation (Al-Awadhi et al., 2013). Standard culturebased techniques to characterize microbial ecology involve the isolation and characterization of microorganisms using synthetic or complex growth media. Traditional cultivation on solid media or via enrichment is thought to favor fast growers at the expense of slow growers and also to mask rare species. Even though culture-based techniques are extremely useful for understanding the physiological potential of isolated organisms, they do not necessarily provide comprehensive information on the composition of microbial communities. However, only .1% of the total number of prokaryotic species grown in given environmental samples are readily cultivable under standard laboratory conditions by standard culturing techniques (Rastogi & Sani, 2011). Due to this documented disparity between cultivatable and in situ diversity, it is often difficult to assess the significance of cultured members in resident microbial communities. The uncultured bacteria could be in an uncultivated or viable but nonculturable (VBNC) state, which indicated that a large number of bacteria were still unknown in varieties of ecological environments (Ramamurthy, Ghosh, Pazhani, & Shinoda, 2014; Su et al., 2018). These VBNC cells fail to grow under many laboratory conditions but are alive and capable of growth when subject to a subset of conditions such as nutrient-rich medium or filtrate from rowing cells (Ramamurthy et al., 2014). These heavily stressed microorganisms show only very weak metabolic activity, often at the very limits of detection, and they lose the ability to form colonies on nonselective plating media or to grow in nonselective broth media. Consequently, culturing fails to reproduce the ecological niches and symbiotic relationships encountered in complex natural environments that are required to support the full spectrum of microbial diversity. In addition, when complex microbial communities are under investigation, enumerating bacteria by traditional microbial culturing techniques may produce erroneous results.

8.3.2 Culture-independent techniques In the past, it was difficult to study microbes in their own environments; microbiologists studied individual species one by one in the laboratory (Jany & Barbier, 2008; Stefani et al., 2015). It now appears that many microbes function in nature as multicellular, often

204 Chapter 8

Cell culturing

Aquatic

Cellular metabolites

Extraction of DNA/mRNA/rRNA

Genomics

Metabolomics and metafluxomics 1H-NMR/ HPLC/GC-MS/LC-MS

Microbial community analysis

Wastewater

Extraction of proteome from microbial communities

Metatranscriptomics cDNA synthesis/microarray/ hybridization/pyrosequencing

METAGENOMICS

Microarray

Genotyping ARDRA/T-RFLP/RAPD/ TGGE/ARISA/DGGE/AFLP

RT-PCR

In situ surveillance

Metagenomics libraries

16 D rRNA gene

Screening functional/sequence based

Bioinformatics Analysis of genome using various bioinformatics and computational tools

In situ technique

Metaproteomics 2D/MALDI-TOF-MS/LC-ESI-MS

Sampling

Contaminated environment Terrestrial

Molecular biosensors

CARD and MAR-FISH

DNA and RNA

Pyrosequencing GS-FLX/SOLiD/SOLEXA

Chain termination sequencing

Next generation sequencing

Figure 8.2 Molecular approaches used for the monitoring of microbial communities during the bioremediation of environmental pollutants.

multispecies, entities, sometimes even physically connected (as in biofilms) and often metabolically connected. The advent of culture-independent molecular methods has enabled the use of natural microbial communities without the need for culturing microorganisms and has introduced new insights into the microbial ecology of different ecosystems (Fig. 8.2; Su, Lei, Duan, Zhang, & Yang, 2012). The primary advantage of cultureindependent methods is their ability to characterize a large proportion of the microbial diversity that can be difficult to observe with standard culture-based studies. These methods use a DNA template as a pool deriving both from live, VBNC, and dead bacterial cells. Further, the pool of polymerase chain reaction (PCR) amplicon can be either cloned and sequenced or subjected to an increasing variety of genetic profiling methods (Nocker, Sossa, & Camper, 2007). Compared with the previous years, there has been a gradual increase in interest in culture-independent (metagenomics) bioremediation studies.

Metagenomics in remediation of contaminated sites 205 8.3.2.1 Polymerase chain reaction based molecular techniques PCR is a powerful genetic technique used in molecular biology to amplify a single copy or a few copies of a segment of DNA by several orders of magnitude and generate thousands to millions of copies of a DNA sequence of interest. PCR-based molecular methods require the extraction and purification of nucleic acids (DNA and/or RNA), followed by the amplification of a target gene or genes via PCR and post-PCR analysis. DNA-based analysis provides information on the total microbial community of the samples, while RNA-based analysis represents only the active part (Agrawal, Agrawal, & Shrivastava, 2015; Gonzalez, Portillo, Belda-Ferre, & Mira, 2012). DNA extraction from an environmental sample is followed by PCR amplification of “marker genes” to obtain taxonomic information of growing the microbial community. The most commonly used marker gene in microbiological research is the smallsubunit ribosomal RNA (SSU rRNA) gene (16S rRNA for prokaryotes and 18S rRNA for eukaryotes) recovered from environmental samples from a variety of habitats (Olsen, Lane, Giovannoni, Pace, & Stahl, 1986). The rRNA gene has different regions, some are highly conserved across all phylogenetic domains (i.e., eubacteria, eukarya, and archaea), other regions are variable between related species (Woese, 1987) and this variability allows for the inference of phylogenetic information from microorganisms inhabiting the different environments. The 16S rRNA gene sequence contains both (nine) highly conserved regions for primer design and nine hypervariable regions (V1 V9) to identify phylogenetic characteristics of microorganisms and was first used in 1985 for phylogenetic analysis. However, it is limited by sequencing technology, and the 16S rRNA gene sequences used in most studies are partial sequences, while 18S rDNA and internal transcribed spacer (ITS) regions are increasingly used to study fungal communities. Target DNA (16S, 18S, or ITS) is amplified using universal or specific primers and the resulting products are separated in different ways. The diversity of the amplified sequences is then resolved by differential electrophoretic migration on agarose or polyacrylamide gels, based on their size (terminal restriction fragment length polymorphism (T-RFLP), ribosomal intergenic spacer analysis) or sequence [denaturing gradient gel electrophoresis (DGGE), temperature gradient gel electrophoresis (TGGE)]. These so-called genetic fingerprinting techniques provide band profiles that are representative of the genetic structure of the community as a whole or of a section of it, as defined by selected primers. These methods are valuable tools for characterizing complex microbial communities and detecting shifts following environmental perturbations and are less time-consuming and labor-intensive than strategies such as small-subunit rRNA gene clone library construction. Individual bands can also be excised from a fingerprint profile, cloned and sequenced, or challenged with a range of probes providing precise information on the phylogenetic groups constituting a community. 8.3.2.1.1 Denaturing gradient gel electrophoresis/temperature gradient gel electrophoresis

DGGE/TGGE is a genetic fingerprinting technique that allows for monitoring changes in microbial communities and it can be used, similarly to other fingerprinting techniques, as

206 Chapter 8 a semiquantitative method to estimate species abundance and richness (Muyzer, 1999). In DGGE or TGGE techniques, the PCR amplicon is obtained from 16S rRNA gene amplification of an environmental sample and electrophoresed on a polyacrylamide gel containing a linear gradient of DNA denaturant (a mixture of urea and formamide) (Muyzer, de Waal, & Uitterlinden, 1993) or using a TGGE (Muyzer, 1999). The final result is a gel with a pattern of bands which is a visual profile of the most abundant species in the studied microbial community. In addition, the specific bands on the gel can be excised and sequenced for subsequent taxonomic identification. The primary advantage of this technique is that it generates a profile of the entire diversity of the microbial community by separating a mixed population of 16S rRNA gene products. However, these techniques have several limitations, biases, and drawbacks. For example, the dominant populations are better revealed, and the bands obtained from multiple numbers of species may be obscured behind a single band, thereby leading to an underestimation of the microbial diversity (Gafan & Spratt, 2005; Green, Leigh, & Neufeld, 2010; Satokari, Vaughan, Akkermans, Saarela, & De Vos, 2001). Recently, DGGE-based analysis of 16S rRNA sequences has been used to investigate and profile complex microbial diversity and to deduce the phylogenetic affiliation among these microbial communities. A compositional shift of bacterial groups in an artificial BTEX (benzene, toluene, ethylbenzene, and xylenes)-contaminated Daejeon forest soil was examined by the 16S rDNA PCR-DGGE method (Ji, Kim, Yoon, & Lee, 2007). Phylogenetic analysis of 16S rDNAs in the dominant DGGE bands showed that the number of Actinobacteria and Bacillus populations increased. To confirm these observations, the authors performed PCR to amplify the 23S rDNA and 16S rDNA against the sample metagenome using Actinobacteria-targeting and Bacilli-specific primer sets, respectively. The results confirmed that a bacterial community containing Actinobacteria and Bacillus was affected by BTEX. The impacts of heavy metals and phytoextraction practices on a soil microbial community were studied during 12-month microcosm experiment using a hyperaccumulating plant Thlaspi caerulescens grown in an artificially contaminated soil (Gremion et al., 2004). The 16S rRNA genes of the bacteria and the β-proteobacteria and the amoA gene (encoding the a-subunit of ammonia monooxygenase) were PCR-amplified and analyzed by DGGE. Principal component analysis (PCA) of the DGGE data revealed that (1) the heavy metals had the most drastic effects on the bacterial groups targeted, (2) the plant induced changes which could be observed in the amoA and in the bacteria 16S rRNA gene patterns, and (3) the changes observed during 12 months in the DGGEpatterns of the planted contaminated soil did not indicate recovery of the initial bacterial community present in the noncontaminated soil. 8.3.2.1.2 Single-strand conformation polymorphism

Single-strand conformation polymorphism (SSCP) analysis is a post-PCR technique that can be used to screen for point mutations and small deletions and insertions that are not limited

Metagenomics in remediation of contaminated sites 207 to a single hotspot but are randomly distributed throughout the exons (Gasser et al., 2006). PCR-coupled SSCP analysis is a powerful approach for investigating a range of pathogens and has proven to be useful for the identification of species and strains in the absence of distinguishing morphological characters (Dong & Zhu, 2005). SSCP has been used to measure the succession of bacterial communities (Peters, Koschinsky, Schwieger, & Tebbe, 2000), rhizosphere microbial communities (Schmalenberger, Schwieger, & Tebbe, 2001; Schwieger & Tebbe, 1998), bacterial population changes in an anaerobic bioreactor (Zumstein, Moletta, & Godon, 2000), and arbuscular mycorrhizal fungi species in roots (Kjoller & Rosendah, 2000; Simon, Levesque, & Lalonde, 1993). 8.3.2.1.3 Restriction fragment length polymorphism/amplified ribosomal DNA restriction analysis

Restriction fragment length polymorphism (RFLP), also called amplified ribosomal DNA restriction analysis (ARDRA), has proved to be a suitable and rapid method for taxonomic studies of microbial communities. The principle of RFLP markers is that any genomic DNA can be differentiated according to the presence or absence of restriction enzyme sites. Restriction enzymes recognize and cut at the particular site on DNA. Due to the accumulation of single-nucleotide mutations in genomic DNA, the restriction enzyme sites on DNA change, resulting in a difference in restriction patterns of two closely related genomes. This technique was originally developed by Botstein, White, Skolnick, and Davis (1980), using restriction endonucleases, proteins that cut the DNA molecule (obtained from PCR amplification of the 16S rDNA fragment) at short, specific sequences, called restriction sites, resulting in different fragments of variable lengths. After a segment of DNA has been cut into pieces with restriction enzymes, researchers can examine the fragments using a laboratory method called gel electrophoresis, which separates DNA fragments according to their size. After separation by agarose gel electrophoresis, digested fragments are transferred to nitrocellulose or nylon filters through southern blotting, followed by hybridization with radioactively labeled DNA probes and visualization using photographic film (Varshney, Mohapatra, & Sharma, 2004). The advantage of RFLP is that fluctuations in the microbial community can be detected by studying the banding pattern. Although it gives no information about types of microorganisms in the sample, it allows us to assess genetic changes, quickly compare different communities, or study the influence of environmental conditions or chemicals on biocenosis richness. Chandra and Kumar (2017b) used the RFLP approach to explore the microbial communities grown in the organic acid and EDCs-rich environment of distillery spent wash. Bacterial community analysis by RFLP revealed that Bacillus and Stenotrophomonas were the dominant autochthonous bacterial communities, belonging to the phylum Firmicutes and γ-Proteobacteria, respectively. The presence of Bacillus and Stenotrophomonas species in highly acidic environments indicated its broad range adaptation. These findings indicated that these autochthonous bacterial communities were pioneer taxa for in situ remediations of this

208 Chapter 8 hazardous waste during ecological succession. Chandra and Kumar (2017a) used the RFLP method to assess the microbial community composition and function during in situ bioremediation of distillery sludge. They demonstrated that Bacillus sp. and Enterococcus were found to be the dominantly growing autochthonous bacterial communities during in situ bioremediation of distillery sludge due to the availability of diverse habitats and metabolization capabilities. 8.3.2.1.4 Terminal restriction fragment length polymorphisms

T-RFLP has multiple advantages and has achieved a rapid gain in popularity: data are quantitative and comparable between laboratories; the final electrophoresis step can be performed with automated sequencing equipment at core sequencing facilities (Tiquia, 2010). In the technique reported here, the initial steps of DNA isolation, PCR amplification, and restriction were similar to those used for RFLP/ARDRA. However, one of the primers used is labeled with a fluorescent dye so that when the preparation is analyzed with an automated DNA sequencer, the sizes of only the terminal restriction fragments could be determined and the amount could be quantified (Schu¨tte et al., 2008). Due to the T-RFLP found in 16S rDNA, the method has been applied with DNA from complex microbial communities and provides a sensitive and rapid means to assess community diversity and obtain a distinctive fingerprint of a microbial community (Ding, Palmer, & Melcher, 2013; Hodgetts, Ball, Boonham, Mumford, & Dickinson, 2007; Marsh, Saxman, Cole, & Tiedje, 2000; Wu et al., 2015). In addition, T-RFLP has been used for the analysis of functional genes such as those encoding for nitrogen fixation and methane oxidation. The distribution of bacterial communities’ T-RFLP fingerprint patterns were evaluated at three proximal hydrocarbon-contaminated sites located within the harbor of Messina (Denaro et al., 2005). Authors demonstrated that TRFL is a suitable technique for monitoring polluted marine environments, typically characterized by low diversity and high relative abundances of a few dominant groups. A systematic evaluation of the value and potential of T-RFLP analysis for the study of microbial community structure and dynamics from PCB-polluted soil has been undertaken by Osborn, Moore, and Timmis (2000). An initial investigation to assess the variability both within and between different polyacrylamide gel electrophoresis runs showed that almost identical community profiles were consistently produced from the same sample. Similarly, very little variability was observed as a result of variation between replicate restriction digestions, PCR amplifications, or between replicate DNA isolations. The greatest variation between profiles generated from the same DNA sample was produced using different Taq DNA polymerases, while lower levels of variability were found between PCR products that had been produced using different annealing temperatures. Incomplete digestion by the restriction enzyme may, as a result of the generation of partially digested fragments, lead to an overestimation of the overall diversity within a community.

Metagenomics in remediation of contaminated sites 209 8.3.2.1.5 Automated ribosomal intergenic spacer analysis

Automated ribosomal intergenic spacer analysis (ARISA) is a culture-independent, PCR-based technique suitable for analyzing structures of microbial communities grown in the environment. This technique was first developed by Fisher and Triplett in 1999 and it is based on the length polymorphism of intergenic spacer sequences between the small (16S) and the large (23S) subunit of rRNA genes amplified with universal eubacterial primers directly on the DNA extracted from the microbial community grown in a contaminated environment. The sequences most commonly used as markers of communities are the genes of the ribosomal operon. The PCR products are analyzed by an automated capillary electrophoresis system that produces an electropherogram, the peaks of which correspond to discrete DNA fragments detected by a laser-based fluorescence detection system. The sensitivity of ARISA is very high (detection level, single-nucleotide difference), and reproducibility is guaranteed by instrumental automatism and commonly used in studying microbial community structure. Wood, Zhang, Mathews, Tang, and Franks (2016) investigated the impact of Cd, and the presence of a Cd-accumulating plant, Carpobrotus rossii (Haw.) Schwantes, on the structure of soil bacterial and fungal communities using ARISA and quantitative polymerase chain reaction (qPCR). Whilst Cd had no detectable influence upon fungal communities, bacterial communities underwent significant structural changes with no reduction in 16S rRNA copy number. The presence of C. rossii influenced the structure of all communities and increased ITS copy number. Suites of operational taxonomic units (OTUs) changed in abundance in response to either Cd or C. rossii, however, they found little evidence to suggest that the two selective pressures were acting synergistically. The Cd-induced turnover in bacterial OTUs suggests that Cd alters competition dynamics within the community. 8.3.2.1.6 Random amplified polymorphic DNA

Random amplified polymorphic DNA (RAPD), also known as arbitrarily primed polymerase chain reaction, is widely adapted for rapid detection of genomic polymorphism. The technique is based on the amplification of the whole-community DNA with either a single or multiple short oligonucleotide primers (8 12 nucleotides in length and designed to a number of random target sites) of an arbitrary or random sequence usually in the size range of 8 15 nucleotides in length (Franklin, Taylor, & Mills, 1999). The number and location of these random primer sites vary for different strains of a bacterial species. Lowstringency conditions allow the primer to hybridize to both strands of template DNA where it is matched or partially matched, resulting in strain-specific heterogeneous DNA products. Complex patterns of PCR products are generated as these random sequence primers anneal to various regions in an organism’s genome. What results from the separation of the amplification products by agarose gel electrophoresis is therefore a semiunique profile of bands garnered from an RAPD reaction, and the band patterns can be compared to

210 Chapter 8 determine percent similarity. The advantage of this approach is that it requires very little sample material and obtains results rapidly. However, this method is less susceptible to base changes in the target DNA. RAPD suffers from poor reproducibility between laboratories largely because of the requirement for consistent PCR amplification conditions including thermal cycler ramp speeds. The complex patterns of RAPD also prevent mixture interpretation and provide challenges in consistent scoring of electrophoretic images even in single-source samples (Wikstro¨m, Ann-Christin, and Forsman, 1999). The reproducibility of the method was evaluated, and its ability to monitor overall changes in complex microbial communities was demonstrated. 8.3.2.1.7 Stable-isotope probing

Microbial biodegradation and biotransformation reactions are essential to most bioremediation processes, yet the specific organisms, genes, and mechanisms involved are often not well understood (Dumont & Murrell, 2005). Stable-isotope probing (SIP) is a technique that is used to identify the microorganisms in environmental samples that use a particular growth substrate. The method relies on the incorporation of a substrate that is highly enriched in a stable isotope, such as 13C, and the identification of active microorganisms by the selective recovery and analysis of isotope-enriched cellular components. DNA and rRNA are the most informative taxonomic biomarkers and 13 C-labeled molecules can be purified from unlabeled nucleic acid by density-gradient centrifugation. SIP enables researchers to directly link microbial metabolic capability to phylogenetic and metagenomics information within a community context by tracking isotopically labeled substances into phylogenetically and functionally informative biomarkers (Uhlik et al., 2013). SIP has been coupled with nucleic acid methods to provide a culture-independent techniques of linking the identity of bacteria with their function in the environment. Soil is either incubated after adding a 13C-labeled substrate or a plant is labeled with 13C-CO2. Afterward soil DNA or RNA is extracted and centrifuged in a density gradient to separate 13C-labeled nucleic acids from the environmental sample. When separated, labeled DNA can be amplified using PCR. Analysis of the PCR products, through cloning and sequencing, for example, allows the microbes that have assimilated the labeled substrate to be identified. The development of SIP was instrumental in circumventing the limitations of culture-based investigations of biodegradation (Friedrich, 2006; Wellington, Berry, & Krsek, 2003). The coupling of SIP with the rapidly advancing field of metagenomics is further increasing its potential benefits to the field of bioremediation. SIP-based approaches hold great potential for linking microbial identity with function, but at present, a high degree of labeling is necessary to be able to separate labeled from unlabeled marker molecules. Jiang, Song, Luo, Zhang, and Zhang (2015) identified a novel phenanthrene-degrading bacteria by DNA stable-isotope probing (DNA-SIP) techniques. In this study, the soil was artificially amended with either 12C or 13 C-labeled phenanthrene (PHE), and soil DNA was extracted on days 3, 6, and 9.

Metagenomics in remediation of contaminated sites 211 T-RFLP results revealed that the fragments of 219- and 241-bp in HaeIII digests were distributed throughout the gradient profile at three different sampling time points, and both fragments were more dominant in the heavy fractions of the samples exposed to the 13 C-labeled contaminant. 16S rRNA sequencing of the 13C-enriched fraction suggested that Acidobacterium spp. and Collimonas spp. were directly involved in the uptake and degradation of PHE at different times. SIP enabled cultivation of the indigenous strain Ralstonia sp. M1 capable of degrading PHE and biphenyl (BP) in industrial wastewater (Li et al., 2019). To identify and obtain the indigenous degraders metabolizing PHE and BP from the complex microbial community within industrial wastewater, DNA-SIP and cultivation-based methods were applied in this study. DNA-SIP results showed that Vogesella and Alicyclobacillus were considered to be the key biodegraders responsible for PHE biodegradation only, whereas Bacillus and Cupriavidus were involved in BP degradation. Additionally, DNA-SIP helps reveal the taxonomic identity of Ralstonia-like degraders involved in both PHE and BP degradation. To target the separation of functional Ralstonia-like degraders from the wastewater, authors modified the traditional cultivation medium and culture conditions. Finally, an indigenous PHE and BP-degrading strain, Ralstonia pickettii M1, was isolated via the cultivation-dependent method and its role in PHE and BP degradation was confirmed by enrichment of the 16S rRNA gene and distinctive dioxygenase genes in the DNA-SIP experiment 8.3.2.1.8 Quantitative polymerase chain reaction

qPCR, also referred to as real-time PCR (RT-PCR) or real-time quantitative PCR (RT-qPCR) was developed as a precise, efficient, and rapid method for nucleic acid detection. qPCR has emerged as a promising tool for studying soil microbial communities. RT-qPCR is a major development of PCR technology that enables the reliable detection and measurement of products generated during each cycle of the PCR process. This technique is based on traditional PCR technology with a few improvements: RT-PCR combines nucleic acid amplification and detection in one step; It can use smaller amounts of a starting material than traditional PCR. Also it is possible to quantify the product based on fluorescent detection. The detection of template DNA (or RNA if a reverse transcription step is added prior to amplification) is live and based on when the PCR product is amplified above the threshold of background or the cycle threshold number (Ct number). Far different from the former end point method of standard PCR techniques, RT-PCR can be quantitative because the PCR product is detected using fluorescent dyes in real time. Yergeau, Sanschagrin, Beaumier, and Greer (2012) used reverse-transcriptase RT-qPCR to quantify the expression of several hydrocarbon-degrading genes. Pseudomonas species appeared as the most abundant organisms in alert soils right after contamination with diesel and excavation (t 5 0) and one month after the start of the bioremediation treatment (t 5 1 m), when degradation rates were at their highest, but decreased after 1 year (t 5 1 year), when residual soil hydrocarbons were almost depleted. RT-qPCR assays confirmed that

212 Chapter 8 Pseudomonas and Rhodococcus species actively expressed hydrocarbon degradation genes in Arctic biopile soils. Lu et al. (2015) used 454 pyrosequencing, Illumina high-throughput sequencing, and metagenomics analysis to investigate bacterial pathogens and their potential virulence in a sewage treatment plant applying both conventional and advanced treatment processes. Pyrosequencing and Illumina sequencing consistently demonstrated that Arcobacter genus occupied over 43.42% of total abundance of potential pathogens in sewage treatment plant (STP). At the species level, potential pathogens Arcobacter butzleri, Aeromonas hydrophila, and Klebsiella pneumoniae dominated in raw sewage, which was also confirmed by qPCR. Rahm et al. (2006) characterized microbial populations at two sites with differing reductive dechlorination abilities. qPCR targeting the 16S rRNA gene of Dehalococcoides strains, known for their unique capability to dechlorinate solvents completely to ethene, revealed a significant population at National Engineering and Environmental Laboratory (INEEL), but no detectable population at Seal Beach prior to bioaugmentation. Detection of Dehalococcoides by qPCR correlated with observed dechlorination activity and ethene production at both sites. qPCR showed that Dehalococcoides was present in even the pristine well at INEEL, suggesting that the difference in dechlorination ability at the two sites was due to the initial absence of this genus at Seal Beach. 8.3.2.2 Nonpolymerase chain reaction based molecular techniques 8.3.2.2.1 Fluorescence in situ hybridization

Fluorescence in situ hybridization (FISH) effectively extends epifluorescence microscopy, allowing for the fast detection and enumeration of specific microorganisms. This method uses fluorescent-labeled oligonucleotides probes (usually 15 25 bp) which bind specifically to microbial DNA in the sample, allowing the visualization of the cells using an epifluorescence or confocal laser scanning microscope. FISH has been successfully used to characterize microorganisms within biofilms and to detect pathogens in drinking water samples (Batte´, Laurent, & Prevost, 2003; Wilhartitz et al., 2007). An improvement of the FISH method is the catalyzed reporter deposition fluorescence in situ hybridization (CARD-FISH) (Pernthaler, Pernthaler, & Amann, 2002). CARD-FISH is useful when dealing with drinking water samples since it can enhance the fluorescent signal from cells in samples with low microbial concentration (Dorigo et al., 2005). In general, FISH is not used as a standalone technique and is mostly used in combination with other methods to characterize microbial communities. An example of these combined techniques is high-affinity peptide nucleic acid-FISH, useful to study pathogens in biofilms due to the enhanced capability of the probe to penetrate through the extracellular polymeric substance matrix (Lehtola et al., 2007). Despite its numerous advantages when compared with culture-dependent techniques, FISH also has several limitations. First of all, knowledge of the nucleotide sequence of

Metagenomics in remediation of contaminated sites 213 the target organisms is needed and the design of new probes and the optimization of the hybridization conditions can be time-consuming and complex. The efficiency of the hybridization might be influenced by the physiological state of the cells and the signal emitted by autofluorescence cells can interfere with the signal emitted by the target microorganisms (Dorigo, Volatier, & Humbert, 2005). The structure of sedimentassociated microbial communities along a heavy-metal contamination gradient in the marine environment has been explored by Gillan, Danis, Pernet, Joly, and Dubois (2005) in marine sediments contaminated for .80 years with Cd, Cu, Pb, and Zn. The abundances of seven phylogenetic groups were determined by using FISH. The results showed that the HCl-extractable Cu, Pb, and Zn were negatively correlated with the abundance of λ-Proteobacteria and Cytophaga/Flavobacterium/Bacteroides (CFB) bacteria. λ-Proteobacteria were not correlated with HCl-extractable metals. Bacteria of the Desulfosarcina Desulfococcus group were detected in every site and represented 6% 14% of the 4,6 diamidino-2-phenylindole counts. Although factors other than metals may explain the distribution observed, the information presented here may be useful in predicting long-term effects of heavy-metal contamination in the marine environment. An anaerobic microbial consortium (referred to as ANAS) that reductively dechlorinates TCE completely to ethene with the transient production of cisdichloroethene (cDCE) and VC was enriched from contaminated soil (Richardson, Bhupathiraju, Song, Goulet, & Alvarez-Cohen, 2002). Three molecular methods were used concurrently to characterize the community structure of ANAS and VCC: clone library construction/clone sequencing, T-RFLP analysis, and FISH with rRNA probes. FISH results suggest that members of the CFB cluster and high guanine plus cytosine (G 1 C) gram-positives were numerically important in ANAS despite their underrepresentation in the clone libraries. However, the nucleic acid-based analyses performed here would need to be supplemented with chemical species data in order to test any hypotheses about the functional roles of various community members. Additionally, these results suggest that an organism outside the Dehalococcoides genus may be capable of dechlorinating cDCE to VC. The microbial communities of three different sulfidic and acidic mine waste tailing dumps located in Botswana, Germany, and Sweden were quantitatively analyzed using qPCR, FISH, catalyzed reporter deposition-FISH (CARD-FISH), Sybr Green II direct counting, and the most probable number (MPN) cultivation technique. The FISH analysis provided reliable data only for tailing zones with high microbial activity, whereas CARD-FISH, qPCR, Sybr Green II staining, and MPN were suitable methods for quantitative microbial community analysis of tailings in general (Dagmar Kock & Axel Schippers, 2008). 8.3.2.2.2 Guanine plus cytosine content

The G 1 C content of DNA can be used to study the bacterial diversity of soil communities (Nusslein & Tiedje, 1999). It is based on the knowledge that microorganisms differ in their

214 Chapter 8 G 1 C content and that taxonomically related groups only differ between 3% and 5% (Tiedje et al., 1999). This method provides a coarse level of resolution as different taxonomic groups may share the same G 1 C range. The advantages of G 1 C analysis are that it is not influenced by PCR biases, it includes all DNA extracted, it is quantitative, and it can uncover rare members in the microbial populations. It does, however, require large quantities of DNA (up to 50 ng) (Tiedje et al., 1999). 8.3.2.2.3 Microarrays

High-throughput technologies are urgently needed for monitoring the formidable biodiversity and functional capabilities of microorganisms in the contaminated environment. Microarrays, also known as DNA chips or biochips, are a powerful genomic technology that is widely used by the scientific community to monitor gene expression under different cell growth conditions, detect specific mutations in DNA sequences, and identify and characterize microbial communities in different microbial ecosystems (Govindarajan, Duraiyan, Kaliyappan, & Palanisamy, 2012). DNA microarrays initially were developed for transcriptome analyses of Arabidopsis cells (Schena, Shalon, Davis, & Brown, 1995). DNA microarrays have revolutionized our ability to simultaneously carry out hundreds or thousands of hybridization reactions at a time. In principle, a DNA microarray is a glass microscope slide onto which thousands of DNA probes targeting genes or gene products of interest can be immobilized on a solid support. Afterward, DNA or mRNA is extracted from cells or tissues, labeled with specific fluorescent molecules, and hybridized to the spotted DNA on the glass slide. The resulting image of the hybridization signal of each probe (fluorescent spots) is visualized by a confocal scanner and digitized for quantitative analysis (Tarca, Roberto Romero, & Draghici, 2006). The DNA microarray approach was first introduced by David Stahl and his colleagues in 1997 to explore the microbial community composition of an environmental sample using a prototype array consisting of nine 16S rRNA-targeted probes for identification of selected nitrifying bacteria (Guschin et al., 1997). The genes encoding functional enzymes involved in biogeochemical cycling processes are very valuable signatures for monitoring the physiological and functional activities of microbial communities in natural environments. The two formats of microarrays, that is, (1) PhyloChip and (2) functional gene arrays (FGAs), have been developed and evaluated for detection and microbial community analyses in complex environments. In PhyloChips (also known as phylogenetic oligonucleotide arrays) target PCR-amplified rRNA gene fragments or directly retrieved community rRNA (genes) can be designed to detect any microorganism. In contrast, microarrays containing functional gene sequences or gene families that encode key enzymes that are diagnostic for a certain metabolic pathway are referred to as FGAs and primarily used for the functional analysis of microbial community activity in the environment (Taroncher-Oldenburg, Griner, Francis, & Ward, 2003; Wu et al., 2001). Recently several pioneering studies were published, which use PhyloChips or FGAs also for functional analysis of microbial communities. To explore the microbial response to heavy-metal

Metagenomics in remediation of contaminated sites 215 contamination (e.g., Cr, Mn, and Zn), the composition, structure, and functional potential of a sedimentary microbial community were investigated by the sequencing of 16S rRNA gene amplicons and a FGAs (Yin et al., 2015). Analysis of 16S rRNA sequences revealed that the composition and structure of sedimentary microbial communities changed significantly across a gradient of heavy-metal contamination, and the relative abundances were higher for Firmicutes, Chloroflexi, and Crenarchaeota, but lower for Proteobacteria and Actinobacteria in highly contaminated samples. Also molecular ecological network analysis of sequencing data indicated that their possible interactions might be enhanced in highly contaminated communities. Correspondingly, key functional genes involved in metal homeostasis (e.g., chrR, metC, and merB), carbon metabolism, and organic remediation showed a higher abundance in highly contaminated samples, indicating that bacterial communities in contaminated areas may modulate their energy consumption and organic remediation ability.

8.4 Metagenomics: a culture-independent insight Microorganisms, the most abundant and diverse group of life on Earth, play a major role in the biogeochemical cycling of compounds highly essential for the functioning of the ecosystem. Several clever cultivation methods have been devised to expand the range of organisms that can be cultured, but knowledge of the uncultured world is slim, so it is difficult to use a process based on rational design to coax many of these organisms into the culture. Metagenomics (also referred to as population genomics, environmental genomics, ecogenomics, and community genomics), is a new field with an additional set of highthroughput sequencing tools, offers an approach to study the diversity, adaptation, and evolution of microorganisms surviving in contaminated sites. This method is unique because it is not dependent on the cloning and sequencing of particular genes. Instead, it provides a detailed survey of all the genes that exist within a particular community, lending insight to both composition and function in a single experiment. The term metagenomics was coined by Handelsman, Rondon, Brady, Clardy, and Goodman (1998). They have accessed the collective genomes and the biosynthetic machinery of soil microflora during a study of cloning the metagenome. The term “metagenomics” is also liberally applied in the literature to a third approach: the thorough analysis of the diversity of specific genes— primarily marker genes such as the 16S rRNA gene (Wang & Qian, 2009) or conserved single-copy genes (Lang, Darling, & Eisen, 2013)—through PCR amplification and highthroughput sequencing of the amplicons. In metagenomics the first step consists of extracting total genomic DNA (metagenome) from the environment, which is often challenging in polluted soils, water, and sediments. Indeed, a high concentration of pollutants (e.g., metals and aromatic hydrocarbons) and low cell density are the main factors that hamper successful DNA recovery. Metagenomics analysis of genomic DNA is based on four approaches (1) genetic and/or functional screening of cloned DNA (referred to as a metagenomics library); (2) large-scale sequencing of metagenomes without

216 Chapter 8 pre-cloning (referred to as “direct sequencing” or “shotgun metagenomics” or shotgun metagenome sequencing); (3) profiling of RNAs and proteins produced by a microbiome (metatranscriptomics and metaproteomics); and (4) identification of a community’s metabolic network (metabolomics) which targets different aspects of the local microbial community associated with a determined environment. Together these approaches have increased our understanding of the unculturable microbial world and have therefore also provided insights into the prokaryotic world that is otherwise obscure. For the metagenomic library preparation, metagenomics entails the extraction of DNA from a community so that all of the genomes of organisms in the community are pooled. These genomes are usually fragmented and cloned into an organism that can be cultured to create “metagenomics libraries,” and these libraries are then subjected to analysis based on DNA sequence or on functions conferred on the surrogate host by the metagenomic DNA. This library-based metagenomic method is also known as library-based targeted metagenomics because it combines the cloning of environmental metagenomic DNA, screening of the clones for a function of interest (sequence-based or function-based screening), and high-throughput sequencing of selected clones. Shotgun metagenomic sequencing allows researchers to evaluate bacterial diversity and detect the abundance of microbes in a complex environment, without culturing the microorganisms. In all cases, metagenomic DNA is then sequenced by one of the presently available high-throughput sequencing platforms (e.g., Roche 454, Illumina, Ion Torrent, Oligonucleotide Ligation Detection (SOLiD), Heliscope, PacBio, and Oxford Nanopore). Lastly, the huge amount of data obtained by sequencing platform is analyzed using a panoply of bioinformatic tools to predict the diversity of the microbial community and/or the functional potential of the microbial community in contaminated environment. There are several in silico softwares, pipelines, web resources, and algorithms that are being used to interpret or correlate the sequencing data (Desai, Pathak, & Madamwar, 2010). Some important web resources such as NCBI Genome, GOLD genome online database, MBGD (Microbial Genome Database), IMG (Integrated Microbial Genomes & Microbiomes), MG-RAST (Metagenomics analysis server), PATRIC, WebMGA, eggNOG (Evolutionary genealogy of genes: Non-supervised Orthologous Groups), FOAM (Functional Ontology Assignments for Metagenomes), SILVA, Rfam, NCBI reference sequences, BSRD, EcRBPome, GreenGene, CAMERA, EnvDB, EnsemblBacteria and computational and bioinformatics tools/software, viz., UCLUST, FUNGIpath, UCHIME, MEGAN, Prodigal, COGNIZER, WebCARMA, UniFrac, JCVI METAREP, and CARMA, are currently used in the analysis of generated metagenomics data.

8.4.1 Functional-based metagenomics The function-based metagenomic approach allows the discovery of novel genes that code for a function of interest, which involves the generation of expression libraries with

Metagenomics in remediation of contaminated sites 217 thousands of metagenomic clones followed by activity-based screenings and metagenomic sequencing. Function-based metagenomics, unlike sequence-driven approaches, does not require that genes have homology to genes of known function, making it the only approach to metagenomics that has the potential to identify entirely new classes of genes for new or known functions. It offers the opportunity to add functional information to the nucleic acid and protein databases. In the functional-based metagenomic analysis, DNA fragments are cloned, expressed in a laboratory host, and screened for enzymatic activities. To clarify how metagenomics can be applied to bioremediation applications and environmental monitoring, we will discuss three case studies that exemplify the application of these techniques. Metagenomic functional analysis of a biostimulated petroleum-contaminated soil was performed by Terro´n-Gonza´lez, Martı´n-Cabello, Ferrer, and Santero (2016) to estimate the relative abundance and diversity of extradiol dioxygenases (Edo), which are key enzymes in the biodegradation of aromatic contaminants. The library, consisting of 6.5 Gb of metagenomic DNA, was screened for Edo activity using catechol and 2,3dihydroxybiphenyl as the substrates. Fifty-eight independent clones encoding Edo activity and forty-one different Edo-encoding genes were identified.

8.4.2 Sequence-based metagenomics Sequence-based metagenomics is used to assemble genomes, identify genes, find complete metabolic pathways, and compare organisms of different communities without culturing them. It can lead to a better understanding of how certain genes help organisms to survive in a contaminated environment. Some studies use a gene of interest or “anchor” to identify metagenomic clones of interest for further analysis. A metagenomic library is constructed and screened using PCR to amplify the anchor. Anchors are often a ribosomal RNA gene, but can also be a metabolic gene (e.g., a polyketide synthase). The clones that contain the anchor are then sequenced or further analyzed to provide information about the genomic context of the anchor. Thus researchers can quickly focus on a clone of interest. In comparison to functional screening, the sequence-based approach depends on the sequence analysis to provide a basis for function prediction. Sequence-based metagenomic analysis can be used for gene identification, genome assemblages, clarifying complete metabolic pathways, and comparing organisms from different communities. Joshi et al. (2014) used a metagenomic approach for understanding the microbial population in petroleum muck. This metagenomic study was based on next-generation sequencing (NGS) using the Ion Torrent platform. The taxonomic analysis revealed the predominance of the eubacteria domain (88.90%), followed by eukarya (0.06%) and archaea (0.03%). Sequences affiliated with the phylum Proteobacteria (99.09%) were most abundant, with Gammaproteobacteria (51.31%) as the major class and Pseudomonas stutzeri as the most abundant species able to metabolize benzoate, cresol, naphthalene, xylene, toluene, and phenol. Other sequences belonged to the phyla Actinobacteria (0.70%), Firmicutes (0.11%), and 0.75% other phyla.

218 Chapter 8 Although sequences were not affiliated with domain archaea as much as other major phyla, Euryarchaeota, Thaumarchaeota, and Crenarchaeota, have developed mechanisms of metal resistance and thus can be used in bioremediation.

8.4.3 Metatranscriptomics Metatranscriptomics is the study of the gene expression (mRNA) and abundance (rRNA) of a microbial community grown in a complex environment (Aguiar-Pulido et al., 2016; Poretsky, Bano, & Buchan, 2005; Tveit, Urich, & Svenninga, 2014). Metatranscriptomics commonly involves reverse transcription to generate cDNA, which can then be sequenced using the same platforms as for metagenomics. Direct RNA sequencing, bypassing cDNA generation and its associated biases, is also available but has not yet been employed in the context of mixed microbial communities. In contrast to proteins, which have a more stable concentration and longer lifetime in the cell in response to external influences, mRNAs provide a more immediate picture of the cells’ responses to changing environmental conditions. As the majority of RNA in a cell is composed of ribosomal and transfer RNAs ( . 95%), metatranscriptomics typically comprises rRNAs depletion steps to enrich for mRNAs. Several challenges, such as the isolation of high-quality RNA samples from some environmental samples, short half-life, and separation of mRNA from other RNA species, are associated with in situ metatranscriptomics analysis of microbial communities during the bioremediation of environmental contaminants (Bashiardes, Zilberman-Schapira, & Elinav, 2016; Simon & Daniel, 2011). However, multiple metagenomic analysis methods may at times produce variable results, even if identical databases are used in the analysis. Thus standardization of RNA isolation, processing, sequencing, and analysis is warranted to enable further dissemination of metatranscriptomics methods and their integration into microbiome research. Whole metagenome sequencing approaches provide information on the taxonomic profile of a microbial community as well as its potential functional profile; in contrast, whole metatranscriptome sequencing describes the active functional profile. A metagenomic and metatranscriptomic study conducted by Men et al. (2017) revealed the structure and dynamics of a dechlorinating community containing Dehalococcoides mccartyi and Corrinoid providing microorganisms under cobalamin-limited conditions. This study provides important insights into the microbial interactions and roles played by members of dechlorinating communities under cobalamin-limited conditions.

8.4.4 Metaproteomics Metaproteomics, also known as community proteomics, environmental proteomics, or community proteogenomics, is the study of all the protein samples recovered from environmental sources. Metaproteomics offers a powerful approach to link microbial community composition to function. However, the success of metaproteomics is strongly

Metagenomics in remediation of contaminated sites 219 dependent on the availability of relevant genomes to enable high protein identification rates. The technological requirements for proteomic measurements include large dynamic range, sensitive protein/peptide detection, accurate mass measurements, high-throughput processing, the ability to deal with very complex mixtures, and the ability to structurally characterize (and resolve) peptide sequences. In this regard, mass spectrometry has emerged as the unchallenged leader in the field, becoming the dominant technological platform for almost all proteomic measurements. This method provides a new way of identifying functional proteins in microorganisms and studying their metabolic activity in complex biological pathways. However, metaproteomics presents some valuable advantages over metabolomics as proteins can be assigned to specific taxa and therefore their detection informs not only on what metabolic pathways are active within an ecosystem but also on the identity of species involved in specific functions. The metaproteomic analysis revealed that bacterial functional proteins, such as ferredoxin- nicotinamide adenine dinucleotide phosphate reductase, acetate kinase, and NADH-quinone oxidoreductase, mainly belonging to phyla Firmicutes, Proteobacteria, Actinobacteria, and Bacteroidetes, are involved in carbohydrate metabolism, energy metabolism, lipid metabolism, and amino acid metabolism. The archaeal functional proteins are mainly involved in methane metabolism in energy metabolism, such as acetyl-CoA decarboxylase, and methyl-coenzyme M reductase, and the acetic acid pathway exhibited the highest proportion of the total. The genus Methanosarcina presents the highest functional diversity in methane metabolism and can produce methane under the influence of multifunctional proteins through acetic acid, CO2 reduction, and methyl nutrient pathways. The study demonstrates metaproteomics as a new way of uncovering community functional and metabolic activity (Jia, Xi, Li, Yang, & Wang, 2017).

8.4.5 Metabolomics Metabolomics is the systematic study of all small low-molecular-weight molecules (metabolites) within a biological system. Metabolites are typically in a state of flux, which implies that their compositions and concentrations vary significantly as a function of time within an ecosystem. The abundance of some compounds (e.g., sugars and amino acids) results from uptake, consumption, and excretion by many different organisms; thus the overall concentration reflects the net metabolic state of the microbial community. Since it provides direct biochemical observation of the community metabolism, metabolomics offers a powerful approach for the characterization of ecosystem phenotypic and traits. To investigate community-level adaptations to the simultaneous challenges of high proton and metal concentrations, authors examined the metabolome of microbial biofilm communities in an acid mine drainage (AMD) environment (Richmond Mine, Iron Mountain, CA). The authors identified key metabolites associated with acidophilic and metal-tolerant microorganisms using stable isotope labeling (SIL) coupled with untargeted, high-resolution

220 Chapter 8 mass spectrometry. They observed .3500 metabolic features in biofilms growing in pH B0.9 AMD solutions containing millimolar concentrations of Fe, SO422, Zn, Cu, and As. SIL improved chemical formula prediction by .50% for larger metabolites ( . 250 atomic mass units), many of which were unrepresented in metabolic databases and may represent novel compounds. Unlike other metaomics, metabolites and metabolic pathways are relatively conserved across species. As metabolic products reflect the interactions between the cell’s genome and its environment, metabolomics provides an unbiased assessment of a cellular state within the context of that particular condition. Since concentrations of intracellular metabolites often reveal aspects of biochemical regulations that are undetectable by other approaches, metabolomics fills in the gaps of the more traditional studies of interactions between genes, proteins, and metabolites in individual cells. Metabolomics has already demonstrated its critical role in bioenergy, environmental interactions, functional genomics and gene discovery, secondary metabolism, genome-wide association mapping, and metabolic modeling in the higher organism and microbial systems. The main challenge of metabolomics is largely technical—the ability to identify and quantify the entire set of intracellular and extracellular metabolites with a molecular mass lower than 1000 daltons. The numbers of these compounds vary among different organisms, from hundreds to hundreds of thousands, and in many cases their identity may be unknown. In contrast to genome, transcriptome, and proteome analyses, products generated from metabolic reactions are highly variable in their chemical structures and properties.

8.4.6 Metagenomics sequencing strategies Genome sequencing technologies have been frequently upgraded since the completion of the human genome at the beginning of the 21st century. Multiple next-generation genomic sequencing strategies have been applied to sequence the metagenomes of different microbial communities. Sequencing technologies were initiated by the Sangers sequencing method, which was widely used during the process of human genome sequencing. Technological drift has gifted NGS techniques like pyrosequencing, ligation sequencing, reverse terminator, and single-molecule sequence by synthesis, providing a high throughput that can read in comparatively less time. A comparative overview of recent metagenomics studies of contaminated environment carried out by various group of researchers are provided in Table 8.1 for a more detailed understanding. However, most metagenomics researchers prefer the pyrosequencing method for sequencing the metagenomes of microbial communities.

8.5 Next-generation sequencing technologies to explore structure and function of microbial communities The application of culture-independent approaches based on the analysis of 16S rRNA and function genes including FISH, DGGE, T-RFLP, and qRT-PCR, microarrays, and clone

Table 8.1: Examples of metagenomics analysis of contaminated sites. S. No. Most abundant taxa/organism

Analysis

1

Metagenomics

2 3 4

5

6 7

Pseudomonas species, Gammaproteobacteria, Alphaproteobacteria, and Actinobacteria Proteobacteria, Sulfuricella, and Thiobacillus Dehalococcoides, Dehalobacter, and Desulfitobacterium Synergistetes, Verrucomicrobia, Cyanobacteria, Actinobacteria, Fusobacteria, Planctomycetes, Chloroflexi, Spirochaetes, TM7, Tenericutes, Acidobacteria, Lentisphaerae, and Euryarchaeota Proteobacteria, Actinobacteria, Chloroflexi, Verrucomicrobia Acidobacteria, Bacteroidetes, Planctomycetes, Cyanobacteria Firmicutes, Gemmatimonadetes, and TM7 γ-Proteobacteria and α-Proteobacteria Thermi, Gemmatimonadetes, Proteobacteria, Bacteriodetes, and Actinobacteria

8 9 10

Proteobacteria and Actinobacteria Bacilli

11

Arcobacter butzleri, Aeromonas hydrophila, and Klebsiella pneumonia Actinobacteria and Proteobacteria, Robiginitalea, Microlunatus, and Alicyclobacillus Proteobacteria, Actinobacteria, Firmicutes, Acidobacteria, Gemmatimonadetes, Chloroflexi, Bacteroidetes, and Nitrospirae Firmicutes, Actinobacteria, Proteobacteria Chloroflexi, Bacteroidetes, Synergistetes, and Acidobacteria Proteobacteria, Bacteroidetes, Firmicutes, Nitrospira, Acidobacteria, Actinobacteria, and Planctomycetes Firmicutes, Proteobacteria, Bacteroidetes, and Actinobacteria Alphaproteobacteria, Betaproteobacteria Gammaproteobacteria, Clostridia, Sphingobacteria, Flavobacteria, Bacilli, Flavobacteria, and Enterococcus aquimarinus

12 13 14 15 16 17

Contaminated site/Matrix Diesel-contaminated Arctic soil

Reference Yergeau et al. (2012)

Metagenomics Cd-contaminated soil Ecogenomics Chlorinated contaminated Contaminated sediments

Feng et al. (2018) Maphosa et al. (2012) Sahu et al. (2018)

Metagenomics Cu-contaminated sites

Jiang et al. (2016)

Metagenomics Petroleum-contaminated soil Metagenomics Hydrocarbon-contaminated soil

Bao et al. (2016) Kumar et al. (2018)

Metagenomics Petroleum hydrocarboncontaminated site Metagenomics Diesel-contaminated microcosms Wastewater treatment plant (Pharmaceuticals)

Mukherjee et al. (2017)

Metagenomics Advanced sewage treatment systems Metagenomics Uranium-contaminated soil Metagenomics Paddy fields under mixed heavymetal contamination Metagenomics Pharmaceutical-enriched wastewater Coal-mine wastewater treatment plants Municipal dumpsite Crude oil and sludge

Jung et al. (2016) Balcom, Driscoll, Vincent, and Leduc (2016) Lu et al. (2015) Yan, Luo, and Zhao (2016) Huaidong et al. (2017) Zhang, Luo, and Lee (2016) Ma et al. (2015) Mwaikono et al. (2016) Albokari, Mashhour, Alshehri, Boothman, and Al-Enezi (2015)

222 Chapter 8 libraries have been conducted to analyze the microbial community structure and gene diversities in various environments. These methods have led to new insights into microbial processes in a contaminated environment and can provide direct evidence for the presence of specific microorganisms influencing pollutant degradation and/or reduction. Although those methods are still useful for lower-diversity communities, those methods may not integrally reflect microbial diversity and couple microbial taxonomy diversity with diversified functions due to low throughput. In recent years, the newly developed highthroughput sequencing technology has been successfully applied in this field which can afford us a huge amount of information to identify the entire profile of microbial communities. High-throughput sequencing methods, viz., 454 pyrosequencing and Illumina sequencing technologies, have become effective tools to fully explore the microbial diversity in the environment (Roesch et al., 2007; Roh et al., 2010). This method has been applied recently as a novel promising method to investigate the genes and genes expression levels of the microbial community in different ecosystems. DNA-based high-throughput sequencing metagenomics has been applied to reveal microbial communities in contaminated soil and water. At the beginning of the metagenomic studies, the use of Sanger sequencing technology provided important progress in the field. However, the advent of NGS technologies capable of sequencing millions of DNA fragments simultaneously, at a low cost, greatly bolstered the field. Comparatively, NGS platforms can recover up to 5000 Mb of DNA sequence per day with costs at about 0.50 $/Mb, while Sanger sequencing methodology allows about 6 Mb of DNA sequence to be created per day with costs about 1000 times higher. Unlike PCR-based approaches, NGS allows researchers to sequence thousands of organisms in parallel. With the ability to combine many samples in a single sequencing run and obtain high sequence coverage per sample, NGS-based metagenomic sequencing can detect very low abundance members of the microbial community that may be missed or are too expensive to identify using other methods. Advances in NGS, enabling massive parallel analysis of DNA sequence information from PCR amplicons, or environmental nucleic acids, open up a new era of proxy development for environmental decontamination and eco-restoration. Kachienga et al. (2018) profiled the indigenous microbial diversity of two South African petroleum-contaminated water aquifer sites and determined the microbial adaptation to hydrocarbon degradation using a metagenomic approach. The sequenced samples revealed that protozoa (62.04%) were found to be the most dominant group, followed by fungi (24.49%), unknown (12.87%), and finally other sequences such as Animalia and Plantae which were ,0.10% domains in the first oil-polluted aquifer site. While in the second site, a protozoon (61.90%) was found to be the most dominant group followed by unknown (16.51%), fungi (11.41%). According to the classification at the genus level, the dominant group was Naegleria (15.21%), followed by Vorticella (6.67%) as the only ciliated protozoan genus, other species such as Arabidopsis (2.97%), Asarum (1.84%), Populus (1.04%) were significantly low and drastically lower in the first site. Regarding the second site, the dominant group was

Metagenomics in remediation of contaminated sites 223 Naegleria (18.29%) followed by Colpoda (9.86%) with the remainder of the genera representing ,2%. Overall results demonstrated the ability of various groups of microorganisms to adapt and survive in petroleum oil-polluted water sites regardless of their respective distributions and play a major role in bioremediation and environmental management. Sangwan, Lata, Dwivedi, Singh, and Niharika (2012) characterized the microbial community responsible for the in situ bioremediation of hexachlorocyclohexane (HCH) using a metagenomic approach. Microbial community structure and function were analyzed using 16S rRNA amplicon and shotgun metagenomic sequencing methods for three sets of soil samples. Certain bacterial (Chromohalobacter, Marinimicrobium, Idiomarina, Salinosphaera, Halomonas, Sphingopyxis, Novosphingobium, Sphingomonas, and Pseudomonas); archaeal (Halobacterium, Haloarcula and Halorhabdus); and fungal (Fusarium) genera were found to be more abundant in the soil sample from the HCHdumpsite. Consistent with the phylogenetic shift, the dumpsite also exhibited a relatively higher abundance of genes coding for chemotaxis/motility, chloroaromatic and HCH degradation (lin genes). Abed, Al-Kharusi, Gkorezis, Prigent, and Headley (2018) described the bacterial diversity, using molecular (Illumina MiSeq sequencing) and cultivation techniques, in the rhizospheric soil of Phragmites australis from oil-polluted wetland in Oman. The obtained isolates were tested for their plant growth-promoting properties. All strains were able to solubilize phosphate and about half were capable of producing organic acids and 1-aminocyclopropane-1-carboxylate deaminase. Around 42% of the strains had the ability to produce indole acetic acid and siderophore. A total of 455,201 16S rRNA sequences were generated by MiSeq sequencing. The most sequences belonged to Proteobacteria, Bacteriodetes, and Firmicutes. Sequences of potential hydrocarbon-degrading bacteria (e.g., Ochrobactrum and Pseudomonas) were frequently encountered. All soils contained sequences of known sulfur-oxidizing (e.g., Thiobacillus, Thiofaba, Rhodobacter, and Sulfurovum) and sulfate-reducing bacteria, although the latter group made up only 0.1% to 3% of total sequences. The obtained isolates from the rhizosphere soils were phylogenetically affiliated to Serratia, Acinetobacter, Xenorhabdus, Escherichia, and Salmonella. The authors conclude that the rhizosphere soils of P. australis in oil-polluted wetlands harbor diverse bacterial communities that could enhance the wetland performance through hydrocarbon degradation, nutrient cycling, and supporting plant growth. Huaidon and his colleagues used the highthroughput Illumina MiSeq sequencing approach to explore the bacterial diversity and community composition of soils in four paddy fields, exhibiting four degrees of mixed heavy metal (Cd, Pb, and Zn) pollution, and examined the effects of these metals on the bacterial communities. Our results showed that 2104 4359 bacterial OTUs were found in the bulk and rhizosphere soils of the paddy fields, with the dominant bacterial phyla (greater than 1% of the overall community) including Proteobacteria, Actinobacteria, Firmicutes, Acidobacteria, Gemmatimonadetes, Chloroflexi, Bacteroidetes, and Nitrospirae (Huaidong et al., 2017).

224 Chapter 8

8.6 Conclusion Microbes in nature usually exist in complex communities, which can be highly changeable in both the abundance and the composition of their constituent species. Interest in the microbial-based bioremediation approaches of environmental pollutants has increased in recent years, as people endeavor to find sustainable ways for the remediation of polluted environments. Since microbes are the drivers of bioremediation, shifts in the composition and activity of a microbial community may impact the fate of a contaminant in the environment. Increased understanding of how microbial communities cope with pollutants could help to assess the potential of contaminated sites to recover from pollution and increase the chances of bioaugmentation or biostimulation trials to succeed. Molecular analysis of environmental communities have revealed that only 1% of the total number of prokaryotic species present in given environmental samples are readily cultivable under standard laboratory conditions, and therefore a majority of microbes in the environment are not readily accessible for the basic research of biotechnological applications. The recent development of culture-independent molecular tools based on the PCR amplification, cloning, and Sanger sequencing of the universally conserved 16S rRNA gene and another marker gene has allowed cultivation-independent investigations of the microbial communities in diverse environments, which could be exploited for the remediation of polluted environments. It is clear from the foregoing that bioremediation techniques are diverse and have proven effective in restoring sites polluted with different types of contaminants. Current sequence databases contain over a million full-length 16S rRNA sequences spanning a broad phylogenetic spectrum that can serve as a benchmark for assessing the bacterial taxa present in contaminated environments worldwide. Although the traditional 16S-cloning-and-sequencing approach has the potential to provide an in-depth view of the richness and evenness of bacterial species within a community, its application is somewhat laborious and costly, with the result that most applications have assayed on the order of only 100 sequences per sample, which may not be sufficient to fully characterize all but the simplest communities. These low-throughput approaches are not able to completely reveal the detailed microbial community structure due to the extremely complex communities and overwhelming genetic diversities in any environment, especially for those low abundant populations although they may play important roles in the system. This has led to the development of alternative techniques for assessing rRNA variation (f.g., ARISA, DGGE, and T-RFLP), but these were often biased and of limited value owing to the inability to cultivate many naturally occurring species. Metagenomic approaches in bioremediation aid in comprehending the characteristics of bacterial communities in different kinds of contaminated environments. Metagenomics entails extraction of DNA from a community so that all of the genomes of organisms in the community are pooled. In addition, metagenomics providing direct access to the pool of environmental genomes without the bias of cultivation, metagenomics offers the possibility to explore the vast

Metagenomics in remediation of contaminated sites 225 diversity of degradation pathways of environmental microorganisms, which remain to a large extent poorly characterized. This could lead, among others, to the design of more efficient customized strains/consortia for targeted use in bioremediation applications. Thanks to the advent of NGS, several metagenomic approaches are nowadays available to dissect the composition and function of microbial populations. Culture-independent methods are expected to be continued to be used more frequently in many fields, including bioremediation and biodegradation of environmental contaminants.

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230 Chapter 8 Pernthaler, A., Pernthaler, J., & Amann, R. (2002). Fluorescence in situ hybridization and catalyzed reporter deposition for the identification of marine bacteria. Applied and Environmental Microbiology, 68, 3094e3101. Peters, S., Koschinsky, S., Schwieger, F., & Tebbe, C. C. (2000). Succession of microbial communities during hot composting as detected by PCR-single-strand-conformation-polymorphismbased genetic profiles of small-subunit rRNA genes. Applied and Environmental Microbiology, 66, 930 936. Poretsky, R. S., Bano, N., Buchan, A., et al. (2005). Analysis of microbial gene transcripts in environmental samples. Applied and Environmental Microbiology, 71(7), 4121 4126. Rahm, B. G., Chauhan, S., Holmes, V. F., Macbeth, T. W., Sorenson, K. S., Jr., & Alvarez-Cohen, L. (2006). Molecular characterization of microbial populations at two sites with differing reductive dechlorination abilities. Biodegradation, 17, 523 534. Ramamurthy, T., Ghosh, A., Pazhani, G. P., & Shinoda, S. (2014). Current perspectives on viable but nonculturable (VBNC) pathogenic bacteria. Frontiers in Public Health, 2, 103. Rastogi, G., & Sani, R. K. (2011). Molecular techniques to assess microbial community structure, function, and dynamics in the environment. In I. Ahmad, F. Ahmad, & J. Pichtel (Eds.), Microbes and microbial technology. New York, NY: Springer. Richardson, R. E., Bhupathiraju, V. K., Song, D. L., Goulet, T. A., & Alvarez-Cohen, L. (2002). Phylogenetic characterization of microbial communities that reductively dechlorinate TCE based upon a combination of molecular techniques. Environmental Science & Technology, 36(12), 2652 2662. Roesch, L. F. W., Fulthorpe, R. R., Riva, A., Casella, G., Hadwin, A. K. M., Kent, A. D., & Triplett, E. W. (2007). Pyrosequencing enumerates and contrasts soil microbial diversity. ISME Journal, 1, 283 290. Roh, S. W., Abell, G. C., Kim, K. H., Nam, Y. D., & Bae, J. W. (2010). Comparing microarrays and nextgeneration sequencing technologies for microbial ecology research. Trends in Biotechnology, 28, 291 299. Roy, A., Dutta, A., Pal, S., Gupta, A., Sarkar, J., Saha, A. C. A., . . . Kaz, S. K. (2018). Biostimulation and bioaugmentation of native microbial community accelerated bioremediation of oil refinery sludge. Bioresource Technology, 253, 22 32. Sahu, N., Deori, M., Ghosh, I., 2018. Metagenomics study of contaminated sediments from the Yamuna River at Kalindi Kunj, Delhi, India. Genome Announcements 6, e01379-17. Sangwan, N., Lata, P., Dwivedi, V., Singh, A., Niharika, N., et al. (2012). Comparative metagenomic analysis of soil microbial communities across three hexachlorocyclohexane contamination levels. PLoS ONE, 7(9), e46219. Satokari, R. M., Vaughan, E. E., Akkermans, A. D., Saarela, M., & De Vos, W. M. (2001). Polymerase chain reaction and denaturing gradient gel electrophoresis monitoring of fecal Bifidobacterium populations in a prebiotic and probiotic feeding trial. Systematic and Applied Microbiology, 24(2), 227 231. Schena, M., Shalon, D., Davis, R. W., & Brown, P. O. (1995). Quantitative monitoring of gene expression patterns with a complementary DNA microarray. Science, 270, 467 470. Schmalenberger, A., Schwieger, F., & Tebbe, C. C. (2001). Effect of primers hybridizing to different evolutionarily conserved regions of the small-subunit rRNA gene in PCR-based microbial community analyses and genetic profiling. Applied and Environmental Microbiology, 67, 3557 3563. Schu¨tte, U. M., Abdo, Z., Bent, S. J., Shyu, C., Williams, C. J., Pierson, J. D., & Forney, L. J. (2008). Advances in the use of terminal restriction fragment length polymorphism (T-RFLP) analysis of 16S rRNA genes to characterize microbial communities. Applied Microbiology and Biotechnology, 80(3), 365 380. Schwieger, F., & Tebbe, C. (1998). A new approach to utilize PCR single-strand-conformation-polymorphism for 16S rRNA genebased microbial community analysis. Applied and Environmental Microbiology, 64, 4870 4876. Siciliano, S. D., Germida, J. J., Banks, K., & Greer, C. W. (2003). Changes in microbial community composition and function during a polyaromatic hydrocarbon phytoremediation field trial. Applied and Environmental Microbiology, 69(1), 483 489. Simon, C., & Daniel, R. (2011). Metagenomic analyses: past and future trends. Applied and Environmental Microbiology, 77, 1153 1161.

Metagenomics in remediation of contaminated sites 231 Simon, L., Levesque, R. C., & Lalonde, M. (1993). Identification of endomycorrhizal fungi colonizing roots by fluorescent singlestrand conformation polymorphism-polymerase chain reaction. Applied and Environmental Microbiology, 59, 4211 4215. Simpanen, S., Dahl, M., Gerlach, M., Mikkonen, A., Malk, V., Mikola, J., & Romantschuk, M. (2016). Biostimulation proved to be the most efficient method in the comparison of in situ soil remediation treatments after a simulated oil spill accident. Environmental Science and Pollution Research International, 23(24), 25024 25038. Smith, C. A., O’Reilly, K. T., & Hyman, M. R. (2003). Cometabolism of methyl tertiary butyl ether and gaseous n-alkanes by Pseudomonas mendocina KR-1 grown on C5 to C8 n-alkanes. Applied and Environmental Microbiology, 69(12), 7385 7394. Stefani, F. O. P., Bell, T. H., Marchand, C., de la Providencia, I. E., El Yassimi, A., St-Arnaud, M., & Hijri, M. (2015). Culture-dependent and -independent methods capture different microbial community fractions in hydrocarbon-contaminated soils. PLoS ONE, 10(6), e0128272. Su, C., Lei, L., Duan, Y., Zhang, K. Q., & Yang, J. (2012). Culture-independent methods for studying environmental microorganisms: Methods, application, and perspective. Applied Microbiology and Biotechnology, 93(3), 993 1003. Su, X. M., Bamba, A. M., Zhang, S., Zhang, Y. G., Hashmi, M. Z., Lin, H. J., & Ding, L. X. (2018). Revealing potential functions of VBNC bacteria in polycyclic aromatic hydrocarbons. Letters in Applied Microbiology, 66, 277 283. Tabak, H. H., Lens, P., Hullebusch, E. D. V., & Dejonghe, W. (2005). Developments in bioremediation of soil and sediments polluted with metals and radionuclides-1. Microbiolal processes and mechanisms affecting bioremediation of metal contamination and influencing meal toxicity. Reviews in Environmental Science and Biotechnology, 4, 115 156. Tarca, A. L., Roberto Romero, M. D., & Draghici, S. (2006). Analysis of microarray experiments of gene expression profiling. American Journal of Obstetrics and Gynecology, 195(2), 373 388. Taroncher-Oldenburg, G., Griner, E. M., Francis, C. A., & Ward, B. B. (2003). Oligonucleotide microarray for the study of functional gene diversity in the nitrogen cycle in the environment. Applied and Environmental Microbiology, 69, 1159 1171. Terro´n-Gonza´lez, L., Martı´n-Cabello, G., Ferrer, M., & Santero, E. (2016). Functional metagenomics of a biostimulated petroleum-contaminated soil reveals an extraordinary diversity of extradiol dioxygenases. Applied and Environmental Microbiology, 82, 2467 2478. Tiedje, J. M., Asuming-Brempong, S., Nusslein, K., Marsh, T. L., & Flynn, S. J. (1999). Opening the black box of soil microbial diversity. Applied Soil Ecology, 13, 109 122. Tiquia, S. M. (2010). Using terminal restriction fragment length polymorphism (T-RFLP) analysis to assess microbial community structure in compost systems. Methods in Molecular Biology, 2010(599), 89 102. Tveit, A. T., Urich, T., & Svenninga, M. M. (2014). Metatranscriptomic analysis of Arctic peat soil microbiota. Applied and Environmental Microbiology, 80(18), 5761 5772. Tyagi, M., da Fonseca, M. M. R., & de Carvalho, C. C. C. R. (2011). Bioaugmentation and biostimulation strategies to improve the effectiveness of bioremediation processes. Biodegradation, 22, 231 241. Uhlik, O., Leewis, M., Strejcek, M., Musilova, L., Mackova, M., Leigh, M. B., & Macek, T. (2013). Stable isotope probing in the metagenomics era: A bridge towards improved bioremediation. Biotechnology Advances, 31(2), 154 165. United States Environmental Protection Agency (USEPA). (1999). Use of monitored natural attenuation at superfund, RCRA corrective action, and underground storage tank sites. U.S. EPA Directive Number 9200.4-17P. United States Environmental Protection Agency (USEPA). (2012). A citizen guide to bioremediation. EPA 542-F-12-003. Varshney, A., Mohapatra, T., & Sharma, R.P. (2004). Molecular mapping and marker assisted selection of traits for crop improvement. In: Srivastava, A., Narula, S. (Eds.), Plant biotechnology and molecular markers (p. 289). Anamaya, New Delhi, India.

232 Chapter 8 Wang, Y., & Qian, P. Y. (2009). Conservative fragments in bacterial 16S rRNA genes and primer design for 16S ribosomal DNA amplicons in metagenomic studies. PLoS One, 4(10), e7401. Wellington, E. M. H., Berry, A., & Krsek, M. (2003). Resolving functional diversity in relation to microbial community structure in soil: exploiting genomics and stable isotope probing. Current Opinion in Microbiology, 6, 295 301. Whitman, W. B., Coleman, D. C., & Wiebe, W. J. (1998). Prokaryotes: The unseen majority. Proceedings of the National Academy of Sciences of the United States of America, 95(12), 6578 6583. Wikstro¨m, P., Ann-Christin., & Forsman, F. M. (1999). Biomonitoring complex microbial communities using random amplified polymorphic DNA and principal component analysis. FEMS Microbiology Ecology, 28(2), 131 139. Wilhartitz, I., Mach, R. L., Teira, E., Reinthaler, T., Herndl, G. J., & Farnleitner, A. H. (2007). Prokaryotic community analysis with CARD-FISH in comparison with FISH in ultra-oligotrophic ground- and drinking water. Journal of Applied Microbiology, 103, 871 881. Woese, C. R. (1987). Bacterial evolution. Microbiological Reviews, 51(2), 221 271. Wood, J. L., Zhang, C., Mathews, E. R., Tang, C., & Franks, A. E. (2016). Microbial community dynamics in the rhizosphere of a cadmium hyper-accumulator. Scientific Reports, 6, 36067. Wu, L., Thompson, D. K., Li, G., Hurt, R. A., Tiedje, J. M., & Zhou, J. (2001). Development and evaluation of functional gene arrays for detection of selected genes in the environment. Applied and Environmental Microbiology, 67, 5780 5790. Wu, Z., Lin, W., Li, B., Wu, L., Fang, C., & Zhang, Z. (2015). Terminal restriction fragment length polymorphism analysis of soil bacterial communities under different vegetation types in subtropical area. PLoS ONE, 10(6), e0129397. Yan, X., Luo, X., & Zhao, M. (2016). Metagenomic analysis of microbial community in uranium-contaminated soil. Applied Microbiology and Biotechnology, 100, 299 310. Yergeau, E., Sanschagrin, S., Beaumier, D., & Greer, C. W. (2012). Metagenomic analysis of the bioremediation of diesel-contaminated Canadian high Arctic soils. PLoS ONE, 7(1), e30058. Yin, H., Niu, J., Ren, Y., Cong, J., Zhang, X., Fan, F., . . . Liu, X. (2015). An integrated insight into the response of sedimentary microbial communities to heavy metal contamination. Scientific Reports, 5, 14266. Zhang, C., & Bennett, G. N. (2005). Biodegradation of xenobiotics by anaerobic bacteria. Applied Microbiology and Biotechnology, 67(5), 600 618. Zhang, D., Luo, J., Lee, Z. M. P., et al. (2016). Characterization of bacterial communities in wetland mesocosmsreceiving pharmaceutical-enriched wastewater. Ecological Engineering, 90, 215 224. Zumstein, E., Moletta, R., & Godon, J.-J. (2000). Examination of two years of community dynamics in an anaerobic bioreactor using fluorescence polymerase chain reaction (PCR) single-strand conformation polymorphism analysis. Environmental Microbiology, 2, 69 78.

Further reading Blackwood, C. B., Marsh, T., Kim, S., & Paul, E. A. (2003). Terminal restriction fragment length polymorphism data analysis for quantitative comparison of microbial communities. Applied and Environmental Microbiology, 69(2), 926 932. Botero, L. M., D’imperio, S., Burr, M., McDermott, T. R., Young, M., & Hassett, D. J. (2005). Poly (a) polymerase modification and reverse transcriptase PCR amplification of environmental RNA. Applied and Environmental Microbiology, 71(3), 1267 1275. Hanreich, A., Schimpf, U., Zakrzewski, M., Schlu¨ter, A., Benndorf, D., Heyer, R., et al. (2013). Metagenome and metaproteome analyses of microbial communities in mesophilic biogas-producing anaerobic batch fermentations indicate concerted plant carbohydrate degradation. Systematic and Applied Microbiology, 36, 330 338. Pace, N. R., Stahl, D. A., Lane, D. J., & Olsen, G. J. (1985). Analyzing natural microbial populations by rRNA sequences. ASM News, 51, 4 12. Zhou, J. (2003). Microarrays for bacterial detection and microbial community analysis. Current Opinion in Microbiology, 6, 288 294.

CHAPTER 9

In situ bioremediation techniques for the removal of emerging contaminants and heavy metals using hybrid microbial electrochemical technologies M.M. Ghangrekar, S.M. Sathe and I. Chakraborty Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, India

9.1 Introduction Rapid urbanization and development have resulted in significant increase in not only the quantity of wastewater generated, but also the complexity of the wastewater. Increase in the consumption of anthropogenic ingredients, namely emerging contaminants (EC) or xenobiotics (dyes, herbicides, hydrocarbons, personal care products, pesticides, etc.) have resulted in the discharge of these ECs into water bodies, groundwater, or soil. The wellestablished primary and secondary biological treatment processes often fail to remove these ECs leading to their uncontrolled discharge into the environment (Behera, Kim, Oh, & Park, 2011; Radjenovic, Petrovic, & Barcelo´, 2007). The presence of xenobiotic compounds in the effluent has demonstrated adverse impacts on the existing aquatic organisms (Godheja, Shekhar, Siddiqui, & Modi, 2016). The continuous discharge of xenobiotic compounds—laden wastewater into a specific location leads to the accumulation of these pollutants. The resulting accumulations of various xenobiotic compounds in the soil and water are of significant concern because of their high toxicity, carcinogenicity, and also because of their ability to bioaccumulate in living organisms (Singh & Ward, 2004). Similar to microorganisms, some plants have the capacity to take up and accumulate metals, pesticides, solvents, crude oil, and many industrial contaminants (Tak, Ahmad, & Babalola, 2013).

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00009-2 © 2020 Elsevier Inc. All rights reserved.

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234 Chapter 9

9.1.1 Bioremediation for pollution control and classification of bioremediation techniques In view of the increase in the number of contaminated sites all over the world a tangible solution is required to minimize the spread of pollutants and at the same time reduce the contaminant concentrations. One promising treatment method is to exploit the ability of microorganisms to remove pollutants from contaminated sites. Bioremediation using microorganisms is an alternative treatment strategy that is effective, minimally hazardous, economical, versatile, and environment-benign. Bioremediation means the use of living organisms, primarily microorganisms, to degrade the environmental contaminants into less toxic forms. Bioremediation comprises the enhanced degradation of toxic compounds by their transformation into less toxic substances for the clearance or immobilization of the contaminants. The process involves utilization of naturally occurring microorganisms which degrade and utilize the toxic compounds for their growth and as a source of energy. Because of this reason, bioremediation can only be effective where environmental conditions permit microbial activity (Brar et al., 2006; Dixit et al., 2015; Vidali, 2001). Many of the xenobiotic compounds can be effectively removed in various conventional processes like adsorption, advanced oxidation, and membrane filtration (Roccaro, Sgroi, & Vagliasindi, 2013). However, these processes are observed to be energy and cost-exhaustive and may have negative environmental impacts when employed for in situ remediation. On the other hand, bioremediation is an eco-friendly low-cost method for the restoration of contaminated sites. Currently, in situ bioremediation techniques in the form of biosparging, bioventing, and bioaugmentation are commonly used for the bioremediation of contaminated sites (Panda & Dhal, 2016). Based on the application of bioremediation, it is classified as either ex situ bioremediation or in situ bioremediation. Ex situ bioremediation techniques are applied for the treatment of groundwaters or soils which have been removed from the contaminated site and treated at some other location using biological and/or physicochemical methods. In situ bioremediation techniques are those which are applicable to the contaminated water or soil at the contaminated site itself with minimal disturbance in terms of excavation. Ex situ bioremediation involves the cleanup of contaminants after their transportation to a suitable site by means of pumping or excavation. However, the transportation of contaminated media from the original place to the treatment place makes such systems highly cost-intensive (Azubuike, Chikere, & Okpokwasili, 2016). Ex situ remediation of contaminated groundwater and surface water, by extracting the contaminated water from the site and remediation of the same in a distant treatment facility, is known as a “pump and treat system.” In addition to the high cost associated with the “pump and treat system,” complete flushing of the pollutant may require a large volume of water to be treated continuously over a long duration of time. Consequently, the remediation by this method is

In situ bioremediation techniques 235 likely to take several decades for the cleaning up of the site (National Research Council, 1993). On the other hand, in situ bioremediation involves on-site treatment using aquatic plants, aquatic animals, and microbes with minimal disturbance to the existing site, which makes the system highly cost-effective (Ateia, Yoshimura, & Nasr, 2016). In situ remediation has been successfully used for the remediation of dyes (Folch, Vilaplana, Amado, Vicent, & Caminal, 2013), chlorinated solvents (Frascari, Zanaroli, & Danko, 2015), heavy metals (Rezania, Taib, Md Din, Dahalan, & Kamyab, 2016), and polyaromatic hydrocarbon (PAH) (Sun et al., 2011) from contaminated media.

9.1.2 Microbial electrochemical technology Microbial electrochemical technology (MET) is the use of an electrochemical process using microbes as the catalyst. The MET systems are based on the interactions between living microorganisms and electrodes. These MET systems can be broadly classified into microbial fuel cells (MFCs), microbial electrolysis cells (MEC), and microbial desalination cells (Mohan et al., 2019). A typical MFC consists of an anode chamber and a cathode chamber separated by a proton exchange membrane (PEM) in which the anodic chamber is inoculated with anaerobic sludge. During the anaerobic respiration by microorganisms, protons and electrons are released; the electrons flow from anode to cathode via an external circuit, while the protons migrate through the PEM (Jadhav, Ghosh Ray, & Ghangrekar, 2017). In the case of MFC, electrical energy is harvested by externally connecting the load, while in MEC external power is supplied. Based on the applicability of bioelectrochemical systems in other setups, they can be classified as constructed wetland-microbial fuel cell (CW-MFC), plant-MFC, soil-MFC, and sediment-MFC.

9.2 In situ bioremediation using microbial electrochemical technologies A few in situ bioremediation techniques coupled with MET, namely CW-MFC, soil-MFC, sediment-MFC, and plant-MFC, are being discussed. The degradation of different xenobiotic compounds or ECs like dyes, pesticides, herbicides, PAHs, and heavy metals have been discussed. These hybrid microbial electrochemical technologies (HMETs) are a conglomeration of natural systems with the METs, in which the advantages of both the systems can be exploited.

9.2.1 Constructed wetlands-microbial fuel cells CW consist of the application of vegetation for in situ treatment of soil and water or wastewater. This time-tested bioremediation technique has been used for the removal of organic pollutants and nutrients. Additionally, its applicability in the removal of heavy metals (cadmium, chromium, zinc, lead, copper, nickel, and iron) has also been explored in

236 Chapter 9 the recent past (Gill, Ring, Casey, Higgins, & Johnston, 2017; Khan, Ahmad, Shah, Rehman, & Khaliq, 2009; Walker & Hurl, 2002). In addition to this, ECs namely surfactants, antibiotics, dyes, and pesticides have also been removed in CW (Hussein & ˇ ´ma, Scholz, 2018; Ramprasad & Philip, 2016; Santos, Almeida, Ribeiro, & Mucha, 2019; Sı Havelka, & Holcova´, 2009; Vymazal & Bˇrezinova´, 2015). The METs in the form of MFC and MEC are cutting-edge new technologies, which can act as bioremediation units for contaminated media (soil, sediment, water, etc.) while simultaneously producing electricity. Practically, CW and MFC are two distinct systems capable of individually degrading the pollutants. By combining CW and MFC to form a single composite hybrid system, the advantage of both systems can be explored for the degradation of organic pollutants as well as ECs. In the work done by Yadav, Dash, Mohanty, Abbassi, and Mishra (2012), it was emphasized that in several natural processes (flooded systems, natural wetlands), redox conditions are prevailing. The prevalence of such redox conditions is favorable for the development of in situ MFC, which may directly be applied at the field scale. The CW technology involves the implementation of a controlled treatment basin, which contains wastewater, media, microorganisms, and vegetation. Typically in a CW, based on the root depth of vegetation and water depth, two distinct zones—aerobic and anaerobic—can be observed. The area in the vicinity of the roots and the exposed top surface are likely to be aerobic, while with an increase in water depth, that is, away from the root zone, anaerobic conditions are witnessed. A typical MFC consists of two distinct zones, aerobic (cathodic chamber) and an anaerobic (anodic chamber) in which the reduction of oxygen and the oxidation of organic matter take place, respectively. This functional similarity of two distinct zones (anaerobic and aerobic zones) in both systems resulted in combining both the systems in a single hybrid setup, which led to simultaneous pollutant degradation and energy recovery (Yadav et al., 2012). In a hybrid CW-MFC system, the cathode is embedded near the roots to facilitate aerobic conditions, while the anode is placed at a greater depth which is anaerobic, making it ideal for an anaerobic reduction. A novel CW coupled with MFC, which takes advantage of MFC as well as phytoremediation, was reported by Yadav et al. (2012) for the first time for the degradation of dye-laden synthetic wastewater. The system achieved 93.15% and 75% of dye and chemical oxygen demand (COD) removal, respectively, at 96 h hydraulic retention time (HRT) while producing a maximum power density (MPD) and current density of 15.73 and 69.75 mA/m2, respectively (Yadav et al., 2012). Subsequent to the initial hybridization of MFC in CW, the CW-MFC has been successfully applied for the removal of organic contaminants from wastewater, which demonstrated enhanced performance compared to stand-alone CW. The CW coupled with electrochemical system exhibited a higher degradation of organic matter, dye, nitrogen, and phosphorus than conventional CW (Fang, Song, Cang, & Li, 2013; Ju, Wu, Huang, Zhang, & Dong, 2014; Srivastava, Yadav, & Mishra, 2015).

In situ bioremediation techniques 237 In a few of the recent studies, CW-MFC in an upflow configuration was designed in which wastewater initially comes into contact with the anode and later with the cathode as the wastewater moves in the upward direction (Liu, Song, Li, & Yang, 2013; Oon et al., 2016). Stand-alone CW has been used for the removal of xenobiotics like dyes, pesticides, phenolic compounds, hydrocarbons, and pharmaceuticals (Dordio & Carvalho, 2013). Similar to CW, stand-alone MFC has also been used for the treatment of xenobiotics in the past (Das, Das, Chakraborty, & Ghangrekar, 2019; Rathour, Kalola, et al., 2019). A hybrid system of CW-MFC is useful for in situ bioremediation owing to a high removal rate of pollutants because the anode acts as an additional electron acceptor. In situ contamination of the polluted sites can be accomplished by using CWMFC for the removal of organic compounds and nutrients, along with heavy metals, dyes, pesticides, and PAH, etc. Degradation of Brilliant Red X-3B up to 92.8% was achieved using CW-MFC at HRT of 3 days (Fang, Song, Cang, & Li, 2015). Experiments were performed by varying the dye proportion from 10% to 90% while keeping COD constant (180 mg/L). A decolorization rate of 97.3% was reported to be highest when the ratio of dye to COD was 0.1. Similar to decolorization, electricity production also dropped with an increase in dye concentration. The attenuation of removal efficiency and electricity production was attributed to the inhibitory effect of increased dye concentration. In another work, Fang, Cheng, Wang, Cao, and Li (2017) analyzed the effect of using different supporting media in CW on the power density and decolorization of Brilliant Red X-3B dye with glucose as the cosubstrate. Gravel, screes, glass beads, and biological ceramics were used as media; gravel resulted in 92.7% decolorization and 0.117 W/m3 MPD, while the use of glass beads resulted in 76.3% decolorization and 0.256 W/m3 MPD. It indicated that media supporting the lowest biomass growth (glass beads) had better electricity production than other media. The Brilliant Red X-3B decolorization efficiency followed an inverse trend in comparison to power generated. For Brilliant Red X-3B dye degradation, oxidation reduction potential has to be lower than 50 mV, and in the middle and cathodic layer the oxidation reduction potential was higher than 50 mV, hence degradation of dye was not observed in these two layers. Decolorization of Acid Red 18 in upflow CW-MFC was explored by Oon et al. (2018), which resulted in 91% decolorization with the initial concentration of 500 mg/L. Despite removal efficiency being higher than 91% for dye concentration varying from 100 to 500 mg/L, the power density reduced with an increase in dye concentration. The Acid Red 18 degradation was due to biological and electrochemical activity and it followed the second-order kinetics model. The MPD without dye was 3.97 mW/m2, which dropped to 0.02 mW/m2 when the dye concentration was increased to 500 mg/L. This declining trend in MPD showed the interference of dye on electrogen activity, which adversely affected the electricity production.

238 Chapter 9 The synthetic samples used in research experiments contain only specific dyes, which are comparatively easier to break down than the complex nature of dyes in real wastewater. Hence with an aim to assess the practicability of a CW-MFC system in treating real-time dye-laden wastewater, Rathour, Patel, Shaikh, and Desai (2019) used real-time dye wastewater. The experiments resulted in 82.2% decolorization, which was 9% higher than stand-alone CW. A rise in dye concentration from 10% to 100% (v/v) in the wastewater resulted in a 16.63% reduction in decolorization, which was similar to that of Fang et al. (2015). The decolorization mechanism can be associated with increased metabolic rates by providing more electrons in CW-MFC as compared to stand-alone CW. Heavy metals (chromium, cadmium, arsenic, nickel) were also removed from CW-MFC; however, removal by stand-alone CW or MFC was not assessed in order to understand the precise mechanism for heavy metal removal. As mentioned in the above section, MFC and CW are capable of solely degrading ECs. However, a combined CW-MFC for the degradation of hydrocarbons was first attempted by Wei et al. (2015). Methyl-tert-butyl ether (MTBE) is the second most widely used oxygenate in gasoline, after ethanol. The accidental release of MTBE along with the gasoline from the underwater storage tanks and pipelines are the most common cause of MTBE contamination in groundwater (Deeb et al., 2003). Such hydrocarbons are easily absorbed in soil strata and subsoil water, hence bioremediation of the same using CW-MFC can be a feasible in situ solution. With the intention to bioremediate MTBE-contaminated groundwater sites, Wei et al. (2015) used the horizontal flow CW-MFC technique to treat benzene and MTBE contamination. The system consisted of four vertical anode modules stacked one above the other inside the CW media. Power generation performance of the top two anodes near the roots was significantly less than the bottom modules due to the availability of dissolved oxygen (DO) at the top. The system showed nearly complete removal of both pollutants prior to achieving the steady-state. The removal was significantly higher near the root zones than away from the roots indicating the significance of higher DO for the degradation of benzene and MTBE. The synergy of the aerobic bacteria and anaerobic electrogens in the root zone was the cause of such a high removal efficiency. In another work, Tu¨rker and Yakar (2017) demonstrated the use of CW-MFC for boron removal which resulted in a 63.4% removal efficiency at the initial concentration of 12.3 mg/L. However, an increase in boron concentration had an adverse impact on the bioelectricity production, which decreased to 39.2 from 75.7 mW/m2 with an increase in boron concentration from 5.5 to 12.3 mg/L. Boron removal was achieved due to sorption on supporting media, accumulation in organic media, and plant uptake. Plants such as Typha latifolia in CW were well adapted to the boron toxicity. A few of the soil enzymes present in CW media also assisted the bioelectricity production. Upflow CW-MFC for the removal of sulfadiazine operated by Song et al. (2018) revealed that 5.46 times improved pollutant removal was obtained in a closed circuit mode than in

In situ bioremediation techniques 239 an open circuit mode after 3 days of HRT. At a lower HRT of 1.5 days, MPD was reduced to 12.84 from 15.41 mW/m2 due to the higher accumulation of sulfadiazine on the electrodes. It demonstrated that the closed circuit and HRT had a significant impact on sulfadiazine removal. In a separate work, membraneless air cathode MFC coupled with CW and planted with water hyacinth was used to treat nitrobenzene (Xie et al., 2018). At a nitrobenzene:COD ratio of 1:16, MPD of 1.53 mW/m2 along with 92.89% nitrobenzene removal was obtained, which was higher than that for the 1:8 and 1:24 ratios. This indicated that cometabolism induced by microorganisms and glucose may have encouraged a higher nitrobenzene removal. Higher nitrobenzene concentrations (1:8 wt. ratio) exhibited an inhibitory effect on the microbes, while at lower nitobenzene concentration (1:24 wt. ratio) the metabolic pathway was entirely shifted toward the biodegradable cosubstrate. As compared to stand-alone MFC, the performance of CW-MFC was higher, indicating the ability of plants to adsorb and bioaccumulate nitrobenzene. The utilization of specific flora in CW-MFC can be targeted to achieve a specific type of decontamination. In an experiment by Zhang et al. (2017), the removal efficiency of tetracycline and sulfamethoxazole in CW-MFC using Oenanthe javanica as wetland vegetation in the cathodic layer was examined. At the initial concentration of 1600 µg/L for both antibiotics, the effluent concentration of sulfamethoxazole and tetracycline was reported as 0 1.65 and 0 2.40 µg/L, respectively. The removal of antibiotics can be due to absorption, microbial degradation, plant uptake, hydrolysis, and electrochemical reactions. The specific reasons were not ascertained in the work, hence further studies can be conducted focusing on the exact pathway of antibiotic degradation in CW-MFC. In another investigation conducted by a different group of researchers, a higher concentration of sulfamethoxazole (2 4 mg/L) was removed by CW-MFC with an efficiency of about 99% (Li, Song, et al., 2018). An anode modified by nano zero-valent iron (nZVI) was used for the removal of phenanthrene and anthracene in CW-MFC planted with Typha orientalis (Wang et al., 2019). At an initial concentration of 0.17 mg/L the removal of PAH was in the range of 88.5% to 96.4%, simultaneously producing 10.9 and 20.6 mW/m2 of MPD, respectively. Nickel and carbon fiber felt-modified nZVI facilitated phytoextraction and biodegradation as the nZVI acted as an electron donor. The growth of Bacillus sp. was detected on the anode, which produced the catalase enzyme responsible for the degradation of polycyclic aromatic hydrocarbons. The presence of Paludibacter, Desulfovibrio, and Lactococcus, which are capable of biodegradation, were also detected in CW-MFCs. An upflow CWMFC was used for the removal of pharmaceutical and personal care products by Li et al. (2019). The results showed 9.3% and 18% higher removal efficiency in a closed circuit as compared to the open circuit for the removal of ibuprofen and bisphenol A (BPA), respectively. The electricity generation performance, including voltage and MPD, declined after an initial surge in voltage (28.63 24.94 mW/m2) with a change in HRT from 16 to 4 h.

240 Chapter 9 The formation of heterotrophic biofilm on the cathode might have led to the mass transfer resistance for the electrons and the protons, which might be the reason for reduced electricity generation at lesser HRT. At HRT of 16 h, ibuprofen and BPA removal was reported to be 95.7% and 91.3%, respectively. Anodic degradation of pollutants was higher than the cathodic degradation. Currently, considerable literature is available on the implementation of the hybrid system of CW-MFC for the degradation of xenobiotics (Table 9.1). The different strategies applied to the CW-MFC system, such as upflow mode CW-MFC, open circuit and close circuit operations, and application of modified anode, enunciates the fact that considerable research is yet to be done in this field. Additionally, there is plenty of scope to widen this field of Table 9.1: Performance evaluation of different CW-MFC hybrid systems for xenobiotics removal. Arrangement

Contaminant (initial concentration)

Removal (%)

Maximum power density Reference

Emerging contaminant Horizontal flow CW-MFC

Wei et al. (2015) Zhang et al. (2017) Tu ¨rker and Yakar (2017) Xie et al. (2018)

Benzene and methyl-tertGreater butyl ether (12 and 3 mg/L) than 98% Tetracycline and sulfamethoxazole (1.6 mg/L) Boron (12 mg/L) 63.4

39.2 mW/m2

Nitrobenzene

92.89

1.53 mW/m2

Sulfamethoxazole (2 4 mg/L) Ibuprofen and BPA (10 mg/L) Phenanthrene and anthracene (0.17 mg/L)

99

15.41 mW/m2 Song et al. (2018) Li, Song, et al. (2018) 28.63 mW/m2 Li et al. (2019)

CW-MFC

Brilliant Red X-3B

92.7

0.117 W/m3

Upflow CW-MFC

Acid Red 18 (500 mg/L)

91

0.02 mW/m2

CW-MFC

Real-time dye-laden wastewater Brilliant Red X-3B

82.2

198.8 mW/m2

92.8

0.0619 W/m3

Hybrid CW-MFC Hybrid CW-MFC Single chambered air cathode membraneless CW-MFC Upflow MFC coupled CW Biofilm electrode reactor followed by CW-MFC Upflow CW-MFC nZVI modified CW-MFC

Sulfadiazine

95.7 and 91.3 88.5 96.4

0.0578 W/m2

Wang et al. (2019)

Dye removal

CW-MFC

CW-MFC, Constructed wetlands-microbial fuel cells; nZVI, nano zero-valent iron; BPA, bisphenol A.

Fang et al. (2017) Oon et al. (2018) Rathour, Patel, et al. (2019) Fang et al. (2015)

In situ bioremediation techniques 241 research focusing bioremediation on fields contaminated with heavy metals. CW as well as MFC as an individual system is capable of removing some of the heavy metals. However, the impact of a hybrid CW-MFC system on the removal of heavy metals is yet to be analyzed. CW proved to be a useful technique for acid mine drainage treatment (RoyChowdhury, Sarkar, & Datta, 2015). For the same, CW-MFC hybrid can be implemented and its effect on removal can be assessed. Both the systems, MFC as well as CW, share a symbiotic relationship and there is immense potential to collectively exploit the advantages of each system to speed up the pollutant degradation rate. The hybrid system’s applicability for in situ conditions and its robustness makes it one of the best choices for in situ bioremediation of not only conventional pollutants but also of xenobiotic compounds.

9.2.2 Sediment-microbial fuel cells In cognizance of the fact that river sediments have organic matter content and that the biodegradation of such organic matter is slow in process, it is prudent to describe the application of MET for the remediation of the same. Electrochemically active bacteria accelerate the metabolism of such organic matter as the availability of terminal electron acceptors is achieved by an electrode inserted in the sediment. This is unlike conventional sediment biomass, in which there is a lack of a terminal electron acceptor as all processes in the sediment are rate limited by the slow diffusive movement of different ions and absorbed compounds. Sediment-MFC provides a low-cost solution for in situ bioremediation of contaminated sediments (Neethu & Ghangrekar, 2017; Sajana, 2016). In a sediment-MFC, the anode is inserted in sediment, while the cathode is placed in the overlying water (Fig. 9.1). This arrangement of the anode acts as an improved electron

Figure 9.1 Schematic of sediment microbial fuel cells.

242 Chapter 9 acceptor in comparison to in situ terminal electron acceptors, such as nitrate, sulfate, and iron. This phenomenon was verified by Yan, Song, Cai, Tay, and Jiang (2012) in a sediment-MFC work in which a comparison of ferric oxide and an anode as an electron acceptor was made. In this work, ferric oxide and an electrode (working as the anode) were introduced individually in a hydrocarbon (phenanthrene and pyrene)-contaminated sediment layer. It was observed that on an individual basis, the anode had a greater performance than the ferric oxide in terms of PAH degradation. The combined effect of ferric oxide and anode both acting as a terminal electron acceptor for the sediment microbes had a greater efficiency than the stand-alone techniques for phenanthrene (99.74%) and pyrene (94.8%) degradation. Placement of the cathode in a sediment-MFC setup is crucial as the DO levels are controlled by a passive gas exchange between the water and atmosphere and the O2 release by the aquatic flora. In order to expedite the oxygen reduction reaction kinetics, the oxygen availability has to be increased to reduce the mass transfer overpotential. In the experiment conducted by Morris and Jin (2012), the cathode was constructed from carbon cloth and it was placed in a partially submerged position to facilitate wicking action. The setup was compared to a fully submerged cathode and in both the sediment-MFCs the effective water depth was maintained at 1 m to stimulate a shallow lagoon or pond condition with 0.2 m sediment bed. The sediment-MFC setups with wicking cathode and submerged cathode were compared to the control sediment layer devoid of any electrode to explore the possibility of enhanced bioremediation of petroleum hydrocarbon. It was observed that the sediment-MFC could achieve higher (12 times) degradation as compared to the control sediment layer, wherein petroleum hydrocarbon removal efficiency was found to be 24.4% and 2.1%, respectively (for an initial concentration of 15,958 mg/kg). In terms of electricity generation, it was observed that the wicking cathode had a higher voltage output (124 mV) than the submerged cathode (34 mV). In situ remediation of different recalcitrant pollutants should also ensure that the intermediates formed can be further degraded in the same system so as to render a complete in situ solution. In an investigation considering the intermediates of PAH degradation, such as toluene and benzene, it was revealed that the aromatic oxidation was faster by Geobacter metallireducens when the contaminated sediment layer was modified with graphite felt as the primary electron acceptor (Zhang, Gannon, Nevin, Franks, & Lovley, 2010). The degradation of aromatics, such as polychlorinated biphenyls (PCBs), absorbed in freshwater sediment collected from a contaminated site was investigated using sediment-MEC (Chun, Payne, Sowers, & May, 2013). The investigation revealed that a major fraction (90%) of the absorbed PCBs found in the collected sediment were di-, tri-, and tetrachlorinated compounds in a combined concentration of 20.2 6 4.0 mg/kg of dry sediment. In the same investigation, the removal efficiency of sediment spiked with PCBs (B100 ppm) was also investigated. The application of voltage in sediment-MEC in a range of 1 3.0 V led to

In situ bioremediation techniques 243 enhanced degradation of PCBs. Application of voltage resulted in O2 generation on the anode beyond 1.5 V. The degradation of the PCBs mainly followed aerobic oxidation in the microaerobic pockets formed near the anode due to this O2 formation and the reductive dechlorination in anaerobic regimes (Chun et al., 2013). Apart from refractory compounds, heavy metal contamination can also be remediated using sediment-MFC. Unlike refractory compounds, heavy metals can only be contained or immobilized so as to avoid spreading of the same in the contaminated media. Such immobilization can be brought about by phase change due to changes in redox condition, pH of the media, or by biotransformation of the metallic ion. Introduction of a potential gradient, through a bioelectrochemical reaction at the anode in the case of a sediment-MFC or through an external power in the case of electrolytic half cells, can help in the retention of such a precipitated phase of the metal. In an examination conducted by Gregory and Lovley (2005), a sediment electrolysis cell (sediment-EC) was operated at a poised potential of 500 mV (vs an Ag/AgCl reference electrode) for conversion of U(VI) to U(IV), which was deposited as a precipitate in the electrode zone. However, it was observed that the precipitated uranium went back to the solution once the poised potential was removed. In comparison to the sediment-EC, sediment-MEC operated with Geobacter sulfurreducens did not show any dissolution of U(IV) even after 600 h. The immobilization occurred due to the reduction of U(VI) to U(IV) and because the anode in sediment-MFC had a continuous supply of electrons from the bioelectrochemical oxidation of organics, the uranium remained in the stationary phase. This discussion shows the advantage of the sedimentMFC for capacitive deionization over the conventional sediment-EC. The bioremediation of metal-contaminated sediment is favorable at a near-neutral pH regime. The removal of cadmium, arsenic, and lead was observed to be 77.7%, 90.86%, and 83.91% in near-neutral pH, respectively (Abbas, Rafatullah, Ismail, & Nastro, 2017). Sediment-MFC provides an excellent option of enhanced bioremediation simultaneously with electricity generation. The voltage generated from sediment-MFC can be mapped to the heavy metal concentration, which can serve as a biosensor for detecting heavy metal levels. Anodic biofilm accelerates the biodegradation of refractory compounds compared to the plain sediment as the anode serves a perpetual sink for electrons. The past investigations throw light on various configurations and operating conditions of sediment-MFC, which can be employed for decontamination of marine and freshwater sediments. The different configurations discussed can be modified according to the treatment requirement.

9.2.3 Soil-microbial fuel cells Soil-based MFC follows the basic principles of conventional MFC, where soil acts as an anaerobic nutrient media containing pollutants. The cathode can be placed at the bottom

244 Chapter 9 with an air cathode arrangement, or on the top of the contaminated soil while the anode is embedded in the contaminated soil at a certain depth with anaerobic conditions prevailing around it. For the in situ bioremediation, however, the cathode is able to be placed at the top while the anode is embedded inside the contaminated soil. The anaerobic degradation of various complex pollutants is possible by anaerobic bacteria; however, in the water-logged and marshy soils, due to the unavailability of a suitable terminal electron acceptor (such as nitrate, sulfate, etc.), the pollutants persist for an extended period (Huang et al., 2011). For most of the contaminant removal in a shorter time, oxygen is a preferred oxidant. Since the addition of oxygen is highly difficult due to low solubility and higher power requirements, it is not a feasible method for the in situ bioremediation of contaminated soils. As mentioned above, MFC has a pronounced potential to enhance the productivity of standalone anaerobic digestion in terms of pollutant removal. The use of MFC in the contaminated soils compliments the process of anaerobic degradation in an innovative approach. This is achieved by providing additional electron acceptors in the form of an electrode, thus speeding up the remediation process. On a similar line, the hybrid system of soil-MFC is not only a low-cost technique but also has the potential to degrade the ECs like PAH, herbicides, and phenolic compounds as detailed in the succeeding section (Fig. 9.2). The early application of soil-MFC for degradation of EC was reported by Huang et al. (2011) with the aim to remediate soil contaminated with phenol. The system consisted of insertion type MFC in soil for simultaneous phenol degradation and electricity production. With 80 mg/L of initial phenol concentration, 90.1% removal efficiency was achieved in soil-MFC with a degradation constant rate of 0.39/d, which was 23 times higher than natural conditions. In another investigation, Cao, Yu, Wang, Zhou, and Li (2017) used External resistance (Huang et al., 2011) Phenol rem.- 90.1% MPD-29.45 mW m–2

Ω

e–

(Wang et al., 2017) Atrazine rem.-91% Current generated0.3 mA (Li et al., 2018b) Herbicide metolachlor rem.-56% Maximum current density-49.6 mA m–2

e–

Air cathode

(Li et al., 2014) 16 priority PAHs rem.- 36% MPD-37 mW m–2

H+ e– Anode

(Cao et al., 2017) Hexachlorobenzene rem.- 71.14% MPD-70.8 mW m–2

(Yu et al., 2019) TPH rem.-31.4% Maximum power density-29.98 mW m–2

Contaminated soil

Figure 9.2 Schematic of soil microbial fuel cells.

In situ bioremediation techniques 245 soil-MFC, for the removal of hexachlorobenzene from the soil. With an initial concentration of 40 mg/kg, a removal efficiency of 71.14% was reported, which was reduced to 50.06% as the concentration of hexachlorobenzene was increased to 200 mg/kg. The addition of the pollutant inhibited the activity of electrogenic bacteria initially, which was restored after an initial acclimatization period. Implementation of MFC has resulted in more than twice the removal obtained in an open circuit arrangement, showing the impact of soil-MFC on the removal of hexachlorobenzene. This emphasized that the removal mechanism was primarily bioelectrochemical oxidation at the anode. Moreover, sodium dodecyl sulfate (SDS) addition (0.0086 mol/L) led to desorption of hexachlorobenzene from the soil particles, thus increasing the mobility of the aromatic compound making it more “bioavailable.” The SDS addition further improved the removal efficiency (88.63%). Greater depth between the cathode and anode results in higher potential losses. However, the less the space between the two electrodes, the less will be the effective depth treated by soil-MFC. Hence to overcome this practical difficulty, and to treat greater depths and volumes of contaminated soil, multiple anodes were used at certain intervals by Li et al. (2014). The stacked multilayered anode configuration was used for the removal of petroleum hydrocarbons by Li et al. (2014) in an air cathode soil-MFC arrangement. Three anodes at a depth of 1, 3, and 5 cm from the cathode with activated carbon as a catalyst were used in saline soil polluted with hydrocarbons. The degradation rates of total petroleum hydrocarbons (TPH), 16 priority PAHs, and total n-alkanes (C8 C40) were 18%, 36%, and 29%, respectively. Furthermore, bacterial analysis determined the presence of Alcanivorax sp. and Firmicutes bacterial species, which might have been associated with hydrocarbon degradation. Such multiple anodes acted as a booster in reducing the entire electrical pathway of charges. The anode material in MFC plays an important role in electricity generation as it is related to electron transfer and the successful growth of bacterial colony over the cathode (Zeng, Zhao, & Li, 2015). Hence to improve the power output of soil-MFC, an anode catalyst was used in soil-MFC by Yu, Li, and Feng (2019). The modified graphite felt anode with Fe3O4 and bentonite-Fe were used for the removal of hydrocarbons from the soil in a soil-MFC setup. Both modifications showed accelerated electron transfer and had catalytic impact on anode half reaction, which resulted in higher pollutant removal. The Fe3O4 and bentoniteFe-modified anode resulted in marginally higher removal of TPH by 1.5% and 4.9%, respectively, than bare graphite. In contrast, the MPD was increased by 2.82 and 1.72 times with bentonite-Fe and Fe3O4-modified anodes, respectively, as compared with bare graphite. This reveals that the TPH removal pathway was mainly bacterial metabolism and catalyst modifications only enhanced the power generation of the system. With the advances in agricultural technology, the use of herbicides is now common for controlling weeds. The excessive use of herbicides on soils can lead to the disturbance of

246 Chapter 9 the biochemical balance of the soil, which can reduce the fertility and productivity of the soil (Marin-Morales, de Campos Ventura-Camargo, & Hoshina, 2013). Many of the chemicals used in herbicides are persistent contaminants which can last a few decades, thus drastically affecting the soil. The concentration of the herbicides is very low while the extent and spread are over a wide area. This makes it very challenging for ex situ remediation, hence the only viable option is in situ bioremediation. The soil contaminated with herbicides was treated using a soil-MFC technique by Wang, Li, Cao, Long, and Li (2017). The experiments revealed that the degradation of atrazine followed first-order decay and was faster in the aerobic layer compared to the anaerobic layer in a soil-MFC. The atrazine concentration was 7.87 mg/kg at the end of the experiment (63 days), which was lower than the other arrangements with an anode at a depth greater than 4 cm. Additionally, the half-life of the pollutant reduced from 53 to 40 days post soil-MFC treatment. Atrazine removal depends on the current generation, as in a closed circuit mode the anode is a major sink for electrons during the microbial degradation of atrazine. Still, the increase in the initial concentration to 750 mg/kg had an inhibitory effect on the electricity performance, which decreased by 1.75 times. In another experiment, Li, Li, Sun, Zhao, and Li (2018) applied air cathode soil-MFC for the removal of the herbicide metolachlor, which showed 2.62 and 1.76 times higher removal than a nonelectrode setup when spiked with 10 and 20 mg/kg of pollutant, respectively. The herbicide metolachlor has high solubility and served as an electron donor, which was evident from the increased charge and power output at 20 mg/kg spikings than at 10 mg/kg. Moreover, the amount of pollutant degraded in soil-MFC at a higher initial concentration was higher than at a lower initial concentration. The highest degradation was observed close to the cathode showing that the cathodic bacterial consortia and DO played an important role in herbicide metolachlor degradation. The soil-MFC hybrid system provides a robust and cost-effective solution for the in situ bioremediation of xenobiotic compounds, while simultaneously harvesting the energy during the cleanup operation. Future research in the field should examine the scalability of the system on the actual site. Additionally, the effect of more than one type of xenobiotic and the removal potential should be assessed by soil-MFC system. The present limitations of soil-MFC system include the mass transfer limitation and applicability of the system to a greater depth of soil, which also needs to be researched in future. The upcoming investigations should also place an emphasis on the by-products of xenobiotic degradation and their toxicity.

9.2.4 Plant-microbial fuel cells The different variants of METs, as discussed in the Section 9.1.2 convert the chemical energy trapped in the wastewater organics to bioelectricity. In the case of plant-MFC, the

In situ bioremediation techniques 247 anodic biofilm derives its source of energy and carbon from the root exudates in the rhizosphere (Nitisoravut & Regmi, 2017). The root exudates are secreted by the plants to hinder the growth of competitive species, stimulate the growth of symbiotic species, and alter the physicochemical nature of the surrounding soil to their own benefit (Nardi et al., 2000). The plant-MFC system acts as an energy conversion device, converting the incoming solar radiation to electrical energy with minimal or zero waste generation. Typically, a plant-MFC system has two units, namely a biocontrol unit and a bioprocess unit. The plant or the biocontrol unit photosynthesizes its own food using solar energy, which can be treated as an input signal. The secreted root exudates are metabolized by the rhizosphere microbes acting as a bioprocess unit, which gives electrons as the output. This is an open loop system in which the solar radiation (input) is converted to electricity (output). This complex system of microorganisms, plant roots, soil as solid phase, and moisture in the root zone as a liquid phase has great potential for the amendment of the contaminated in situ media (Strik et al., 2011). The various pathways of in situ pollutant abatement include adsorption in bacterial biomass and plant roots, phase transformation, bacterial degradation, oxidation, etc. The detailed pathway of each type of pollutant degradation requires a detailed understanding of this microecosystem. For example, elemental pollutants, such as heavy metals, are mainly reduced to an insoluble form prior to adsorption in the soil biomass or soil itself. On the other hand, a few soluble metallic ions are removed through bioassimilation in the plant root tissues. The plant-MFC itself is a hybrid technology of plant bioremediation technique and MET; however, it can be considered as a foundation for the further design of different in situ bioremediation systems. For example, in work done by Arends et al. (2014) it was revealed that methane generation from rice cultivation can be reduced by shifting the electron transfer pathway toward exogenous electron transfer instead of the production of methane. This was a specific application of plant-MFC for the reduction of greenhouse gas emissions. Further to this work, Md Khudzari, Garie´py, Kurian, Tartakovsky, and Raghavan (2019) demonstrated that introducing a biochar anode in the plant-MFC as compared to a commercially available carbon felt anode reduces methane generation by 39%. However, the plant-MFC with a biochar anode designed by Md Khudzari et al. (2019) produced 16% less bioelectricity compared to the control plant-MFC, which was operated with carbon felt anodes. Further work would be required to overcome the hurdle of lower electricity generation of replacing the carbon felt anode with a biochar anode. Apart from greenhouse gas emission reduction, plant-MFCs have been extensively evaluated for the removal of toxic metals, pesticides, and aromatic hydrocarboncontaminated sites. Past reviews have discussed the different architectural and design aspects of plant-MFCs. The reviews have also discussed the effects of different plants used in the plant-MFC designs, the effects of light and dark cycles, and the effects of different

248 Chapter 9 Table 9.2: Performance evaluation of different plant-MFC hybrid systems for xenobiotics and heavy metal removal. Arrangement

Contaminant

Removal (%)

Maximum power density

Reference

Emerging contaminant Plant-MFC

54 48

Phenanthrene Pyrene (1000 mg/kg each) Oil degradation In presence of glucose In presence of surfactant Scarlet RR dye Textile industry effluent Pyrene Benzo[a]pyrene

52 83 89 87 87 76

Plant-MFC

Chromium

99

Plant-MFC

Chromium Anodic removal Cathodic removal Arsenic Lead Cadmium Chromium

35 16 100 63 56 46

Floating phytobed MFC Macrophyte based plant-MFC

Zhao et al. (2019)

0.077 W/m2

Kadam et al. (2018) Yan et al. (2015)

Heavy metals

Floating phytobed MFC

Habibul et al. (2016) Guan et al. (2019) 0.076 W/m2

Kadam et al. (2018)

MFC, Microbial fuel cells.

soil types on the different plant-MFC configurations. Concise literature is available on the working principle and different configurations of the plant-MFCs. It is evident from the discussion that plant-MFC has excellent scope for the mitigation of in situ pollution, which needs to be further exploited (Table 9.2). Hence this section is dedicated to the removal of pollutants such as EC and heavy metals using plant-MFC with reference to the various existing architectures and designs from past investigations. Removal of aromatics and hydrocarbons: In situ removal of PAH from contaminated soil using plant-MFC was investigated by Zhao, Deng, Hou, Li, and Yang (2019). The innovative work combined the addition of surfactant to the PAH-contaminated soil and further treatment in plant-MFC. The incorporation of a surfactant increased the solubility of the crude oil absorbed in soil, which was essential for increasing the bioavailability of the PAH to the rhizosphere microbes. Degradation of PAH was reported to be 40.8% in plantMFC. The effect of cosubstrate (glucose) addition in the contaminated soil led to higher degradation (52.1%) than plant-MFC devoid of any external carbon sources. The addition of cosubstrate at lower concentrations 500 mg/kg had the maximum effect on the PAH

In situ bioremediation techniques 249 degradation as compared to 1000 and 1500 mg/kg doses of glucose. The addition of glucose at a dose of 500 mg/kg was described as the most optimum concentration of glucose cosubstrate. The role of cosubstrate (glucose) shifted from aiding in cometabolism to becoming the dominant substrate to be utilized by the biota. Hence at higher glucose doses, the microbes shifted the metabolism toward cosubstrate digestion. Apart from in situ removal of ECs from the soil, floating variants of plant-MFC can also be utilized in the removal of contaminants in effluent directly. The application of such plantMFCs requires the selection of plants which can be integrated with the MFC in an aquatic environment. Two different plant species, namely Chrysopogon zizanioides and Typha angustifolia, were implemented in a phytobed MFC treating textile effluent by Kadam et al. (2018). Such a system facilitates the utilization of both the microbial consortium and the phytoremediation techniques for decontamination of the wastewater. In the investigation by Kadam et al. (2018) the decolorization of scarlet RR dye was carried out using the floating phytobed plant-MFC setup. The decolorization efficiency of the scarlet RR using individual plants C. zizanioides and T. angustifolia was reported as 85% and 81%, respectively. The consortium of the two plants achieved a higher decolorization efficiency of 89%. All the decolorization efficiencies reported were achieved in 60 h retention time. Such integrated systems are capable of handling real textile wastewater also, as reflected in the aforementioned investigation wherein the textile industry effluent was decolorized to an extent of 87% after 60 h of exposure. The investigation further elaborates the degradation mechanism of the contaminant dye in the form of bioabsorption followed by biotransformation in the plant root tissues. The identification of several enzymes in the plant tissues capable of dye degradation led to the conclusion that the plant releases several enzymes as a response to the stress induced due to the absorbed dye. The integrated floating phytobed and MFC system could achieve higher removal efficiencies as compared to a stand-alone floating phytobed reactor. The removal of hydrocarbons, namely pyrene and benzo[a]pyrene, by utilizing plant-MFC was achieved in a PAH contaminated site (Yan, Jiang, Cai, Zhou, & Krumholz, 2015). The research revealed that the synergistic effect of macrophyte Acorus calamus and MFC embedded in the soil strata was 70% more effective than soil-MFC or macrophyte alone for the remediation PAH-contaminated soil. It was emphasized that the PAH was removed through biological metabolism into the root zone. Microbes responsible for the degradation of PAH were identified in the biota of the rhizosphere (Yan et al., 2015). Removal of heavy metals: In accordance with the previous section about soil MFCs, in which in situ microbes regulate the entire bioremediation phenomena, in plant-MFCs the plant plays a major role. Heavy metals can be immobilized in the soil due to phase change (from liquid to solid) with a change in oxidation state or they can be bioassimilated in the plant root zone. In research conducted by Habibul et al. (2016) hexavalent chromium was

250 Chapter 9 reduced to Cr(III) by acting as an electron acceptor at the cathode of the plant-MFC system. The trivalent chromium underwent a phase change and deposited as Cr(OH)3 over the electrode, as revealed by X-ray photoelectron spectroscopy. Reduction and subsequent immobilization of chromium was achieved by the anodic oxidation acetate in the first phase of operation and by anodic oxidation of root exudates in the second phase of operation. It was observed that the during long-term operation for chromium removal, the addition of acetate as an external electron donor increased the Cr(VI) removal efficiency to 5 mg/L/day compared to 1.66 mg/L/day without the addition of acetate. In both phases of operation (acetate as energy source in first phase and root exudates as energy source in second phase), Cr(VI) reduction at the cathode was nearly 100%, thus indicating that plant-MFC systems can be exploited for the in situ remediation of chromium-contaminated soil and water. Another aspect of such heavy metal removal from aqueous phase in plant-MFC is that a portion of the heavy metal is bioassimilated in the plant tissues. This was confirmed in the aforementioned work, where the presence of total Cr was confirmed in the plant tissue (Habibul et al., 2016). This was again corroborated in a separate investigation utilizing plant-MFC for the removal of the Cr(VI) using a wetland plant-MFC (Guan, Tseng, Tsang, Hu, & Yu, 2019). The high removal efficiency (99%) of Cr(VI) was attributed to bioelectrochemical reduction at the cathode and bioassimilation by the plant investigated. In both the investigations cited above, it was emphasized that the removal of Cr(VI) was achieved with high efficiency without any external power or resource input. Removal of lead (63%), arsenic (100%), cadmium (56%), and chromium (46%) were also observed in the floating phytobed plant-MFC system (Kadam et al., 2018).

9.3 Future scope of research In the HMETs, the removal of organic matter, nutrients, and ECs, such as dyes, herbicides, PAHs, and heavy metals, have been reported. However, the majority of the experiments were conducted in lab- and pilot-scale setups. For field-scale applications, the various issues discussed below need to be addressed. Although the studies have reported considerable removal of ECs, the future works should place an emphasis on the simultaneous removal of by-products released during degradation. Additionally, the degradation rate kinetics need to be further improved so as to reduce the time required for bioremediation. This can be achieved by the augmentation of an external power source to enhance the removal of ECs. The possibility of the use of energy tapped from renewable energy sources like solar and wind for augmentation can be further explored. As discussed, in situ HMETs have the ability to bioremediate the target pollutants at greater depths of media. This has to be addressed in the future research by suitable electrode modifications to simultaneously enhance electrochemical properties and reduce the cost of the system. Field-scale applications should also cater to the long-term stability of the system in terms of removal

In situ bioremediation techniques 251 efficiency and durability of electrodes. Further research can also be focused on the use of HMET as a biosensor for detecting the concentration levels of these target contaminants.

9.4 Summary The applications of the HMETs described in this chapter show that such hybridization of the METs with the natural system can help to achieve the detoxification of contaminated sites through in situ treatment. Natural systems, when augmented with these biopowered electrochemical systems, can act as enhanced in situ remediation systems. In the majority of investigations cited in this chapter it was found that the degradation mechanism was biological attenuation, be it refractory organic compound degradation or immobilization of toxic metals. The application of HMETs has immense potential, which can further be exploited for field-scale applications.

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CHAPTER 10

Gene-targeted metagenomics approach for the degradation of organic pollutants Raghawendra Kumar, Dinesh Kumar, Labdhi Pandya, Priti Raj Pandit, Zarna Patel, Shivarudrappa Bhairappanavar and Jayashankar Das Gujarat Biotechnology Research Centre (GBRC), Department of Science and Technology (DST), Government of Gujarat, Gandhinagar, India

10.1 Introduction The rapid development of industry and urbanization around the world have led to the identification and proper understanding of the relationship between public health and environmental contamination. Growing industries are the key players in many developing countries; unfortunately, they are also major source of environmental pollution. Industries use many different types of chemicals for the processing of raw materials which are economical and effective in a short period of time. In general, they use cheap, poorly or nonbiodegradable chemicals which generate toxicity and industry has ignored the problem. There are numerous reports available showing that industrial waste is highly toxic/ hazardous to human health as well as other living beings. Industrial wastewaters or solid waste contain various types of organic and inorganic pollutants that are hazardous to human health and generate huge environmental pollution (Goutam et al., 2018; Saxena, Chandra, & Bharagava, 2016). Discharge of wastewater into the environment is characterized by a high chemical oxygen demand, biological oxygen demand, total dissolved solids, total suspended solids, and a variety of recalcitrant organic and inorganic pollutants (Bharagava, Chowdhary, & Saxena, 2017; Bharagava, Saxena, Mulla, & Patel, 2017; Gautam, Kaithwas, Bharagava, & Saxena, 2017). There are various organic pollutants generated by industries, such as azo dyes, polyaromatic hydrocarbons (PAHs), phenols, chlorinated phenols, endocrine disrupting chemicals, polychlorinated biphenyls (PCBs), pesticides, and xenobiotics, whereas inorganic pollutants include a variety of toxic heavy metals, such as cadmium, chromium, arsenic, lead, and mercury (Chandra, Bharagava, Kapley, & Purohit, 2011; Saxena & Bharagava, 2017). Before discharging wastewater or solid waste into the environment it should be properly treat

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00010-9 © 2020 Elsevier Inc. All rights reserved.

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258 Chapter 10 because industrial waste contains a high concentration of recalcitrant organic pollutants that are nonbiodegradable in nature. The bioremediation of organic waste using the microbial algae, fungi, and bacteria is an ecofriendly waste management technique to remove the organic and inorganic pollutants that utilizes the inherent potential of microbes to convert highly toxic complex organic pollutants into less toxic forms (Bharagava, Chowdhary, et al., 2017; Bharagava, Saxena, et al., 2017; Maszenan, Liu, & Ng, 2011; Saxena & Bharagava, 2016). This process involves the mineralization and detoxification of waste in a sequential way and the compounds generated after conversion are carbon dioxide, water, and methane (Reshma, Spandana, & Sowmya, 2011). The process of bioremediation totally depends upon the metabolic potential of the microbes (Antizar-Ladislao, 2010). But from time to time the bioremediation/biodegradation of organic waste is very low due to unavailability of a native microbial community to degrade the pollution in a proper time duration. This problem can be solved by using a nonnative microbial community or enzymes to speed up the degradability. However, this process is very tedious as nonnative microorganisms take time to survive in a contaminated ecosystem and then remove the contamination, thus modifying the ecosystem (Techtmann & Hazen, 2016). Gene-targeted metagenomics approach have been applied to identify the potential microbial community that is active in bioremediation/ biodegradation in a contaminated environment. Next-generation sequencing (NGS) technology-based approaches in metagenomics provide reliable and valuable information about active enzymes or genes which play an important role in degradation. Therefore the main objective of this chapter is to provide the basic knowledge of metagenomic approaches and their applications to better understand the microbial community structure and functions during the bioremediation of environmental pollutants in a contaminated matrix.

10.2 Gene-targeted metagenomics The culture-independent gene-targeted metagenomics approach was developed to direct the exploration of the vast genetic diversity of microbial communities (Kumar, Mishra, & Jha, 2019; Yousuf, Kumar, Mishra, & Jha, 2014a, 2014b). This technique has proved to be a highly efficient, accurate, rapid, and sensitive method to analyze the target genes in a complex microbial community, which can bypass or replace the culture-based approach (Simon & Daniel, 2011). However, there is limitation to conclude that a particular gene is a representative for the complexity of the whole microbial community in a particular environment (Suenaga, 2012). Metagenomics helps to identify the variation in genes’ sequence composition, evolutionary rates, genome size, and metabolic capabilities (Dinsdale et al., 2008; Raes, Hermans, Williams, & Eelen, 2007; Von Mering et al., 2007). For better understanding about how microbes react in a different environment, a

Gene-targeted metagenomics approach for the degradation of organic pollutants 259 comparative study of microbial community among the physical dissimilar environment was carried out (Kumar et al., 2019). The term “metagenome” was first time used by Handelsman et al. (1998) for the mixture of microbial genomes extracted directly from an environmental sample (soil, water, sediment).

10.3 Methods used for metagenomics studies Metagenomics studies can be broadly classified into two groups: (1) whole metagenomics and (2) targeted metagenomics, and this latter group can be further classified into a sequences-based approach or function-based on the sequencing of the selected target. Several project based on the random sequencing of microbial domains, such as the bacteria and archaea, and of viruses, have been reported (Sharpton, 2014). Whole metagenome analysis gives information about the enzymes and the metabolites that actively participate in key ecological processes under different environmental pressures. But this information is lacking due to insufficient sequencing data present in common platforms, particular regarding the active enzymes which well work in such conditions (Hemme et al., 2010). When mutations accrue in particular key enzymes they may optimize adaptive phenotypes for a specific niche (Chattopadhyay et al., 2013). A genetargeted metagenomics study also revealed that the same gene occupied by the microorganism which is now in own it may be shifting of the endogenous gene or horizontal gene transfer between the bacteria. A systematic workflow, overview and computational analysis pipeline including the sampling, experiment design based on the tentative objectives are explained in (Fig. 10.1). The gene-targeted metagenomics-based approach is suitable for the collection and construction of a particular gene or a specific group of enzymes that is useful for the particular ecological niche in adaptive evolution (Yousuf, Kumar, Mishra, & Jha, 2014b; Kumar et al., 2019). 1. Gene-targeted (sequence-based): a sequences-based approach is important to understand the distribution of phylogenetic and functional marker genes. In the genetargeted metagenomics study a particular selected DNA pool is sequenced and these selection processes are usually done based on the sequence-based approach or functional-based approach. By the targeting of the sequence-based approach, analysis of selected sequences may provide wide coverage and extensive redundancy of the sequences for target enzymes and reveal specific genome areas directly linked to an ecological function even if the gene abundance is very low (Suenaga, 2012). A polymerase chain reaction (PCR)-based approach has been widely used for the capture of specific genes from the cloning of metagenomics DNA. The advantage of gene-targeted metagenomics is the it is a well-known established technique that requires a primer for a specific gene and a probe can be designed for high-throughput screening. Probes and primer are designed based on the conserved region of the

260 Chapter 10

Figure 10.1 Overview of metagenomics studies in all sample types.

particular gene or gene fragment so that only the known genes will be identified and only the known fragments will be amplified. Overcoming this problem by combining PCR detection of conserved regions and genome walking in the gene flanking regions

Gene-targeted metagenomics approach for the degradation of organic pollutants 261 makes it possible to obtain entire genes or gene fragments. The NGS is more recent technology, such as Roche 454 sequencing (Margulies et al., 2005), the SOLiD system of Applied Biosystems (Bentley, 2006), and the Genome Analyzer of Illumina Miseq (Bentley, 2006), in which the identification of the diverse collection of functional and phylogenetic marker genes can be achieved without the construction of a metagenomics library (Harismendy et al., 2009; Shendure & Ji, 2008). 2. Functional-based: the function-based metagenomics approach provides much more information about the enzyme of interest expressed in some host, such as antibiotic production, and is usually done in BAC and fosmid libraries. The screening of the metagenome library reveals the unidentified gene or gene family, which cannot be detected in a sequences-based approach (Ferna´ndez-Arrojo et al., 2010; Ferrer et al., 2005). Several enzymes or gene clusters were identify based on functional screening of metagenome. PCR screening of a marine sediment metagenome fosmid library of about 11,000 clones using the degenerated primer identified three fosmid DNA fragments which contain important enzymes for sulfate reduction. The screening of 77,000 clones of a soil metagenome library allowed the identification of nine denitrification gene clusters (Demane`che et al., 2009; Ginolhac et al., 2004; Mussmann et al., 2005). A gene-targeted metagenomics study also revealed that the same gene occupy by the microorganism which is now in own it may be shifting of endogenous gene or horizontal gene transfer between the bacteria. 3. Whole genome metagenome sequencing: the whole genome metagenome-based approach is used to target all genes from all organisms found in an environmental sample. The total metagenomics DNA or genomics complement of the particular environment and its community is revealed by genomic sequencing. In this approach the total metagenomics DNA is isolated from an environmental sample followed by fragments using either a physical or enzymatical process to prepare the library. These libraries are sequenced to identify the total genomics contained within the particular samples. Shotgun metagenomics sequencing is a powerful technique which can give accurate information regarding the functional potential of the microbial community but it has limitations related to sequencing depth. To access the complete gene cluster of environmental samples requires deep sequencing. The entire genome of entire microbial community is required for a complete analysis of taxonomy as well as the functional potential of the microorganism. The analysis of the shotgun metagenomics data can be a vast and complex process as it involves accurately assembling and annotating the diverse gene sequences. Many of the genes do not have proper homology in the current database (Delmont, Simonet, & Vogel, 2012). The aim of the numerous studies is to link the microbial communities with their functions and taxonomy using the marker genes. This process is tedious in metagenomics sequencing approach unless sufficient sequencing depth is covered, as required, metagenome assembly of sequences with different K-mers to obtain long contigs and to obtain the metagenome-assembled bins,

262 Chapter 10 which gives more accurate information regarding functional potential of particular species within microbial community (Sharon and Banfield, 2013: Imelfort et al., 2014). Recently, nowadays, the PacBio system and Nanopore sequencing platforms are used for the sequencing of large fragments.

10.4 Bacterial community abundance Quantitative real-time PCR (qPCR) is a precise, highly sensitive, and fast method to determine the bacterial community abundance (copy number) of targeted genes (Dhaene et al., 2010). It is a promising and advanced tool for studying the microbial community’s abundance in an environmental sample (Kabir et al., 2003; Stubner, 2002; Yousuf et al., 2014a, 2014b). In qPCR the amplification of a targeted gene in PCR is observed in real time, and the reaction is performed in two ways: (1) using SYBR Green as a nonspecific fluorescent dye which intercalates with the double-stranded DNA; and (2) using sequencespecific DNA probes labeled with a fluorescent reporter that allows monitoring of amplification. The qPCR approach in gene-targeted metagenomics is unique among the methods of community analysis in environmental sample. This approach gives a quick and relative quantification of specific phylogenetic groups of microorganisms in a given environmental sample. The amount of amplification and fluorescence measured throughout the cycle is directly proportional to the amount of PCR product and fluorescence light emitted (Raeymaekers, 2000). Real-time PCR is more appropriate and accurate in the estimation of gene abundance, which focuses on the logarithmic phase rather than the endpoint abundance of the product accumulation. qPCR is less affected by the reaction or depletion of reagent in the case of amplification efficiency (Gruntzig et al., 2001). A standard curve is prepared in each qPCR assay, using the known concentration of DNA (usually a linearized plasmid). This standard curve is further used for the calculation of gene abundance of the microbial community marker genes or functional genes (Kumar et al., 2019; Yousuf et al., 2014a, 2014b).

10.4.1 Biodegradation pathway involved in the degradation of organic compounds The degradation of organic pollution using microorganisms is not a single-step process. It requires multiple steps and multiple enzymes for the complete removal or degradation of organic compounds. Microorganisms have the capacity to evolve different anaerobic and aerobic pathways for the degradation of aromatic compounds. The process and catabolic pathways involved in degradation require two steps. The first step is the thermodynamic activation of the functional group (benzene ring), and then the cleavage of the functional group.

Gene-targeted metagenomics approach for the degradation of organic pollutants 263 Organic waste

Primary oxygenase enzymes attack in the presence of oxygen

Start degradation by peripheral pathways

TCA cycle

Figure 10.2 Overview of aerobic pathway degradation by microorganisms.

10.4.1.1 Aerobic pathway In the aerobic pathway for the degradation of organic waste, oxygenase plays the crucial role in both steps. Monooxygenase and dioxygenase are actively participating in the degradation of organic waste via the insertion of one or both atoms of oxygenase enzymes into the organic substrate (McLeod & Eltis, 2008). An overview of the microbial degradation by aerobic process is briefly given in Fig. 10.2, describes the enzymatic process of organic waste degradation A bacterial cell naturally evolves to degrade natural occurring organic waste by the use of enzymes or a group of enzymes. But one enzyme or its pathway is not able to degrade all types of waste. Microorganisms can adapt to different environment and modify their own systems to degrade the organic waste. The natural catabolic processes of microorganisms are essential to maintain the global carbon biogeochemical cycle. Diverse genera of bacteria have been reported to degrade toxic organic aromatic compounds, including Acinetobacter, Alcaligenes, Pseudomonas, Nocardiaand, and Rhodococcus, through an aerobic process. The members of Pseudomonas have been extensively studied and have been found to have outstanding performance in the biodegradation of aromatic organic compounds, ranging from benzene to pyrobenzene (Ghosh, Qureshi, & Purohit, 2019). Moreover, different microbial phyla degrade a wide range of hazardous or toxic waste containing numerous kinds of aromatics, aliphatics, heavy metals, and complex hydrocarbons. The use of the gram-negative bacterial strains was found to improve the bioremediation of hydrocarbon-contaminated sludge

264 Chapter 10 (Zhang et al., 2011). Proteobacteria are predominant among the bacteria strains, including Gammaproteobacteria and Betaproteobacteria, and some others strains found to be effective for sewage sludge are Pseudomonas, Advenella, Klebsiella, Mesorhizobium, Bordetella, Brucella, Stenotrophomonas, Ochrobactrum, Achromobacter, Mycobacterium, Raoultella, and Pusillimonas (Ghosh et al., 2019; Tan & Ji, 2010). These strains efficiently utilized the crude oil as an energy and carbon source (Mishra, Bera, & Mandal, 2014). 10.4.1.2 Anaerobic pathways Microorganisms are well-known decomposers of the persistent environment-polluting compounds. Therefore exploiting them for bioremediation purposes offers an ecofriendly approach compared to physical and chemical treatment methods. Industrial pollution due to xenobiotic compounds is a great cause of concern. In a recent study, compounds such as isophthalate/3-carboxybenzoate were degraded by the anaerobic microbe, Syntrophorhabdus aromaticivorans, which further can be utilized for the identification of the characteristic enzyme and can help in reducing the negative effect on environment. This suggests that the metabolic machinery of the microorganism is a potential tool for the removal of the xenobiotic and aromatic pollutants from the environment. Similarly, research on their derivatives differentiate the fundamental approaches adopted by the aerobic and anaerobic processes for their enzymatic degradation (Boll, Geiger, Junghare, & Schink, 2019; Junghare, Spiteller, & Schink, 2019). Metagenomics and metatranscriptomics studies reveal the genes encoding enzymes and transcripts for the alkylbenzene sulfonate degradation pathways in a comparative methodology using RNASeq data analysis of the anaerobic digester. The microbes were found to be taxonomically close to Smithella, Acinetobacter, and Syntrophorhabdus, indicating the relevance of the metabolic active anaerobic species and pathways (Delforno et al., 2019). The bioaccumulation or magnification of these recalcitrant compounds is another challenge posed to the environment due to their toxicity and impact on human health. Metabolic pathways, such as beta-ketoadipate pathways, provide the essential mechanisms for the primary substrate consumption. This is also present in a wide range of taxonomies, such as bacteria and fungi (Bhatt et al., 2019). Anaerobic conditions lack molecular oxygen, therefore the microbial community utilizes inorganic forms of elector acceptors, such as nitrates, sulfates, or metal ions, and generates methane, nitrogen, sulfite gases, and inert forms of compounds which are less soluble in water. Therefore making them less susceptible to microbial degradation and further understanding the degradation pathway mechanism should be useful in the efficient removal of the aromatic compounds from polluted sites (Phale, Sharma, & Gautam, 2019). The metagenomics approach has revealed the degradation potential of the harmful phthalate isomers and sulfate-reducing by Desulfosarcina cetonica, which utilizes the

Gene-targeted metagenomics approach for the degradation of organic pollutants 265 phthalate/sulfate as the only carbon and energy source. The research studies highlight the degradation capacity of the anaerobic microorganisms for bioremediation (Geiger et al., 2019). The strains of denitrifying Pseudomonas spp. were demonstrated to be capable of anaerobic benzoate degradation. They were also reported to grow by using 2-fluorobenzoate as the sole source of carbon and energy (Schennen, Braun, & Knackmuss, 1985). Another study using an in vitro enzymatic assay and 13C-based flux analysis revealed that the thermophilic bacteria, Geobacillus thermoglucosidasius M10EXG, is able to degrade C5 and C6 sugars and is resistant to high ethanol concentrations. These methods have been used to explore the extremophile microbial strains for industrial applications (Tang et al., 2009). Recent advances in metabolic pathway engineering and construction using a systems biology approach, has enabled the quantification of the functional marker’s genes for the assessment of the degradative potential for the specific contaminants by respective microbial species. PCR-based assays or biochemical assays coupled with NGS are key to the discovery and monitoring of the polluted environment. Similarly, the discovery of new enzymatic biocatalysts and microbial regulatory network systems could be used for the design of contaminant-specific degradation pathways using a synthetic and computational biology approach (Bhan, Xu, & Koffas, 2013; Dı´az, Jime´nez, & Nogales, 2013; Rossmassler, Snow, Taggart, Brown, & Susan, 2019). However, the mechanism of anaerobic digestion is still not well understood due to the association of complex microbial communities. Metagenomics is the study of genomes isolated directly from the environmental source, and gives insights into the microbial communities associated with the anaerobic pathways.

10.4.2 Functional metagenomics The NGS technology is useful for the identification of multiple gene sequences from a shotgun metagenomics library. In addition to sequence-based metagenomics, the functional metagenomics approach is a potent, culture-independent method for characterizing environmental pollutant-degrading enzymes, genes, and transcripts. In this approach, a metagenomics library is prepared by cloning gene sequences from the total microbial community genomic DNA extracted from a desired sample in a suitable expression vector. Further, the cloned library is transformed into a host species and assayed for specific organic pollutant degradation capacity by plating on selective media that are lethal to wildtype hosts based on the substrate specificity. The selected inserts from the surviving recombinant, organic, and polyaromatic compound degrading colonies are then processed for high-throughput sequencing. They are subsequently analyzed, genomic feature predicted, assembled, and annotated for their role in organic pollutant degradation from the polluted sites. The most common recalcitrant organic pollutants, such as chlorinated phenols, PAHs, benzene, toluene, organic halide and halogenated hydrocarbons compounds, PCBs, and polybrominated diphenyl ethers, are well-known for their toxic effects on the

266 Chapter 10 soil, environment, and various life forms (Maphosa et al., 2012; Suenaga, Ohnuki, & Miyazaki, 2007) The functional metagenomics libraries can be of two types based on insert size, that is, small insert libraries and large insert libraries. The small insert libraries are created using plasmids with an insert size of 10 kb approximately. The large insert metagenomics libraries are created using fosmid vectors with 25 45 kb inserts, cosmids with 15 40 kb insert vectors, and bacterial artificial chromosomes with 100 200 kb inserts as vectors. The large insert metagenomics libraries are suitable for the identification and characterization of multigene-encoded proteins and enzymatic products, operons, and the entire biochemical pathways and usually utilize low-copy number and inducible vectors. Similarly, large inserts also facilitate gene neighborhood analysis, phylogenetic and taxonomic assignment of sequences. The application of biotechnology based on the intended objectives results in distinctive outcomes, which have been very surprising due to the exploration strategies by the global research community. In this way, the translation application of the science and technology is explosively growing due to the robust genomic information produced and is significantly driving the research priorities. Similarly, the functional metagenomics-based approach has been progressively developed in recent times due to the advances in sequencing technology and automated robotic platforms introduced in the microbial research. The permutated resolve of genome-wide analysis of the whole genome datasets of prokaryotes and eukaryotic microbial communities is to identify the role of specific genes, transcription factors, proteins, metabolic pathways, and biocatalytic enzymes for their role using biotechnological strategies in agriculture, health, and environment (Chakraborty, Wu, & Hazen, 2012; Culligan, Sleator, Marchesi, & Hill, 2014; Lam, Cheng, Engel, Neufeld, & Charles, 2015). The many aspects of the uncharacterized genomic information are intended to identify the novel functional properties and utility in cellular, molecular, and biological compartments. Further, they are intended to distinguish their role in inducible model systems, overexpression, and upscale from pilot production to industrial application and their use as an enzymatic biocatalyst for the remediation applications. The industry effluents from pharmaceuticals, chemical and fertilizers, petrochemicals, leather and tannery, textile dyes, and pigments-based industries are most commonly overloaded with organic and polyaromatic-based compounds and hazardous environmental pollutants. Therefore identifying the key enzymatic biocatalysts for their degradation should be highly useful for the overall reduction in threats to the environment and ecology (Coughlan, Cotter, Hill, & Alvarez-Ordo´n˜ez, 2015). Also, this should prove to be of great significance for industry for the treatment of effluents at the source without further contaminating water bodies and thus allowing a successful ecofriendly approach. In common effluent treatment plants and industry effluent treatment plants (ETPs), the most commonly used treatment methods are chemical in nature, posing a greater risk in the formation of recalcitrant composite solid sludge. So replacing or finding alternative and ecofriendly options is always a welcome step. The major equivalent challenges remaining for the biological treatment methods are their efficiency and efficacy at variable factors

Gene-targeted metagenomics approach for the degradation of organic pollutants 267 such as flow rate, pH, and temperature among many other parameters. The application of functional metagenomics has advantages in terms of the discovery of the significantly improved versions of the natural biocatalysts and their encoding genes from the identified potential sample collection sites (Jayaraman, Thangaiyan, Mani, Nagarajan, & Muthukalingan, 2019; Ufarte, Laville, Duquesne, & Potocki-Veronese, 2015). The most distinct feature of the polluted sites is the inherent high load of pollutants so a hypothetical case arises that microbes present in the highly polluted sites should have the capacity of exclusive remediation of the targeted compounds from the environments. Therefore exploring and mining of the functional microbial diversity of exclusively polluted sites should be an effective strategy for the development of bioremediation technology for organic and polyaromatic compounds. Similarly, in a distinct approach, the microbial community of gut microbiomes of rumens of domesticated and wild animals has been explored for the identification of methanogenic and cellulose-degrading enzymes and their encoding genes (Ufarte´ et al., 2018). Further, the coupling of functional metagenomics with high-throughput proteomics and systems biology tools, metabolomics, flux-balance analysis, synthetic biology, metabolic and catabolic gene interaction networks, microarray chips, and mass spectrometry should be useful in the validation and characterization of the potential genomic and proteomic signatures for the development of efficient bioremediation technology to be useful for organic pollutants. It would be highly recommended to use biochemical and microbiological techniques together with an integrated functional metagenomics approach to capitalize on new technological advancements in biotechnology for environment safeguarding and protection. While research studies indicate the strength and usefulness of functional metagenomics, it has certain definite limitations, such as a gene has to be functional outside its native microbial host to be identified by using the functional metagenomics approach. On several occasions the differences between a recombinant expression host, such as Escherichia coli, and the original host, which obviously may be a different microorganism, essentially do not confer the same phenotype for the same gene of interest. Similarly, the regulatory genes may not have the same phenotype in recombinant expression systems. Therefore having multiple efforts might be required for the optimization of the suitable expression vector system for the gene or enzyme of interest, among many other permutations and combinations. Apart from intended objectives and suitable methodologies for the development of metaomics-based technologies, the whole genome shotgun sequencing approach has been used for total community profiling from the environmental DNA and RNA samples of the industrially polluted sites and municipal dumping sites where unsegregated waste pools to form highly complex mixtures. The challenges of total genomic DNA isolation and spectroscopic quantification from these polluted sites equally remains challenging for the researchers because of the difficulties caused by the recalcitrant impurities present in it. Similarly, the gene-targeted approach has been helpful in the discovery of full-length dioxygenaseencoding genes from a metagenomics study of the contaminated soils (Eyers et al., 2004;

268 Chapter 10

Figure 10.3 Overview of the study of functional metagenomics.

Suenaga et al., 2012). The sampling strategy for the isolation of environmental DNA could be flexible as novel microbial species with specific activity might be from different extreme and diverse sources, such as marine and hot-springs, with desirable characteristics and a culture-based approach allows the study of only a fraction of the species, that is, those that can grow at laboratory conditions in a defined medium. Therefore the culture-independent approach has certain advantages in terms of robustness and genome coverage of the noncultivated microbial species. A tentative overview of the functional metagenomics is explained in Fig. 10.3, describes the quantification and downstream analysis of the targeted gene encoding key enzyme of interest.

10.5 Conclusion The gene-targeted metagenomic-based studies in organic waste can be used to find novel native functional genes or species that can be used as microbial inoculants for the removal of environmental pollutants. Microbes play a significant role in the detoxification and degradation of a variety of organic pollutants and also help in the biogeochemical process for mineral recycling in the ecosystem. Thus prior knowledge of microbes present in contaminated sites helps to achieve a better understanding about the bioremediation mechanism of a specific pollutant and to find out the potential enzyme/genes which are

Gene-targeted metagenomics approach for the degradation of organic pollutants 269 involved in the degradation of an organic pollutant. Gene-targeted metagenomics approaches are accurate, quicker, and highly effective approaches that bypass the limitations of the conventional molecular techniques to reveal the microbial community structure, microbial community composition, and functional gene abundance present in contaminated matrix. NGS technologies are very popular nowadays with molecular biologists/environmental microbiologists because they are easy to handle, scalable, and can generate large amounts of date in a short time. This technology requires various bioinformatic tools to calculate accurate and reliable results of a sample in an informative way. The application of metagenomics approaches and NGS technology for the characterization of environmental microorganism might open a new window to discover or identify novel microorganisms, catabolic genes, and their products (enzymes) for the bioremediation/biodegradation of organic pollutants.

10.6 Future perspective Gene-targeted metagenomics has been proved to be a highly efficient, accurate, rapid, and sensitive method to identify genes or gene clusters in environmental samples. It is useful to create enzyme pools and it also indicates how enzymes or genes adapt to different ecological conditions. Both gene-targeted and functional screening of the metagenome have become well established in recent years. The development of new technology, such as highthroughput screening of desired gene or enzymes, is a continuous process. For example, microarray-based technologies coupled with microfluidic devices, cell compartmentalization, flow cytometry, and cell sorting have been proposed as promising new tools. The application of NGS in metagenomics has provided large sets of data sequences and their better interpretation has revealed the unknown microbial community, and taxonomic and functional diversity in different environments.

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Gene-targeted metagenomics approach for the degradation of organic pollutants 271 Ginolhac, A., Jarrin, C., Gillet, B., Robe, P., Pujic, P., Tuphile, K., Bertrand, H., Vogel, T. M., Perriere, G., Simonet, P., & Nalin, R. (2004). Phylogenetic analysis of polyketide synthase I domains from soil metagenomic libraries allows selection of promising clones. Appl. Environ. Microbiol., 70(9), 5522 5527, 1. Goutam, S. P., Saxena, G., Singh, V., Yadav, A. K., Bharagava, R. N., & Thapa, K. B. (2018). Green synthesis of TiO2 nanoparticles using leaf extract of Jatropha curcas L. for photocatalytic degradation of tannery wastewater. Chemical Engineering Journal, 15(336), 386 396. Gruntzig, V., Nold, S. C., Zhou, J., & Tiedje, J. M. (2001). Pseudomonas stutzeri nitrite reductase gene abundance in environmental samples measured by real-time PCR. Appl Environ Microbiol, 67, 760 768. Handelsman, J., Rondon, M. R., Brady, S. F., Clardy, J., & Goodman, R. M. (1998). Molecular biological access to the chemistry of unknown soil microbes: a new frontier for natural products. Chemistry & biology, 5(10), R245 R249. Harismendy, O., Ng, P. C., Strausberg, R. L., Wang, X., Stockwell, T. B., Beeson, K. Y., . . . Frazer, K. A. (2009). Evaluation of next generation sequencing platforms for population targeted sequencing studies. Genome Biology, 10(3), R32. Kabir, S., Rajendran, N., Amemiya, T., & Itoh, K. (2003). Real-time quantitative PCR assay on bacterial DNA: In a model soil system and environmental samples. J Gen Appl Microbiol, 49, 101 109. Imelfort, M., Parks, D., Woodcroft, B. J., Dennis, P., Hugenholtz, P., & Tyson, G. W. (2014). GroopM: An automated tool for the recovery of population genomes from related metagenomes. PeerJ, 30(2), e603. Jayaraman, S., Thangaiyan, S., Mani, K., Nagarajan, K., & Muthukalingan, K. (2019). Identification of a novel gene through the metagenomic approach to degrade the targeted pollutant. Microbial biodegradation of xenobiotic compounds (p. 204) CRC Press. Junghare, M., Spiteller, D., & Schink, B. (2019). Anaerobic degradation of xenobiotic isophthalate by the fermenting bacterium Syntrophorhabdus aromaticivorans. The ISME Journal, 13(5), 1252. Kabir, S., Rajendran, N., Amemiya, T., & Itoh, K. (2003). Real-time quantitative PCR assay on bacterial DNA: In a model soil system and environmental samples. J Gen Appl Microbiol, 49, 101 109. Kumar, R., Mishra, A., & Jha, B. (2019). Bacterial community structure and functional diversity in subsurface seawater from the western coastal ecosystem of the Arabian Sea, India. Gene, 15(701), 55 64. Lam, K. N., Cheng, J., Engel, K., Neufeld, J. D., & Charles, T. C. (2015). Current and future resources for functional metagenomics. Frontiers in Microbiology, 29(6), 1196. Maphosa, F., Lieten, S. H., Dinkla, I., Stams, A. J., Smidt, H., & Fennell, D. E. (2012). Ecogenomics of microbial communities in bioremediation of chlorinated contaminated sites. Frontiers in Microbiology, 3, 351. Margulies, M., Egholm, M., Altman, W. E., Attiya, S., Bader, J. S., Bemben, L. A., . . . Dewell, S. B. (2005). Genome sequencing in microfabricated high-density picolitre reactors. Nature, 437(7057), 376. Maszenan, A. M., Liu, Y., & Ng, W. J. (2011). Bioremediation of wastewaters with recalcitrant organic compounds and metals by aerobic granules. Biotechnology Advances, 29, 111 123. McLeod, M. P., & Eltis, L. D. (2008). Genomic insights into the aerobic pathways for degradation of organic pollutants. (2008). Norfolk: Caister Academic Press. Mishra, S., Bera, A., & Mandal, A. (2014). Effect of polymer adsorption on permeability reduction in enhanced oil recovery. Journal of Petroleum Engineering, 2014, 1 9. Mussmann, M., Richter, M., Lombardot, T., Meyerdierks, A., Kuever, J., Kube, M., Glo¨ckner, F. O., & Amann, R. (2005). Clustered genes related to sulfate respiration in uncultured prokaryotes support the theory of their concomitant horizontal transfer. Journal of bacteriology., 187(20), 7126 7137, 15. Phale, P. S., Sharma, A., & Gautam, K. (2019). Microbial degradation of xenobiotics like aromatic pollutants from the terrestrial environments. Pharmaceuticals and personal care products: waste management and treatment technology (pp. 259 278). Elsevier. Raes, F., Hermans, D., Williams, J. M., & Eelen, P. (2007). A sentence completion procedure as an alternative to the Autobiographical Memory Test for assessing overgeneral memory in non-clinical populations. Memory, 15(5), 495 507. Raeymaekers, L. (2000). Basic principles of quantitative PCR. Mol Biotechnol, 15, 115 122.

272 Chapter 10 Reshma, S. V., Spandana, S., & Sowmya, M. (2011). Bioremediation technologies. In World congress of biotechnology, India. Rossmassler, K., Snow, C. D., Taggart, D., Brown, C., & Susan, K. (2019). Advancing biomarkers for anaerobic o-xylene biodegradation via metagenomic analysis of a methanogenic consortium. Applied Microbiology and Biotechnology, 103(10), 4177 4192. Saxena, G., & Bharagava, R. N. (2016). Ram Chandra: Advances in biodegradation and bioremediation of industrial waste. Clean Technologies and Environmental Policy, 18, 979 980. Saxena, G., & Bharagava, R. N. (2017). Organic and inorganic pollutants in industrial wastes, their ecotoxicological effects, health hazards and bioremediation approaches. In R. N. Bharagava (Ed.), Environmental pollutants and their bioremediation approaches (pp. 23 56). Boca Raton, FL: CRC Press, Taylor & Francis Group. Saxena, G., Chandra, R., & Bharagava, R. N. (2016). Environmental pollution, toxicity profile and treatment approaches for tannery wastewater and its chemical pollutants. Reviews of Environmental Contamination and Toxicology, 240, 31 69. Schennen, U., Braun, K., & Knackmuss, H.-J. (1985). Anaerobic degradation of 2-fluorobenzoate by benzoatedegrading, denitrifying bacteria. Journal of Bacteriology, 161(1), 321 325. Sharon, I., & Banfield, J. F. (2013). Genomes from metagenomics. Science, 342, 1057 1058. Available from https://doi.org/10.1126/science.1247023. Sharpton, T. J. (2014). An introduction to the analysis of shotgun metagenomic data. Front.PlantSci, 5, 209. Available from https://doi.org/10.3389/fpls.2014.00209. Shendure, J., & Ji, H. (2008). Next-generation DNA sequencing. Nature Biotechnology, 26(10), 1135. Simon, C., & Daniel, R. (2011). Metagenomic analyses: Past and future trends. Applied and Environmental Microbiology, 77(4), 1153 1161. Stubner, S. (2002). Enumeration of 16S rDNA of Desulfotomaculum lineage 1 in rice field soil by real-time PCR with Sybr Green detection. J Microbiol Methods, 50, 155 164. Suenaga, H. (2012). Targeted metagenomics: A high-resolution metagenomics approach for specific gene clusters in complex microbial communities. Environmental Microbiology, 14(1), 13 22. Suenaga, H., Ohnuki, T., & Miyazaki, K. (2007). Functional screening of a metagenomic library for genes involved in microbial degradation of aromatic compounds. Environmental Microbiology, 9(9), 2289 2297. Tan, Y., & Ji, G. (2010). Bacterial community structure and dominant bacteria in the rhizosphere of two endemorelict plants capable of degrading a broad range of aromatic substrates. Applied Microbiology and Biotechnology, 91, 1227 1238. Tang, Y. J., Sapra, R., Joyner, D., Hazen, T. C., Myers, S., Reichmuth, D., . . . Keasling, J. D. (2009). Analysis of metabolic pathways and fluxes in a newly discovered thermophilic and ethanol-tolerant Geobacillus strain. Biotechnology and Bioengineering, 102(5), 1377 1386. Techtmann, S. M., & Hazen, T. C. (2016). Metagenomic applications in environmental monitoring and bioremediation. Journal of Industrial Microbiology & Biotechnology, 43(10), 1345 1354. Ufarte, L., Laville, E., Duquesne, S., & Potocki-Veronese, G. (2015). Metagenomics for the discovery of pollutant degrading enzymes. Biotechnology Advances, 33(8), 1845 1854. Ufarte´, L., Potocki-Veronese, G., Cecchini, D., Tauzin, A. S., Rizzo, A., Morgavi, D. P., . . . Klopp, C. (2018). Highly promiscuous oxidases discovered in the bovine rumen microbiome. Frontiers in Microbiology, 4(9), 861. Von Mering, C., Hugenholtz, P., Raes, J., Tringe, S. G., Doerks, T., Jensen, L. J., . . . Bork, P. (2007). Quantitative phylogenetic assessment of microbial communities in diverse environments. Science, 315(5815), 1126 1130. Yousuf, B., Kumar, R., Mishra, A., & Jha, B. (2014a). Unravelling the carbon and sulphur metabolism in coastal soil ecosystems using comparative cultivation-independent genome-level characterisation of microbial communities. PLoS ONE, 9(9), e107025.

Gene-targeted metagenomics approach for the degradation of organic pollutants 273 Yousuf, B., Kumar, R., Mishra, A., & Jha, B. (2014b). Differential distribution and abundance of diazotrophic bacterial communities across different soil niches using a gene-targeted clone library approach. FEMS Microbiology Letters, 360(2), 117 125. Zhang, Z., Hou, Z., Yang, C., Ma, C., Tao, F., & Xu, P. (2011). Degradation of n-alkanes and polycyclic aromatic hydrocarbons in petroleum by a newly isolated Pseudomonas aeruginosa DQ8. Bioresource Technology, 102, 4111 4116.

Further reading Bell, T. H., Joly, S., Pitre, F. E., & Yergeau, E. (2014). Increasing phytoremediation efficiency and reliability using novel omics approaches. Trends in Biotechnology, 32(5), 271 280. Dholakiya, R. N., Kumar, R., Mishra, A., Mody, K. H., & Jha, B. (2017). Antibacterial and antioxidant activities of novel actinobacteria strain isolated from Gulf of Khambhat, Gujarat. Frontiers in Microbiology, 7(8), 2420. Handelsman, J. (2004). Metagenomics: Application of genomics to uncultured microorganisms. Microbiology and Molecular Biology Reviews, 68(4), 669 685.

CHAPTER 11

Current status of toxic wastewater control strategies Sushma Chityala*, Dharanidaran Jayachandran*, Ashish A. Prabhu and Veeranki Venkata Dasu Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering, Indian Institute of Technology Guwahati, Guwahati, India

11.1 Introduction Polluting the environment has been a customary practice by humans since the dawn of their kind. The early forms of water pollution were mainly due to human and animal wastes. The outbreaks of epidemics caused by toxicity were not understood or were only poorly understood. Environmental pollution has increased greatly over the past few centuries predominately due to the industrial revolution. Although the industrial revolution had positive consequences, especially in reducing human labor, it has inflicted mayhem on the environment. There is no doubt that air and water are the two major facets of the environment have been hugely affected. However, it is beyond the scope of this chapter to examine the features of the air pollution and possible control strategies. Therefore we restrict ourselves to water pollution and the control of toxic wastewater. Toxic wastewater, when consumed, inhaled, or absorbed through the skin, could prove to be fatal. It is indistinguishable from any other type of wastewater unless analyzed chemically. The toxic nature of this type emanates from dangerous pathogens, microbial toxins, and potentially toxic elements, such as cadmium, nickel, lead, chromium, and copper (Xu, 2016). The remediation of these toxic entities is mandatory for treating toxic wastewater. The majority of these toxins are carcinogens, which with continuous exposure can cause cancer.



Contributed equally.

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00011-0 © 2020 Elsevier Inc. All rights reserved.

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276 Chapter 11

11.2 Causes and effects of toxic wastewater pollution Humans are the worst contributors to water pollution. The inputs for the toxic wastewater comes from the various sectors of human activity, such as pulp and paper mills, sugar mills, mining, industries, direct usage of pesticides, and fertilizers in the soil. According to the United Nation’s water statistics, 90% of the wastewater in developing countries and 80% of the wastewater worldwide flows untreated into water bodies. Moreover, 2 million tons of human waste is disposed of into water bodies every day. The food sector has been found to be an obvious contributor to water pollution. It has been found that in both high-income and low-income countries, the food sector contributes about 40% 2 54% of organic water pollutants. Also, nitrogen pollution in rivers has increased roughly by 10% 2 20% globally because of the use of fertilizers in food production (Awaleh & Soubaneh, 2014). Table 11.1 shows the composition of selected parameters in different types of raw toxic wastewater. An increase in nitrogen pollution in rivers can lead to an increased nitrate level in the drinking water, leading to a lethal disorder, that is, methemoglobinemia or the “blue baby syndrome.” Especially in rural parts of Eastern Europe, where agriculture was intensifying, infants were found to suffer from this disorder (Kumar & Puri, 2012). In 1912 the people of Japan around the Jinzu river basin were found to have “Itai-Itai” disease. This disease was due to widespread cadmium poisoning caused by mining (Bernhoft, 2013). Similarly, in 2002 pregnant women were found to have a decreased bone mineral density because of cadmium intake (Ohta, Ichikawa, & Seki, 2002). There are more than a dozen common diseases, a few of which are lethal, caused by water pollution. The most common diseases Table 11.1: Composition of selected parameters in different types of toxic wastewater. Parameter

Unit

Domestic wastewater

Tannery wastewater

Suspended solids Total dissolved solids (TDS) Volatile Compounds COD BOD Phosphate TKN Sulfate Chloride Odor Temperature pH

mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L

350 850 .400 1000 400 15 40 50 100 Pungent 20 6.5 8.5

1244 21,620



C

2533 977 62 118 860 6528 Obnoxious 30 7.9 9.2

Pharmaceutical wastewater 800 1000 500 690 39,000 60,000 3 6 1000 1575

30 7 8

BOD, Biological oxygen demand; COD, chemical oxygen demand; TKN, total Kjeldahl nitrogen. Source: Partly adapted from Akarsubasi, A. T., et al. (2005). Water Research, 39, 1576 1584; Mandal, T., Dasgupta, D., Mandal, S., & Datta, S. (2010). Journal of Hazardous Materials, 180, 204 211; Rawat, I., Ranjith Kumar, R., Mutanda, T., & Bux, F. (2011). Applied Energy, 88, 3411 3424.

Current status of toxic wastewater control strategies 277 are cholera, amoebiasis, dysentery, diarrhea, malaria, hepatitis A, polyomavirus infection, polio, arsenicosis, skeletal fluorosis, and typhoid.

11.3 Current interventions in toxic wastewater control Toxic wastewater control requires strategies at various levels in order to achieve the overall goal of reducing pollution. It is important to take precautions from the start to reduce the toxicity of wastewater. This includes abatement in the use of chemical fertilizers in agriculture, use of proper production methods in industry, and proper disposal of domestic sewage. With the advancement in our understanding and betterment of technology day by day, there is always an augmentation in the treatment schemes of wastewater. The various treatment strategies in wastewater control are dispersed under three main categories: physical, chemical, and biological treatment methods. Biological methods are processes involving biological agents for the treatment of wastewater. Three such novel and promising biological strategies used at present in various industries for treating toxic wastewater have been reviewed thoroughly in this chapter. These methods when compared to the majority of the existing conventional methods are environment-friendly and cost-effective. Furthermore, they efficiently convert the organic matter and utilize the elements as cofactors to form useful bioproducts. The chemical method is a process that uses chemicals such as ozonation/H2O2, ultraviolet (UV) photolysis/H2O2, and photo-Fenton processes for the treatment of wastewater. This method completely removed pesticides, beta-blockers, and pharmaceuticals. Apart from chemical method, intermittent hybrid systems involving both chemical and biological processes showed higher efficacy in removing a few emerging contaminants. Such systems involve an initial ozonation treatment step followed by biologically activated carbon or an ultrasound system (Ahmed et al., 2017). The penultimate section covers the miscellaneous methods available for treating the toxic wastewater. However, the reader is strongly encouraged to go through the literature to understand more about the other interventions in detail.

11.3.1 Treatment using aquatic systems Aquatic systems consist of shallow ponds in which one and/or more species of watertolerant vascular plants such as duckweed or water hyacinths are grown (Tchobanoglous, 1987). Water hyacinth plant systems were capable of removing high levels of chemical oxygen demand (COD), suspended solids, total Kjeldahl nitrogen, and trace phosphorous (e.g., 80%, 81%, 79%, and 89%, respectively) from domestic wastewater (Abdel-Raouf, Al-Homaidan, & Ibraheem, 2012; Orth & Sapkota, 1988). A pond covered with duckweed mat seems to be able to purify wastewater jointly with bacteria. The bacterial decomposition causes anaerobiosis in the water. The duckweed mat

278 Chapter 11

Figure 11.1 Wetland plant scheme for treating wastewater (Samer, 2015). Source: From Qasim, S. R. (1999). Wastewater treatment plants: Planning, design, and operation. Lancaster, PA: Technomic Publishing Company.

observe excess minerals and grows on nitrate, thus removing excess nitrogen compounds from wastewater (Abdel-Raouf et al., 2012). It has been shown that duckweed species such as Lemna and Spirodela are can be used in the treatment process (Culley & Epps, 1973). Fig. 11.1 shows the wetland plants for treating wastewater. The main minerals, such as C, N, and P, will be converted into protein by duckweed since it has the ability to remove the organic matter by utilizing and assimilating them as carbohydrates and various amino acids (Hillman, 1976).

11.3.2 Treatment using microalgae Microalgae cultures can play an interesting role in the treatment of toxic wastewater, especially as a tertiary treatment method. Microalgae species such as Chlorella sp. and Dunaliella sp. were used for toxic wastewater treatment and mass production (Abdel-Raouf et al., 2012). Microalgae are mainly used in the culture of high-value products such as genetically engineered products and pharmaceuticals (Al-Saif, Abdel-Raouf, El-Wazanani, & Aref, 2014; Javanmardian & Palsson, 1991). Wastewater treatment with microalgae is particularly attractive because of their photosynthetic capabilities, converting solar energy into useful biomasses and incorporating the nutrients (i.e., nitrogen and phosphorus) that cause eutrophication in ponds, thus leading to a pollution-free environment (de la Noue & de Pauw, 1988).

11.3.3 Treatment using vermifiltration Vermiculture or worm farming consists of some species of earthworms, such as Eisenia fetida and Lumbricus rubellus, making vermicompost, which is also known as worm compost, vermicast, or worm castings. Worm humus or worm manure is the end-product of the breakdown of organic matter and is considered to be a nutrient-rich biofertilizer and soil conditioner. Vermiculture can be implemented to transform livestock manure, food leftovers, and organic matters into a nutrient-rich biofertilizer (Samer, 2015). Fig. 11.2 shows the schematic view of vermifiltration.

Current status of toxic wastewater control strategies 279

Figure 11.2 Schematic diagram of a vermifilter with the column containing earthworms, whose fruitful gut microbes produce excrement that homogenizes the organic matter (Samer, 2015). Source: Xing M., Li X., Yang J., Lv B., & Lu Y. (2012). Performance and mechanism of vermifiltration system for liquid-state sewage sludge treatment using molecular and stable isotopic techniques. Chemical Engineering Journal, 197, 143 150.

The potential use of earthworms to break down and manage sewage sludge began in the late 1970s (Li, Xing, Yang, & Huang, 2011). The introduction of earthworms to the filtration systems led to them being termed vermifiltration systems and was advocated by Jose´ Toha in 1992 (Li et al., 2008). Vermifiltration is widely used in the treatment of wastewater and appears to have high treatment efficiency, including the synchronous stabilization of wastewater and sludge (Xing, Li, Yang, Lv, & Lu, 2012). Vermifiltration is a feasible method to reduce and stabilize liquid-state sewage sludge under optimal conditions (Xing et al., 2012; Xing, Zhao, Yang, Huang, & Xu, 2011). Vermicomposting involves the joint action of earthworms and microorganisms (Xing et al., 2012) and enhances the breakdown of sludge.

11.3.4 Other interventions in toxic wastewater control Although different contaminants require unique methods of treatment, there are a few contaminants which can be removed by any of the three major treatment methods mentioned above. Cyanide is such a contaminant that can be removed by the use of various methods such as physical treatment, that is, diluting the cyanide through the barren/ freshwater rinse method or by using chemical treatments, such as ozonation, acidification, chlorination, sulfidization, acidification volatilization reneutralization process, or through different biological treatments, viz., phytoremediation and microbial remediation. The

280 Chapter 11 degradation of cyanide also happens naturally through oxidation, photodecomposition, adsorption, and volatilization (Mekuto, Ntwampe, & Akcil, 2016). Another process of wastewater treatment includes the use of bioreactors. One such method is the use of a batch biofilter granular reactor (SBBGR) developed by the Water Research Institute (IRSA) of the Italian National Research Council (CNR) in the last decade. Its ability to combine the attached biomass systems and the periodic systems makes it a potential candidate for treating and reusing municipal wastewater. The SBBGR system efficiently removes suspended solids, COD, and nitrogen. Using this system microbes such as Escherichia coli, Giardia lamblia, total coliforms, Clostridium perfringens, Cryptosporidium parvoum, and somatic coliphages were also reduced to levels less than the permissible amounts given by the World Health Organization. Coupling of the SBBGR system with disinfection processes such as UV radiation and per acetic acid showed promising results in E. coli removal making it suitable for the reuse of the water in agriculture (De Sanctis, Del Moro, Levantesi, Luprano, & Di Iaconi, 2016). Alum, a hydrated potassium aluminum salt, is very well-known for its coagulating property in wastewater treatment (Gungor, Karakaya, Gunes, Yatkin, & Evrendilek, 2016). Conventional methods involving bioreactors for treating wastewater have been performed under batch conditions. However, the continuous systems lack efficiency when it comes to complete conversion. In addition, the batch systems become efficient in converting the biomass, although they might be time-consuming. Manuel (2005) compared different control strategies associated with biomass inhibition during toxic wastewater treatment. They found that discontinuous sequencing batch reactors were superior for treating the wastewater treatment than continuous reactors. The removal of emerging contaminants from wastewater treatment plant effluents was clearly reviewed by Awaleh and Soubaneh (2014). Biological processes involving membrane bioreactors, activated sludge, and aeration processes were found to remove endocrine disruption chemicals—exogenous substances that modify the functions of the endocrine system (RWL Water, 2010; Samer, 2015). Industrial and municipal wastewaters and concentrated slurries have been traditionally treated by aerobic and/or anaerobic biological processes (Eckenfelder, Goodman, & Englande, 1972). Other technologies employ artificial wetlands for the reduction of biological oxygen demand (BOD) and total suspended solids (Moo-Young & Chisti, 1994; Zachritz & Fuller, 1993). Biological phosphate removal from wastewater was carried out using modified activated sludge systems (Hieltjes & Lijklema, 1980; Zheng, Zhao, Zhou, Fu, & Li, 2013). The dairy industry is considered one of the most polluting sectors because of its water consumption and increasing wastewater generation. The effluent from dairy consists of high concentrations of organic compounds and nutrients (Perle, Kimchie, & Shelef, 1995). The conventional treatment techniques include the primary treatment to

Current status of toxic wastewater control strategies 281 remove solids, oils, and fats and then secondary biological treatment to remove organic matter and nutrients, followed by tertiary treatment for polishing (Marquardt et al., 2012). Vourch, Balannec, Chaufer, and Dorange (2008) used a reverse osmosis technique to treat wastewater from the dairy industry and analyzed the reusability of water. Membrane bioreactors were very efficient in treating the dairy wastewater by retaining complete biomass.

11.4 Wastewater reuse Toxic wastewater or any form of wastewater is not only an entity to be treated and discarded but is also an important alternative source of water that can be used in other sectors. The use of wastewater in irrigation and other industrial sectors has been a common practice in both developed and developing countries. However, such retreated water is avoided for domestic uses. In developed countries, much of the wastewater is treated prior to its use in the irrigation of different kinds of crops. It is also used in industries, construction, aquaculture, etc. The use of untreated wastewater may have adverse effects on an ecological system, and these effects need to be identified and assessed (Table 11.2). Table 11.2: Overview of the composition and adverse effects of usage of untreated wastewater in agriculture. Pollutant

Parameter

Impact

Nutrients

Nitrogen, phosphorous, and potassium BOD and COD

Excessive amounts of N and P can cause excessive growth of undesirable aquatic species (eutrophication) Depletion of dissolved oxygen due to the decomposition of biomolecules Can form sludge and plug the irrigation systems

Biodegradable organics Suspended solids SS including volatile solids (SS) Pathogens Bacteria, parasites, and other indicator organisms Stable organics Phenols, pesticides, chlorinated hydrocarbons, etc. Heavy metals Cadmium, nickel, zinc, and mercury Hydrogen ion pH activity Sodium, calcium, and Dissolved boron inorganic substances Residual chlorine Free and combined chlorine

Can cause communicable and parasitic diseases Highly toxic to the environment and persevere for longer periods Bioaccumulation, high toxicity, and other health effects Affects metal solubility and alkalinity of soils Can cause salinity that inhibits plant growth and soil permeability Can cause ground water contamination

BOD, Biological oxygen demand; COD, chemical oxygen demand. Source: Partly adapted from Asano, T. (1987), GeoJournal, 15, 273 282.

282 Chapter 11 Municipal wastewater treatment is a well-developed engineering science and sophisticated techniques are available for treatment (Asano, 1987; Asano & Levine, 1996). In the absence of high concentrations of industrial waste, treatment can be done primarily by sedimentation followed by biological processes, but it is economically unfeasible in developing countries. Land-based wastewater treatment systems are considered to be better compared to the nonland-based systems, provided the land is available at low prices. Waste stabilization ponds are a popular form of land-based treatment system, and basically consist of anaerobic, facultative, and maturation ponds (Juanico, Azov, Teltsch, & Shelef, 1995; Yagoubi et al., 2000). The quality of the effluent depends on the type of pond in which wastewater is treated. Anaerobic and facultative ponds produce an effluent which is suitable for restricted irrigation (for some types of crops). Further treatment of effluent in maturation ponds makes it suitable for unrestricted irrigation (can be used to irrigate all types of crops). However, the disadvantages of land-based treatment systems are the high cost of land and water loss through evaporation.

11.5 Conclusion The bioconversion and treatment of toxic wastewater into harmless substances or reusable water for agriculture and industries has a significant role in reducing environmental pollution and improved resource utilization. Hybrid systems provide more advantages than the individual conventional systems by being more stable, providing superior treatment, being sustainable, and saving energy. Various hybrid systems currently in use are hybrid methods such as physical biological, physical chemical, chemical biological, or physical chemical biological systems. The development of hybrid systems to a large-scale level is understandably the next major direction in wastewater treatment. Both in situ and bioreactor-based treatment processes are experiencing rapid development and increasing deployment in practical applications. The current infancy of many of these bioprocesses will, however, be overcome only through time and extensive research.

Acknowledgments The authors would like to thank the Department of Biosciences and Bioengineering, IIT Guwahati, for providing all the support for successfully completing the chapter.

References Abdel-Raouf, N., Al-Homaidan, A. A., & Ibraheem, I. B. M. (2012). Saudi Journal of Biological Sciences, 19, 257 275. Ahmed, M. B., Zhou, J. L., Ngo, H. H., Guo, W., Thomaidis, N. S., & Xu, J. (2017). Journal of Hazardous Materials, 323, 274 298. Akarsubasi, A. T., et al. (2005). Water Research, 39, 1576 1584.

Current status of toxic wastewater control strategies 283 Al-Saif, S. S. A., Abdel-Raouf, N., El-Wazanani, H. A., & Aref, I. A. (2014). Saudi Journal of Biological Sciences, 21, 57 64. Asano, T. (1987). GeoJournal, 15, 273 282. Asano, T., & Levine, A. D. (1996). Water Science and Technology, 33, 1 14. Awaleh, M. O., & Soubaneh, Y. D. (2014). Hydrology: Current Research. Available from https://doi.org/ 10.4172/2157-7587.1000164. Bernhoft, R. A. (2013). The Scientific World Journal, 2013, e394652. Culley, D. D., & Epps, E. A. (1973). Journal of Water Pollution Control Federation, 45, 337 347. De Sanctis, M., Del Moro, G., Levantesi, C., Luprano, M. L., & Di Iaconi, C. (2016). Sciences of the Total Environment, 543, 206 213. Eckenfelder, W. W., Jr., Goodman, B. L., & Englande, A. J. (1972). Advances in Biochemical Engineering, 2, 145 180. Gungor, K., Karakaya, N., Gunes, Y., Yatkin, S., & Evrendilek, F. (2016). Desalination and Water Treatment, 57, 2413 2421. Hieltjes, A. H. M., & Lijklema, L. (1980). Journal of Environmental Quality, 9, 405 407. Hillman, W. S. (1976). Science, 193, 453 458. Javanmardian, M., & Palsson, B. O. (1991). Biotechnology and Bioengineering, 38, 1182 1189. Juanico, M., Azov, Y., Teltsch, B., & Shelef, G. (1995). Water Research, 29, 1695 1702. Kumar, M., & Puri, A. (2012). Indian Journal of Occupational and Environmental Medicine, 16, 40 44. Li, X., Xing, M., Yang, J., & Huang, Z. (2011). Journal of. Hazardous Materials, 185, 740 748. Li, Y. S., et al. (2008). Ecological Engineering, 32, 301 309. Mandal, T., Dasgupta, D., Mandal, S., & Datta, S. (2010). Journal of Hazardous Materials, 180, 204 211. Manuel, J. (2005). https://www.semanticscholar.org/paper/Control-Strategies-for-treating-Toxic-WastewaterBetancur-Moreno/6e4708f11ab99307da3c288274e5dd6e4213b7c1. (Accessed 30.12.16). Marquardt, L., Rohlfes, A. L. B., de, N., Baccar, M., de Oliveira, M. S. R., dos, N. S. P., & Richards, S. (2012). Tecno-Lo´gica, 15, 79 83. Mekuto, L., Ntwampe, S. K. O., & Akcil, A. (2016). Science of the Total Environment, 571, 711 720. Moo-Young, M., & Chisti, Y. (1994). Resources Conservation and Recycling, 11, 13 24. de la Noue, J., & de Pauw, N. (1988). Biotechnology Advances, 6, 725 770. Ohta, H., Ichikawa, M., & Seki, Y. (2002). Tohoku Journal of Experimental Medicine, 196, 33 42. Orth, H. M., & Sapkota, D. P. (1988). Water Research, 22, 1503 1511. Perle, M., Kimchie, S., & Shelef, G. (1995). Water Research, 29, 1549 1554. Rawat, I., Ranjith Kumar, R., Mutanda, T., & Bux, F. (2011). Applied Energy, 88, 3411 3424. RWL Water. (2010). https://www.rwlwater.com/biological-wastewater-treatment/. (Accessed 30.12.16). Samer, M. (2015). https://www.intechopen.com/books/wastewater-treatment-engineering/biological-andchemical-wastewater-treatment-processes. (Accessed 30.12.16). Tchobanoglous, G. (1987). Aquatic plant systems for wastewater treatment: Engineering considerations (1st ed.). Orlando, FL: Magnolia Publishing Inc. Vourch, M., Balannec, B., Chaufer, B., & Dorange, G. (2008). Desalination., 219, 190 202. Xing, M., Li, X., Yang, J., Lv, B., & Lu, Y. (2012). Chemical Engineering Journal, 197, 143 150. Xing, M., Zhao, L., Yang, J., Huang, Z., & Xu, Z. (2011). Environmental Engineering Science, 28, 619 626. Xu, P. (2016). Ecological Engineering, 87, 150 156. Yagoubi, M., Yachioui, M. E., Foutlane, A., Bourchich, L., Jellal, J., & Wittland, C. (2000). Journal of Water Supply: Research and Technology Aqua, 49, 203 209. Zachritz, W. H., & Fuller, J. W. (1993). Water Environment Research, 65, 46 52. Zheng, C., Zhao, L., Zhou, X., Fu, Z., & Li, A. (2013). Treatment technologies for organic wastewater. London: Intech.

CHAPTER 12

Latest innovations in bacterial degradation of textile azo dyes Shantkriti Srinivasan1, Kanyaga Parameswari M2 and Siranjeevi Nagaraj3 1

Department of Biotechnology, Kalasalingam Academy of Research and Education, Krishnankoil, India, 2Department of Zoology, Cotton University, Panbazar, Guwahati, Assam, India, 3 Nencki Institute of Experimental Biology, Polish Academy of Sciences, Warsaw, Poland

12.1 Introduction Textile and dyeing industries are inevitable in the progress of developing countries. They greatly use the synthetic azo dyes due to their cost-effectiveness, ease of synthesis, versatility, broad applicability, and good physical properties, such as fastening and endurance to fading of color by sunlight, water, and chemicals (Correia, Stephenson, & Judd, 1994). Azo dyes contain
N 5 N
bond along with a cyclic ring (usually benzene or naphthalene). Based on the functional group substituted to this cyclic ring, it forms a plethora of azo dyes with slight variations that confer the desired color and properties. The sobering reality, however, is that the usage of azo dyes is accompanied with a large volume of discharge, mostly as recalcitrant pollutants, which is a major threat to the aquatic ecosystem and pollutes the potable water bodies considerably (Carmen & Daniela, 2012). The effluent from these industries is produced primarily by the wet processing operations which include dyeing, scouring, bleaching, and other finishing processes. Bleaching and dyeing are the maximum water-requiring and yet nonconsumptive processes, thus most water is discharged as hazardous effluents after the processing. Statistically, textile industries are one of the most potential sources and largest producers of wastewater contributing to high chemical oxygen demand (COD) (Li, Zhang, Li, & Cao, 2015). Importantly, they act as reservoirs of metabolites that possess carcinogenic and mutagenic properties (Alves de Lima et al., 2007). Thus, azo dyes pose a major threat to the environment because of their persistent color, high load of concentration, biorecalcitrance, and potential toxicity to other living organisms (Senthil, Muruganandham, & Jaabir, 2014; Senthil, Muruganandham, Kathiravan, Ravikumar, & Jaabir, 2013). A recent study in the cotton knitwear industrial cluster in Tirupur, Tamil Nadu, reported that the constant discharge of inadequately treated textile effluent into water bodies or land has increased the

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00012-2 © 2020 Elsevier Inc. All rights reserved.

285

286 Chapter 12 salinity of the soil rendering it unfit for agriculture, adversely affected fisheries, and contaminated potable water (Jayanth, Karthik, Logesh, Srinivas, & Vijayanand, 2011). The effluent generated by dyeing alone is around 35
175 L/kg with an average of 105 L/kg. In developing nations, a lack of appropriate water resources has prompted farmers to use industrial wastewater which may contain elevated levels of toxins (Kumar et al., 2016). These cumulative circumstances necessitated the urge to establish an efficient treatment of textile industry effluents, in particular the azo dyes. Currently, various physical and chemical methods are aimed at removing the azo dyes (Fig. 12.1). Chemical methods such as oxidation with hydrogen peroxide, ozonation, use of sodium hypochlorite, etc.; photochemical methods such as ultraviolet (UV) treatment in the presence of hydrogen peroxide; physical methods such as adsorption using activated carbon, peat, and silica gel, membrane filtration, ion exchange, etc. (Holkar, Jadhav, Pinjari, Mahamuni, & Pandit, 2016; Robinson, McMullan, Marchant, & Nigam, 2001).

Figure 12.1 Comprehensive representation of azo dye degradation.

Latest innovations in bacterial degradation of textile azo dyes 287 These processes are relatively efficient in treating large volumes of discharged pollutants and reasonably promising, but they are not economical processes; complete degradation is not possible; and not ecofriendly methods, resulting in the accumulation of less dangerous sludge as an end product, which is just an amelioration but not the solution to this devastating problem (Forgacs, Cserhẚti, & Oros, 2004; Zhang, Yediler, Liang, & Kettrup, 2004). Hence, an ecofriendly and efficient alternative for degradation of azo dyes is required. Living organisms, mainly lower order species, are highly capable of degrading azo dyes. Basically, the mechanism of decolorization in these microorganisms is enzymatic degradation involving reductases and oxidases (Wu, Li, & Yang, 2012). Utilization of these naturally present, ecofriendly sources by optimizing processing conditions would be an attractive option. Azo dye degradation by algae, yeast, fungi, and bacteria (El-Sheekh et al., 2009; Lucas, Amaral, Sampaio, Peres, & Dias, 2006; Verma & Madamwar, 2003; Wesenberg, Kyriakides, & Agathos, 2003) has been widely reported (Fig. 12.1). The objective of this review is to update the recent studies in the biodegradation of textile azo dyes, particularly by bacteria as a supplement to existing authoritative reviews (Pandey, Singh, & Iyengar, 2007; Saratale, Saratale, Chang, & Govindwar, 2011). In this review, studies involving a single bacterium for biodegradation of textile azo dyes and analytical characterization of the degraded metabolites were included.

12.2 Bacteria in degradation of azo dyes 12.2.1 Bacteria as source Bacterial degradation of azo dyes has particularly gained importance due to the relatively high levels of biodegradation, cost-effectiveness, and applicability to a broad range of azo dyes (Saratale et al., 2011). Phylogenetically they are ancient organisms with an immense potential for adapting to various conditions, making them favorable candidates. Naturally, in order to confront the stressed environment they secrete enzymes which degrade the toxic substances. Bacteria are easy to grow with a shorter doubling time under aerobic or anaerobic conditions, which induce the expression of various azoreductases (Solı´s, Solı´s, Pe´rez, Manjarrez, & Flores, 2012). Bacterial laccases, an oxidative enzyme involved in azo dyes decolorization, are a better alternative than pH-sensitive fungal laccases due to their stability at high temperature, high pH, and chloride tolerance (Fang et al., 2011; Sharma, Goel, & Caplash, 2007; Telke, Ghodake, Kalyani, Dhanve, & Govindwar, 2011). These bacterial enzymes have heat stability, wider substrate specificity, short production time, and the ability to work at a neutral to basic pH, thereby being suited to treat dye-contaminated textile wastewaters (Hilden, Hakala, & Lundell, 2009; Lee, Choi, & Xu, 2003; Reiss, Ihssen, & Thony-Meyer, 2011; Wells, Teria, & Eve, 2006). As a result, complete

288 Chapter 12 mineralization occurs or less toxic by-products accumulate, which could be processed further. Bacteria also possess the ability to tolerate and degrade higher dye concentrations, albeit the time taken for decolorization increases with the dye concentration (Jadhav, Phugare, Patil, & Jadhav, 2011). These characteristics and the requirement of a simple substrate for growth and maintenance make bacteria ideal candidates for industrial use. The mechanism by which bacteria act on the toxic compound may be anaerobic or aerobic and the consecutive process of reduction is mainly by azoreductase enzymes that convert azo dyes (containing
N 5 N
bonds) to aromatic amines or other less toxic intermediates (Chang & Kuo, 2000; Van Der Zee & Villaverde, 2005). Thus the use of bacteria as a source of ecofriendly elimination of azo dyes is a preferred choice among the alternatives. Toxicity assessment of the treated effluent is performed to gauge the overall effectiveness of the treatment.

12.2.2 Mechanism of azo dye degradation by bacteria Azo dyes are xenobiotic compounds with electron-withdrawing groups, such as azo (N 5 N) linkages and sulfonic groups (
SO3
), rendering them electron deficient and, thereby, less susceptible to oxidative bacterial degradation. Various intracellular and extracellular oxidoreductive enzymes, such as laccase, veratryl alcohol oxidase, azoreductase, nicotinamide adenine dinucleotide-dichlorophenolindophenol (NADH-DCIP) reductase, and tyrosinase, are involved in the azo dye degradation (Jadhav, Kalyani, Telke, Phugare, & Govindwar, 2010; Jadhav et al., 2011; Jadhav, Dawkar, Tamboli, & Govindwar, 2009; Oturkar et al., 2011; Parshetti, Telke, Kalyani, & Govindwar, 2010; Telke, Kalyani, Jadhav, Parshetti, & Govindwar, 2009). The presence of oxygen usually inhibits azo bond reduction activity (Saratale, Saratale, Kalyani, Chang, & Govindwar, 2009). The anaerobic enzyme-mediated machinery is initiated by the cleavage of the azo linkage by an anaerobic azoreductase and electron transfer by a redox mediator, which ferries the electron between the extracellular dye and the intracellular reductase (Chacko & Subramaniam, 2011; Kodam, Soojhawon, Lokhande, & Gawai, 2005; McMullan et al., 2001). Azoreductases require the presence of reducing equivalents, such as flavin adenine dinucleotide, NADH, and nicotinamide adenine dinucleotide phosphate, to mediate this reaction (dos Santos, Cervantes, & Van Lier, 2007; Van der Zee & Cervantes, 2009). However, some strains of certain aerobic bacteria have been shown to reduce the azo linkage by oxygen-insensitive or aerobic azoreductases (Nachiyar & Rajakumar, 2005). The majority of azo dyes are high-molecular-weight compounds, hence making them impermeable to cell membranes, suggesting that the dye reduction might necessarily not involve the intracellular uptake of dyes (Pandey et al., 2007; Pearce, Lloyd, & Guthrie, 2003). Documented evidence suggests that azoreductase activity is associated with various types of

Latest innovations in bacterial degradation of textile azo dyes 289 genes, indicating that it can involve more than one reductase, depending on the microorganism and the culture conditions (Ramalho, Scholze, Cardoso, Ramalho, & Oliveira-Campos, 2002). The role of laccases is very significant in decolorizing azo dyes. Laccase-producing bacteria have been reportedly isolated from various sources, such as laccase-producing Aeromonas hydrophila W-11 from the activated sludge of the textile industry (Wu, Kim, Lee, & Lee, 2010) and Bacillus sp. strain SF from a textile wastewater drain (Gudelj et al., 2001). It has been observed that laccase oxidizes azo dyes without the cleavage of the azo bond and that the intermediate formed by laccase action is further degraded by veratryl alcohol oxidase (Jadhav et al., 2011; Pereira et al., 2009). However, the role of laccases in bacterial cell metabolism is not known yet (Sharma et al., 2007). Some purified forms of laccases have also been employed successfully in the decolorization of azo dyes (Kalme, Jadhav, Jadhav, & Govindwar, 2009). Tyrosinase has been associated with the dye degradation properties of many bacterial species, such as Listeria (Kuberan, Anburaj, Sundaravadivelan, & Kumar, 2011), Bacillus laterosporus (Kurade, Waghmode, & Govindwar, 2011), and Bacillus lentus B1377 (Oturkar et al., 2011). Veratryl alcohol oxidases and lignin peroxidase (LiP) have been reported to be involved in the decolorization of azo dyes by many bacteria including Pseudomonas aeruginosa (Jadhav et al., 2011), Sphingobacterium sp. (Tamboli, Kagalkar, Jadhav, Jadhav, & Govindwar, 2010), and Acinetobacter calcoaceticus (Ghodake, Jadhav, Tamboli, Kagalkar, & Govindwar, 2011).

12.2.3 Phases of treatment Azo dyes’ degradation by bacteria is carried out in anaerobic, aerobic, and microaerophilic conditions. It has been observed that anaerobic nonspecific cleavage of the azo bond leads to the formation of toxic and carcinogenic products (McMullan et al., 2001). The intermediate by-products produced under anaerobic conditions are further degraded aerobically (Rajaguru, Kalaiselvi, Palanivel, & Subburam, 2000). Under aerobic conditions electrons get transported to the electron acceptor, thereby resulting in azo dye reduction and decolorization (Robinson et al., 2001). Many investigators have explored various combinations of aerobic, microaerophilic, and anaerobic phases to achieve the best possible outcomes. Sequential anaerobic
aerobic phases has been explored by many authors (Lodato et al., 2007; Van Der Zee & Villaverde, 2005), wherein the amines formed in the anaerobic phases have been reported to be biotransformed in the aerobic phase (Elbanna, Hassan, Khider, & Mandour, 2010; Steffan, Bardi, & Marzon, 2005). Some researchers have investigated and reported almost 95% to complete biodegradation of dyes under microaerophilic or combined anaerobic/microaerophilic conditions (Chan, Rashid, Koay, Chang, & Tan, 2011; Xu, Guo, & Sun, 2007). The application of a shaking condition

290 Chapter 12 followed by static incubation has also been reported. High decolorization of Golden Yellow HER by B. laterosporus (85%) as well as the consortium GG-BL (100%) was observed in a combined aerobic (12 h) to microaerophilic (12 h) batch process by the action of oxidoreductive enzymes (Waghmode, Kurade, Khandare, & Govindwar, 2011). Limited literature is available on sequential aerobic (shaking condition)
microaerophilic (static incubation) treatment of the azo dyes by single adaptable microorganisms. In one such study seed cultures were inoculated preculture in the respective medium for 12 h and then in a medium with the probe dye for 6 h. These primary stages were maintained in shaking conditions to allow cells to reach late exponential or early stationary growth phase, after which shaking was stopped and the cultures were maintained in static incubation for decolorization to occur (Hsueh, Chen, & Yen, 2009; Zhang et al., 2010). In an experiment on decolorization of the azo dyes Solophenyl Red 3BL and Polar Brilliant Red B by A. hydrophila, it was observed that cell growth was better in shake-flask cultures than under static conditions. However, the percentage of color reduction was very high under static culture than under shaking conditions (Ogugbue, Sawidis, & Oranusi, 2012). This shows that both shaking and static conditions contribute to an efficient decolorization process. For instance, in a study involving degradation of dyes under both shaking and microaerophilic conditions, the pace of decolorization was found to be faster under shaking conditions and it was observed that decolorization reaction depended on the presence of molecular oxygen (Ren, Guo, Zeng, & Sun, 2006). Thus oxygen conditions evidently play a very significant role in the azo dyes bioremediation process and optimizing the aerobic/anaerobic phases is the key to achieving maximum decolorization and degradation of the dyes to nontoxic end products.

12.2.4 Recent studies in bacterial mediated azo dye degradation In recent years several novel bacterial strains capable of azo dyes’ degradation have been isolated. In this review we have included documented studies wherein the degraded metabolites have been characterized. Some of the extensively exploited bacterial strains in the biodegradation of azo dyes are mentioned in Table 12.1. Aeromonas sp. was found to degrade Reactive Black dye to α-ketoglutaric acid with transient accumulation of 4-aminobenzenesulfonic acid (sulfanilic acid), 4-amino, 3-hydronapthalenesulfonic acid, and 4-amino, 5-hydronapthalene 2,7 disulfonic acid (Shah, 2014). A. hydrophila SK16 completely biodegraded Joyfix Red RB by the enzymatic action of azoreductase to form sodium 3-aminobenzenesulfonate, sodium 4- amino-5hydroxynaphthalene-2-sulfonate, sodium (3E,5Z)-4-amino-6-hydroxyocta-1,3,5,7-tetraene-2sulfonate, and sodium (3E,5Z)-4-amino-6-hydroxyhexa-13,5-triene-2-sulfonate (Kumar et al., 2016). Pseudomonas sp. SUK1 has been reported to produce various enzymes and efficiently degrade different dyes. Notable examples are aminopyrine N-demethylase and

Table 12.1: Details of bacteria involved in biodegradation of azo dyes and their degraded metabolites identified in various studies. S. no. Study

Dye

Organism

Year, region

Dry and wet lab techniques

Drimaren Red CL5B

Aeromonas hydrophila 2018, Asia UV
vis spectroscopy, (India) HPLC, FTIR, GC-MS, MTCC 1739 and Enzyme assays Lysinibacillus sphaericus MTCC 9523

Degraded products Naphthalene, sodium 2-amino-3hydroxybenzenesulfonate, sodium (3E)-4aminopenta-1,3-diene-2-sulfonate, 2aminophenol naphthalene, phenol, cyclohexa-1,3-diene, 1,3,5-triazine-2,4 (1H,3H)-dione, 2-aminophenol Phenol, propanal, benzene

1.

Srinivasan and Sadasivam (2018)

2.

Srinivasan (2018) Remazol Yellow RR Lysinibacillus sphaericus MTCC 9523 Srinivasan et al. Reactive Yellow Aeromonas hydrophila (2017) F3R SK16 Kumar et al. Joyfix Red RB Aeromonas hydrophila (2016) SK16

2018, Asia UV
vis spectroscopy, (India) HPLC, FTIR, GC-MS, Enzyme assays 2017, Asia UV
vis spectroscopy, (India) HPLC, FTIR, GC-MS 2016, Asia FTIR, GC-MS (India)

5.

Shah (2014)

Reactive Black

Aeromonas spp.

2014, Asia UV
vis, FTIR, LC-MS (India)

6.

Thakur et al. (2014) Pathak et al. (2014)

Red HE7B

Bacillus sp.

Reactive Black-B (RB-B)

Morganella sp. HK-1

2014, Asia HPLC GC/MS, Enzyme assays 2014, Asia UV
vis, FTIR, GC-MS

8.

Zhao et al. (2014)

Methyl red (MR)

Bacillus sp. strain UN2

2014, Asia LC-MS, Enzyme assays (China)

(2Z)-but-2-ene, 1,3,5-triazine, aniline, and naphthalene Sodium 3-aminobenzenesulfonate, sodium 4- amino-5-hydroxynaphthalene-2sulfonate, sodium (3E,5Z)-4-amino-6hydroxyocta-1,3,5,7-tetraene-2-sulfonate and sodium (3E,5Z)-4-amino-6hydroxyhexa-13,5-triene-2-sulfonate α-Ketoglutaric acid, 4aminobenzenesulfonic acid (sulfanilic acid), 4-amino, 3-hydronapthalenesulfonic acid, and 4-amino and 5-hydronapthalene 2,7-disulfonic acid 8-Nitroso 1-naphthol, 2-diazonium naphthalene Disodium 3,4,6-triamino-5hydroxynaphthalene-2,7-disulfonate, 4aminophenylsulfonylethyl hydrogen sulfate, naphthalene-1-ol, aniline and benzene N,N0 -dimethyl-p-phenyle-nediamine and 2aminobenzoic acid

9.

Karunya et al. (2014) Ng et al. (2014)

Mordant Black 17

Moraxella osloensis

Congo red

Shewanellaxia menensis BC01

2014, Asia UV
Vis, TLC, FTIR, (India) GC-MS 2014, Asia UV
vis, FTIR, GC-MS (China)

Naphthalene, naphthol, naphthoquinone, salicylic acid, and catechol 4,40 -diamino-1,10 -biphenyl and 1,20 diamino naphthalene 4-sulfonic acid

3. 4.

7.

10.

(Continued)

Table 12.1: (Continued) S. no. Study

Dye

Organism

Year, region

11.

Kurade et al. (2013)

Disperse Brown 118

Brevibacillus laterosporus

2013, Asia HPTLC and FTIR, GC(India) MS

12.

Olukanni et al. (2013)

Congo red

Bacillus thuringiensis RUN1

2013, Asia FTIR, GC-MS

13.

Garg and Tripathi Orange II (Acid (2013) Orange 7)

2013, Asia FTIR (India)

14.

Lim et al. (2013)

Acid Orange 7

15.

Kolekar et al. (2012) Jadhav et al. (2012) Mate and Pathade (2012) Mansour et al. (2011) Anjaneya et al. (2011)

Reactive Blue 59

Bacillus cereus (MTCC 9777) RMLAU1 Enterococcus faecalis strain ZL Alishewanella sp. Strain KMK6 Pseudomonas aeruginosa BCH Enterococcus faecalis YZ66 Pseudomonas putida mt-2 Bacillus sp. AK1 and Lysinibacillus sp. AK2

Tamboli et al. (2010) Sarayu and Sandhya (2010) Kalyani et al. (2009a) Kalyani et al. (2009b) Dawkar et al. (2009)

Direct Red 5B (DR5B) Remazol Orange

16. 17. 18. 19.

20. 21. 22. 23. 24.

Remazol orange C.I. Reactive Red 195 Acid orange 52 (AO52) Metanil Yellow

Reactive Red 2 Methyl Orange Navy blue 2GL

Sphingobacterium sp. ATM Pseudomonas aeruginosa Pseudomonas sp. SUK1 isolate Pseudomonas sp. SUK1 Bacillus sp. VUS

2013, Asia (Malaysia) 2012, Asia (India) 2012, Asia (India) 2012, Asia (India) 2011, Europe 2011, Asia (India)

Dry and wet lab techniques

RSM, UV
vis, FTIR, HPLC, FESEM FTIR, HPLC, GC-MS UV
vis, FTIR, HPTLC, GC-MS UV
vis, FTIR,TLC HPLC, GC-MS HPLC

UV
vis, FTIR, TLC, HPLC, GC-MS, Enzyme assay 2010, Asia NMR, FTIR, GC-MS (India) 2010, Asia FTIR, HPLC, NMR (India) 2009, Asia UV
vis, IR (India) spectroscopy, HPLC 2009, Asia GC-MS (India) 2009, Asia UV
vis, FTIR, HPLC, (India) GC-MS

Degraded products N-carbamoyl-2-[(8-chloroquinazolin-4-yl) oxy] acetamide and N-carbamoyl-2(quinazolin-4-yloxy)acetamide Sodium (4- amino-3-diazenylnaphthalene1-sulfonate), phenylbenzene. sodium (4amino-3-diazenylnaphthalene-1-sulfonate), 2-(1-amino-2-diazenyl-2-formylvinyl) benzoic acid Sulfanilic acid

Sulfanilic acid 3-(4,5-Dimethylthiazol-2-yl)-2,5diphenyltetrazolium bromide Benzene, naphthalene-1-carboxylic acid Phthalic acid, trihydroxy-1-naphthalene N,N0 -dimethyl-p-phenylenediamine and 4aminobenzenesulfonic acid Metanilic acid, p-aminodiphenylamine

p-Aminobenzenesulfonic acid and naphthalene-1-ol Methyl metanilic acid and 4-aminobenzoic acid 2-Naphthol 1,4-Benzenediamine, N,N-dimethyl 4-Amino-3-(2-bromo-4, 6-dinitrophenylazo)-phenol and acetic acid 2(-acetoxy-ethylamino)-ethyl ester

25.

Gomare and Govindwar (2009)

Methyl red (MR)

Brevibacillus Laterosporus MTCC 2298

2009, Asia FTIR, GC-MS (India)

26.

Saratale et al. (2009)

Scarlet R

2009, Asia UV
vis, HPLC, FTIR, (India) GC-MS, Enzyme assays

27.

Khalid et al. (2008)

28.

Wang et al. (2008)

Reactive Black-5, Direct Red-81, Acid Red-88 and Disperse Orange-3 Reactive Black 5

Proteus vulgaris NCIM-2027 and Micrococcus glutamicus NCIM2168 Shewanella putrefaciens strain AS96 Rhodopseudomonas palustris W1

2008, Asia Liquid (China) chromatography
mass spectrometry, cyclic voltammetry

29.

Xu et al. (2007)

Fast Acid Red GR

30.

Mansour et al. (2007)

2007, Asia HPLC, GC-MS (China) 2007, UV
vis, toxicity assays Europe

31.

Bafana et al. (2007)

Acids yellow 17, violet 7 and orange 52 Direct Black 38 (DB38)

Shewanella decolorationis S12 Pseudomonas putida mt-2

2007, Asia UV
vis
NIR (India) spectrophotometry, HPLC, GC-MS

4-Aminobiphenyl

32.

Platzek et al. (1999)

Consortium of Cardiobacterium hominis and Pseudomonas stutzeri Staphylococcus aureus

1999, HPLC, GC-MS Europe (Germany)

o-Tolidine (3,3’-dimethylbenzidine, OT), 3,3’-dimethyl-4-amino-4’-hydroxybiphenyl and 3,3’-dimethyl-4-aminobiphenyl

Direct Blue 14

2008, United States

HPLC

Heterocyclic substituted aryl amine, p-(N,N di formyl)-substituted para-di amino benzene derivative and p-di-amino benzene derivative 1,4-Benzenediamine

1-Amino-2-naphthol, sulfanilic acid, and nitroaniline

2-(4-Aminobenzenesulfonyl) ethanol (pbase) and 7-amino-8-hydroxy-1,2naphthoquinone-3, 6-disulfonate-1,2diimime (TAHNDSDP1), 1-2-7-triamino-8hydroxy-3-6-naphthalinedisulfate (TAHNDS), and dihydroxynaphthoquinone-3,6disulfonatediimine (TAHNDSDP2) Catechol and 4-aminobenzoic acid Sulfanilic acid, N,N’-dimethyl-pphenylenediamine, and 4’-aminoacetanilid

294 Chapter 12 NADH-DCIP reductase enzymes which degrade Red BLI (Kalyani, Patil, Jadhav, & Govindwar, 2008); LiP, azoreductase, and dichlorophenol indophenol reductase degrade Methyl Orange into 1,4-benzenediamine, N,N-dimethyl, LiP, and azoreductase degrade Reactive Red 2 to 2-naphthol (Kalyani, Telke, Dhanve, & Jadhav, 2009a). Proteus vulgaris NCIM-2027 (PV) and Micrococcus glutamicus NCIM-2168 (MG) have been reported to degrade Scarlet R into 1,4-benzenediamine with the enzymatic activity of riboflavin reductase and NADH-DCIP reductase (Saratale et al., 2009). Brevibacillus laterosporus MTCC 2298, with the aid of LiP, laccase, aminopyrine Ndemethylase, NADH-DCIP reductase, and malachite green reductase degraded Methyl Red into heterocyclic substituted aryl amine, p-(N,N di formyl)-substituted para-di amino benzene derivative, and p-di amino benzene derivatives (Gomare & Govindwar, 2009). Bacillus sp. VUS has been reported to degrade Navy Blue 2GL into 4-amino-3-(2-bromo4,6-dinitro-phenylazo)-phenol and acetic acid 2-(acetoxy-ethylamino)-ethyl ester by LiP, laccase, and reductases (Dawkar, Jadhav, Ghodake, & Govindwar, 2009). Veratryl oxidase, laccase, DCIP (2,6-dichlorophenol-indophenol) reductase, riboflavin reductase, and azoreductase polyhydroxyalkanoate synthase from Sphingobacterium sp. ATM degrade Direct Red 5B into p-amino benzenesulfonic acid and naphthalene-1-ol (Tamboli et al., 2010). Azoreductase and NADH-dichlorophenol-indophenol reductase from Alishewanella sp. strain KMK6 degrades Reactive Blue 59 into 3-(4,5-dimethylthiazol-2-yl)-2,5diphenyltetrazolium bromide (a yellow tetrazole) (Kolekar & Kodam, 2012). Laccase and azoreductase from Bacillus thuringiensis RUN1 degraded Congo Red into sodium (4-amino-3-diazenylnaphthalene-1-sulfonate) and phenylbenzene. Sodium (4-amino-3diazenylnaphthalene-1-sulfonate) was further oxidized by the orthocleavage pathway to yield 2-(1-amino-2-diazenyl-2-formylvinyl) benzoic acid (Olukanni et al., 2013). Azoreductase, laccase, and NADH-DCIP reductase from Bacillus sp. strain UN2 degraded Methyl Red into N,N0 dimethyl-p-phenylenediamine and 2-aminobenzoic acid (Zhao et al., 2014). Azoreductase and laccase enzymes from Bacillus sp. isolate degraded Red HE7B into 8-nitroso 1-naphthol, 2-diazonium naphthalene (Thakur et al., 2014).

12.2.5 Analytical methods in azo dye degradation Analytical methods are significant in monitoring the process of biodegradation by estimating the degraded compounds. Widely used analytical methods are UV
Visible spectroscopy (UV
Vis), thin-layer chromatography, high-performance layer chromatography, high-performance thin-layer chromatography, Fourier transform infrared spectroscopy, liquid chromatography
mass spectrometry, gas chromatography
mass spectrometry (GC-MS), and nuclear magnetic resonance (NMR) spectroscopy. Scope of each analytical method is given in Table 12.2. Azo dyes and their intermediates obtained through biodegradation absorb different wavelengths in the UV
Visible region as assessed

Latest innovations in bacterial degradation of textile azo dyes 295 Table 12.2: Analytical techniques in bacterial azo dye degradation studies. Analytical methods

Rationale

UV
vis spectroscopy (UV
vis)

Azo dye and their processed intermediates obtained after biodegradation absorb different wavelengths in the UV
visible region Thin-layer chromatography (TLC)/high-performance Chromatographic separation of azo dye and their thin-layer chromatography (HPTLC)/highprocessed intermediates performance liquid chromatography (HPLC) Fourier transform infrared spectroscopy (FTIR) Nature of interactions in the dye based on their functional group moieties can be found by Fourier transform infrared spectroscopy Liquid chromatography
mass spectrometry Mass spectrometry is coupled with chromatography (LC-MS)/gas chromatography
mass to measure molecular weight of the metabolites spectrometry (GC-MS) formed and identify mass/charge ratio as result of biodegradation Nuclear magnetic resonance spectroscopy Structural details of metabolites formed as a result of biodegradation can be elucidated

by UV
Vis spectroscopy. Chromatography can be used for the separation of azo dyes and their processed intermediates. The nature of interactions in the dye based on their functional group moieties can be identified by infrared spectroscopy. In GC-MS, MS is coupled with chromatography to measure the molecular weight of the metabolites formed as the result of biodegradation. NMR can be used to elucidate the structural details of metabolites formed after biodegradation.

12.2.6 Analysis of efficiency of bacterial dye degradation by toxicity tests It is a well-established fact that azo dyes are toxic to the environment and living beings in many ways. The bacterial degradation of azo dyes and the associated toxicity can be due to different mechanisms, such as the reduction and cleavage of the azo bond linkage, resulting in the production of aromatic amines that after metabolic oxidation form electrophilic compounds which covalently bind to DNA. Azo dyes with free aromatic amine groups may be metabolically oxidized without cleavage of the azo bonds. Since these mechanisms are dye specific, the bioremediation system becomes unpredictable with regard to the toxicity levels (Saratale et al., 2011). Toxicity depends on the chemical structure and nature of the particular dye and slight variations in the same can alter the toxicity (Solı´s et al., 2012). For instance, 3-methoxy-4-aminoazobenzene is known to cause hepatic cell cancer in mice and induces mutations in bacteria, while 2-methoxy-4aminoazobenzene is noncarcinogenic and a weak mutagen in bacteria (Ferraz et al., 2011). It has also been observed that metabolites produced by dye degradation are in many cases more toxic than the dye itself (Solı´s-Oba, Eloy-Jua´rez, Teutli, Nava, & Gonza´lez, 2009).

296 Chapter 12 The results of reported cytotoxicity studies are not unequivocal in the sense that some investigators have found that the degradative products are more toxic than the parent dye; for example, the toxicity of Acid Violet 7, a potent mutagen, is found to apparently increase after static biodegradation by Pseudomonas putida due to the production of 4-aminoacetanilide and 5-acetamido-2-amino-1-hydroxy-3,6-naphthalene disulfonic acid (Mansour et al., 2010). However, investigations conducted by some authors suggest that the degraded metabolites are less toxic than the parent compound, implying detoxification of the dyes. For instance, in a study conducted by Jadhav et al. (2011) it was found that biodegradation of the dye Remazol red by P. aeruginosa BCH resulted in the formation of metabolites of a less toxic nature than the parent dye. These studies strongly suggest that the toxicity levels can differ with each microorganism and the dyes involved and therefore toxicity evaluation must be performed for every single biodegradation system. It can be assessed by measuring the toxicity of the dyes before and after the treatment. Various evaluation systems like phytotoxicity, cytotoxicity, microbial toxicity, genotoxicity, and oxidative stress response in plants can be employed for toxicity studies. In phytotoxicity studies, performance of the processed dye intermediates are evaluated by their ability to improve seed germination in comparison with the parent dye (Dawkar et al., 2009; Gomare, Tamboli, Kagalkar, & Govindwar, 2009; Kurade, Waghmode, Kabra, & Govindwar, 2013; Lim, Bay, Aris, Majid, & Ibrahim, 2013). In cytotoxicity studies, the effect of processed dye intermediates on cell viability (usually by MTT assay) is assessed in comparison with the unprocessed dye. In particular, parameters such as cell proliferation rates and the extent of DNA damage (usually by Comet assay) are measured to interpret the toxic effect of the intermediates (Kolekar & Kodam, 2012). Microbial toxicity studies using known cultures is a better approach as the results can be reproduced and compared. Selecting the same microorganism as the decolorizer is more effective as the results will also reflect the efficiency of the degradation operation and appropriateness of the biodecolorizer selected (Chen, 2006).

12.3 Computational inputs in enhancing biodegradation 12.3.1 Choice of strains: adapted versus nonadapted strains It is evident from various documented literature mentioned in the previous sections that a considerable number of bacterial strains isolated from the areas surrounding the textile industries (Table 12.1), which have adapted themselves to the effluent-enriched environment, have shown the potential to decolorize azo dyes and that the enzyme activities are associated with different types of genes. Various bacteria of Bacillus, Aeromonas, and Pseudomonas genus isolated from different effluent sites have been

Latest innovations in bacterial degradation of textile azo dyes 297 reported for decolorizing a variety of dyes but isolation of these cultures is a laborious and time-consuming process (Banat et al., 1996). Effective bioremediation involves the optimization of environmental conditions for ideal microbial growth and faster degradation activity, thereby enzymatically attacking the pollutants and converting them to harmless products. Hence these bacteria can degrade environmental pollutants under optimized physicochemical conditions (nutrients, pH, aeration, electron donors or acceptors, etc.). The bacteria can be indigenous to a contaminated site or isolated elsewhere and brought to the site where they transform the toxicants through their metabolic processes (Shinde, 2013). A. hydrophila isolated from different geographic regions and environments has been reported for the degradation of various dyes (Bouraie & Din, 2016; Chen, Wu, Liou, & Hwang, 2003; Hsueh et al., 2009; Idaka, Ogawa, Horitsu, & Tomoyeda, 1978; Srinivasan, Shanmugam, Surwase, Jadhav, & Sadasivam, 2017; Yatome, Ogawa, Itoh, Sugiyama, & Idaka, 1987). The use of appropriate strains under optimized environmental conditions can degrade any organic compound naturally or by evolution and adaptation (Gayle, 1952). Thus, bacteria like A. hydrophila, which contain oxidoreductive enzymes that have been isolated and identified to possess good biodegrading activity, can be used to decolorize and degrade a variety of dyes through their metabolic degradation pathways (Rathoure, 2017). It has been established that the horizontally mobile bacterial gene pool plays a significant role in the adaption to stressful environments due to the presence of a variety of chemical compounds (Frost, Leplae, Summers, & Toussaint, 2005; Thomas & Nielsen, 2005). Genes for catabolic enzymes are carried by mobile genetic elements, such as transposons and plasmids, or may be present in the chromosome itself, and horizontal gene transfer aids in the rapid microbial transformation of xenobiotic compounds (Sinha et al., 2009). However, some authors have reported azo dyes decolorization by nonadapted bacterial strains isolated from sources other than the textile effluents. In an experimental study conducted by Leena and Raj (2008) the potentials of textile effluent adapted and nonadapted soil isolates to decolorize Reactive Black-B were investigated. The results showed that, though to a lesser degree, decolorization was also accomplished by the nonadapted strains. It has also been demonstrated that plasmids extracted from dye-degrading bacterial strains isolated from an adapted environment were able to transform non dyedegrading strains successfully, conferring on the transformed cells the ability to degrade dyes, indicating that the genes responsible for decolorization are extrachromosomal (Gounder, Jeyakumar, Devi, Singh, & Iyer, 2015). It is empirical that textile effluent-adapted soil bacteria found in polluted sites are efficient in degrading azo dyes. However, screening, identification, and optimization processes are laborious and do not always guarantee the isolation of successful strains that could be translated to industrial utilization. On the other hand, based on the literature and enormous curated datasets already available in culture collections and in silico predictive databases, selecting a suitable characterized nonadapted strain to test their potential enzyme inductivity for azo dye degradation is worth trying, thereby reducing the preliminary screening time. The nonnative strains can also reduce any dye containing an

298 Chapter 12 aromatic ring. A study conducted by Olukanni, Osuntoki, and Gbenle (2006) showed that nonadapted bacterial strains of Bacillus exhibited a relatively higher potential than the textile effluent-adapted isolates. Plasmid screening in both the adapted and nonadapted strains yielded negative results. Therefore it is inferred that genes coding for dye-degrading enzymes need not necessarily be coded by plasmids. They may be chromosomal and inherently present in these microorganisms and get overexpressed under stress conditions in the presence of toxic dyes resulting in their degradation. Thus well-characterized nonadapted strains possessing azo dye-degrading abilities naturally or through genetic manipulation can be regarded as potential candidates for application in textile effluent bioremediation

12.3.2 In silico analysis as a valuable tool Bioinformatics has offered a faster and more robust way of screening the target for bioremediation than any other technique. Modern systems biology integrating homology modeling, molecular docking, and dye-enzyme interaction studies can assist in bioremediation strategies by forecasting various parameters like nature, interactions, structure, and function of enzymes, theoretical mechanisms, and toxicity of ligands and proteins to screen as well as elucidate an efficient technology transfer to real-time setup (Sridhar & Chandra, 2014). It can be utilized for screening of a pollutant’s susceptibility to degradation by already characterized enzymes (Suresh, Kumar, Kumar, & Singh, 2008). The in silico approach can predict the chemical nature of the contaminant, novel xenobiotic biodegradation pathways, and microorganisms capable of biotransformation at the cellular and environmental level (Jaimini, Shabnam, & Sarkar, 2012). A better remediation approach uniting theoretical in silico techniques followed by experimental confirmation is suggested as a time and cost-effective method since a manifold increase in xenobiotics makes it impractical to screen them all experimentally for bioremediation (Reena, Dhall, Kumar, & Kumar, 2014). Understanding of the substrate specificity provides an opportunity to predict the likelihood of degradation. As the substrate specificity for dye-degrading enzymes laccase and azoreductase isolated from different sources varies, it can be utilized for reducing different pollutants. Although docking studies are extensively used for drug discovery, their usage in predicting targets for bioremediation will provide a new impact (Drews, 2000; Sridhar, Chinnathambi, Arumugam, & Suresh, 2013). Visualizing the interaction of dyes with enzymes computationally will provide deep insights into the mechanism of biodegradation. In silico techniques effectively utilize computational algorithms to perform virtual screening of protein
ligand interactions. Structure-based virtual screening has been successfully applied in various fields of biology. Research on azo dye biodegradation can potentially use docking algorithms to identify the ligands that fit the target protein under investigation (Kroemer, 2007). Some studies have successfully applied the in silico techniques of docking and catalytic site residue analysis to identify and study

Latest innovations in bacterial degradation of textile azo dyes 299 the enzymes involved in the biodegradation of various azo dyes curbing the need for expansive preliminary screening of microbial strains that may have selective applications (Srinivasan et al., 2017). A study by Ramanathan, Shanthi, and Rao (2009) identified catalytic residues of azobenzene reductase from Bacillus subtilis and also its function, conservation, hydrogen bonding, B-factor, solvent accessibility, and flexibility of the residues. Crucial amino acid residues having pro or inhibitory or enhancing effects on the enzyme
dye interaction are hypothesized by molecular docking studies. Gao, Ding, Shao, Xu, and Zhao (2015) showed that in the GtAZR gene derived from G. thermoglucosidasius C56-Y593, amino acids Asp-79 and Thr-80 are responsible for azoreductase activity. Another docking study by Wang, Liu, and Qin (2012) revealed that Acid Yellow 23 interacted with the His 57 and Lys 224 residue of trypsin leading to the inhibition of enzyme activity. Also, a study by Philem and Adhikari (2012) showed through a docking study that alr2106, alr1063, and alr2326 have the best docking results, which would enhance the azoreductase activity and give better dye degradation. Haghshenas, Kay, Dehghanian, and Tavakol (2016) employed molecular dynamics simulation to study the interaction between AzrC of mesophilic Bacillus sp. B29 and five azo dyes. In docking studies of flavin mononucleotide -dependent NADH-azoreductase enzymes of Escherichia coli with azo dyes to analyze the dye-enzyme affinity, Thakuria, Jungai, and Adhikari (2015) predicted that Phe-172, Glu-174, Lys-145, Asp-146, and Lys-169 were the catalytic sites in the enzyme. It is clear that higher efficiency in degradation is the most desired goal of all studies. In summary, there are two possibilities to achive this. One is to look for novel adapted isolates from the effluent-polluted sites. The other is to improve the enzyme from wellcharacterized strains which are commercially available and then increase their degradation capacity with computational predictions for preferable mutants. The former option of looking for adapted isolates does not always guarantee the identification of an efficient isolate and biochemical characterization of these isolates is time consuming. However, the nonadapted, well-documented strains can be employed with improved biodegradation capability by advances in the technology to make highly potent recombinant enzymes. Dehghanian, Kay, and Kahrizi (2015) showed that various mutants of AzrC (azoreductase from mesophilic gram-positive Bacillus sp. B29) are made using Molegro Virtual Docker 6.0 and assessed for enhanced affinity in enzyme
dye interactions based on AutoDock 4.2 free binding energy. Results showed AzrC quadruple mutations with higher biodegradation capacity which could be tested experimentally. The most recent development in the area of bioremediation of azo dyes is the use of computational methods like using ab initio, density functional theory calculation, and computational fluid dynamics modeling (Laurence, Uribe, & Utyuzhnikov, 2005; Mahmoodi, Arami, & Gharanjig, 2009). The application of in silico methods has been

300 Chapter 12 reported in laccase-mediated remediation of dyes (Sridhar et al., 2013). Recently, Srinivasan, Sadasivam, Gunalan, Shanmugam, and Kothandan (2019) and Srinivasan & Sadasivam (2018) have reported a combinatorial approach of in silico enzyme-docking studies followed by microbial bioremediation of textile azo dyes for corroboration. In another study two novel strains of Lysinibacillus sphaericus SK13 and A. hydrophila SK16 were isolated and in silico enzyme modeling and enzyme
dye interaction analysis of the enzymes laccase and azoreductase were performed (Kumar et al., 2016). The results of the in silico analysis based on docking score were correlated and found to be in agreement with the biological decolorization percentage obtained by UV
Vis spectral analysis. This showed that the in silico molecular docking approach can be applied to large-scale screening of bacterial strains to choose the best decolorizer among given samples, eliminating the need for time-consuming and cumbersome preliminary screening by the conventional methods of analyzing the decolorization percentages by various dye and bacterial strain combinations. In conclusion, the integration of in silico methods, particularly docking studies, with in vitro experiments is the most effective novel approach in this field of bioremediation.

12.4 Alternative front-runners: fungi, yeast, and algae-mediated azo dye degradation Filamentous fungi produce a wide range of intra- and extracellular enzymes which help them to metabolize different carbon and nitrogen sources that allow them to survive in varied environmental niches including dye effluents (Humnabadkar, Saratale, & Govindwar, 2008). This ability of fungi to degrade a wide range of organic compounds is due to the production of nonspecific ligninolytic enzymes like LiP, manganese peroxidase (MnP), and laccase (Christian, Shrivastava, Shukla, Modi, & Vyas, 2005). Lignin peroxidase oxidizes nonphenolic compounds while MnP and laccase oxidizes phenolic compounds (McMullan et al., 2001). The most explored fungi for azo dye biodegradation is the white-rot fungi, Phanerochaete chrysosporium. Other fungal species which have attracted considerable interest are Trametes (Coriolus) versicolor (Champagne & Ramsay, 2005) and Aspergillus ochraceus (Humnabadkar et al., 2008; Saratale, Kalme, & Govindwar, 2006). But there are certain intrinsic limitations presented by fungal bioremediation of azo dyes. These include the extended growth cycle, the need for nitrogen-limiting conditions, the maintenance of fungi in bioreactors (Saratale et al., 2011; Stolz, 2001), and the sensitivity of fungal laccases to high pH (Sharma et al., 2007). Various species of yeasts have been explored for their ability to decolorize azo dyes by adsorption, enzymatic degradation, or an amalgamation of both (Solı´s et al., 2012). In their experiment with new fungal isolates Yang et al. (2003) have observed that the yeasts Debaryomyces polymorphus and Candida tropicalis completely decolorize Reactive Black 5

Latest innovations in bacterial degradation of textile azo dyes 301 and detected manganese peroxidase (MnP) activity. The presence of azo dyes induces the enzymatic machinery and results in the expression of enzymes such as MnP and tyrosinase (Pajot, Farin˜a, & de Figueroa, 2011). Saccharomyces cerevisiae has been reported to bioaccumulate Remazol Blue and Black B textile dyes when grown in molasses (Aksu, 2003). Cyanobacteria or algae are present in many habitats with reports of algal growth in industrial effluents (Dubey, Dubey, Viswas, & Tiwari, 2011). Algae are believed to degrade azo dyes via induced azoreductase activity (Vijayaraghavan & Yun, 2007). Decolorization by algae is brought about by the assimilation of chromophores to produce biomass, decolorization by CO2 and H2O transformation, and adsorption of chromophores on algal biomass. Yan and Pan (2004) observed that Chlorella pyrenoidosa, Chlorella vulgaris, and Oscillateria tenuis are able to degrade more than 30 azo compounds into simpler aromatic amines, whereas Acuner and Dilek (2004) suggested that many species of Chlorella and Oscillatoria are capable of degrading these aromatic amines into simpler forms.

12.5 Future perspective This review is an update to the existing reviews in the field of biodegradation of azo dyes to fill the knowledge gap and also expand our understanding in the progress of textile azo dye biodegradation. The synergistic use of abiotic (physical and chemical) and biotic (microorganism) methods to eliminate azo dyes are suggested (Harrelkas et al., 2008; Lucas, Dias, Sampaio, Amaral, & Peres, 2007; Nigam, Armour, Banat, Singh, & Marchant, 2000; Ulson, Bonilla, & de Souza, 2010; You & Teng, 2009). Also a systems biology approach of using in silico techniques can be incorporated to save time, cost, and energy, thus reducing the preliminary screening process, as specific enzymes are responsible for dye degradation. Modeling and docking of enzymes from organisms selected based on previous literature and availability in lab with dye structures is a novel yet efficient approach to select the most efficient organism
dye combination. Also software can be used as tools to perform phylogenetic analysis, degradative pathway prediction, etc. The future prospects of in silico methodology include the creation of a dynamic computational interface/platform/ tool, which researchers could use to virtually analyze the azo dye degrading proteins of a particular organism, perform enzyme modeling, and enzyme
dye docking studies. The results of the in silico studies can then be used to confirm the dye structure and select the best suited organism available in their lab/locality to perform the microbial bioremediation. Finally, it is essential to analyze the nature of the degraded compounds by toxicity analysis before implementation of the bioremediation process on a large scale. Several studies suggest that transformation of laboratory research to real-time textile industry effluents would be an attractive, ecofriendly, and economical option to restore polluted aquatic systems.

302 Chapter 12

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6145.

CHAPTER 13

Development in wastewater treatment plant design Bapi Mandal*, Anwesha Purkayastha*, Ashish A. Prabhu and Veeranki Venkata Dasu Biochemical Engineering Laboratory, Department of Biosciences and Bioengineering Indian Institute of Technology Guwahati, Guwahati, India

13.1 Introduction Design of wastewater treatment plants (WWTPs) and wastewater treatment processes/ systems have been evolving because the objectives of wastewater treatment and related regulations are changing constantly. The prime focus of wastewater treatment is to let human and industrial effluents to be disposed with adequate measures to reduce health risks and environmental hazards (Batstone, 2002). An effective way of exploiting wastewater is using it for irrigational purposes; as this way it shall serve as both disposal and utilization. However, prior using it for any purpose, an appropriate treatment is provided to municipal wastewater. Municipal wastewater usually contains the following major constituents: (1) colloidal, suspended, and floatable material; (2) oxygen demanding materials (e.g., organics); (3) microorganisms (pathogens); (4) nutrients (N and P); (5) hazardous constituents (e.g., xenobiotic compounds, heavy metals, etc.); (6) contaminants of emerging concern; and (7) energy-rich material (e.g., organics). Historically, the major purpose of wastewater treatment was for removal of constituents (1) to (3) before the 1970s, for that of (1) to (4) before the 1980s, for that of (1) to (5) plus development of a “clean sludge” since the 1990s, for that of (1) to (6) since the 2000s, and of (1) to (6) plus recovery of (4) and (7) since 2010s (Zhang, Surampalli, Tyagi, & Benergi, 2017). Conventional wastewater treatment processes primarily consist of three major steps: (1) primary treatment; (2) secondary (biological) treatment (including aerobic and/or anaerobic); and (3) tertiary treatment (Koivunen & Heinonen-Tanski, 2005). However, there have been evidential developments in the treatment plant design for better and effective effluent treatment. Therefore to reflect the evolution in design of WWTPs and related treatment processes and 

Contributed equally

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00013-4 © 2020 Elsevier Inc. All rights reserved.

311

312 Chapter 13 systems, this section discusses the conventional WWTP followed by the recent modifications in them.

13.1.1 Conventional wastewater treatment technology 13.1.1.1 Primary treatment The primary treatment of wastewater involves physical processes such as screening, grit removal and sedimentation (Kurbiel, 1998). 13.1.1.1.1 Screens

A screen is a device consists of openings of uniform size. Screens are used to retain the solid wastes found in the influent wastewater to the treatment plant. The primary function of screens is to remove coarse materials from the flow stream that might subsequently damage the process equipment and lower overall treatment process. 13.1.1.1.2 Grit chambers

Grit chambers are provided to protect the mechanical instrument from abrasion, reduce accumulation of heavy deposits in pipelines and channels and to reduce the frequency of digesters. Grit chambers may be placed either before or after the screens. 13.1.1.1.3 Primary settlers or sedimentation

The principle of primary settlers is based on sedimentation due to gravity. Majority of the solid wastes remain to be suspended in a flowing sewage. To that end, stopping the flow of the sewage would settle down the solid wastes. The sedimentation of the particles depends on the size, shape, specific gravity of particles and velocity of sewage. 13.1.1.2 Secondary treatment Secondary treatment mainly consists of biological and chemical treatments. We would prime our focus on biological systems with aerobic and anaerobic treatment systems. A basic difference between aerobic and anaerobic wastewater treatment processes have been listed in Table 13.1. 13.1.1.2.1 Aerobic treatment system

Aerobic biological treatment of wastewater can be further sub-categorized into suspended growth and attached growth systems. 1. Suspended growth system In a suspended-growth system like activated sludge processes, the waste flows through the free-floating microorganisms, gathering into biological flocs that settle out of the wastewater (Bull, Steritt, & Lester, 1984). The settled flocs then retain the

Development in wastewater treatment plant design 313 Table 13.1: Major differences in aerobic and anaerobic treatment systems. Parameter Process principle

Applications

Reaction kinetic Net sludge yield Post treatment Foot-print Capital investment Example technologies

Aerobic treatment

Anaerobic treatment

1. Microbial reactions take place in the presence of oxygen 2. Reactions products are carbon dioxide, water and excess biomass Wastewater with low to medium organic impurities and for wastewater that are difficult to biodegrade, for example, municipal sewage, refinery wastewater etc. Relatively fast Relatively high

Microbial reactions take place in the absence of oxygen Reactions products are carbon dioxide, methane and excess biomass Wastewater with medium to high organic impurities and easily biodegradable wastewater, for example, food and beverage wastewater rich in starch/sugar/alcohol

Typically direct discharge or filtration/ disinfection Relatively large Relatively high

Relatively slow Relatively low (generally one fifth to one tenth of aerobic processes) Invariably followed by aerobic treatment Relatively small and compact Relatively low with pay back

Continuously stirred tank reactor/digester, Activated sludge, for example, Extended upflow anaerobic sludge blanket (USAB), Aeration, Oxidation Ditch, membrane Ultra High Rate Fluidized Bed reactors. bioreactor (MBR), Fixed Film Processes, For example, EGSBTM, ICTM etc. for example, Trickling Filter/Biotower, biological aerated filter (BAF), moving bed bioreactor (MBBR) of hybrid processes, for example, integrated fixed film activated sludge (IFAS)

microbes, which later can be recycled for further treatment. In the current section, we describe activated sludge in detail. a. Activated sludge A high biomass concentration and an increased process kinetic rates support the removal of pollutants that is much faster, highly efficient and is carried out in an integrated way with increasing volumetric capacity. The effluent requirement is based on single and multi-activated sludge system where aerobic system enables significant biological removal of nutrients like nitrogen and phosphorus. It can be further classified as according to flow characteristic into two groups: (a) with continuous flow and (b) with cyclic operation (sequencing batch reactor, SBR and its modifications) (Randall, 1998). i. Continuous flow multistage biological reactors with activated sludge The development of the multi-stage reactor was inspired by J. Barnard’s patent in 1970 which was later constructed and followed in many developed nations. For pre-dinitrification of the return sludge, an additional anoxic chamber is a modification tagged along with the three-stage systems developed by Barnard “Bardenpho”. The process of surplus phosphorus accumulation by

314 Chapter 13 microbes are carried out in two stages in a sequence of anaerobic and aerobic conditions. In order to intensify biological denitrification and phosphorus removal further Bardenpho technology is growing worldwide in the direction increasing the biodegradable carbon concentration in influent wastewater (Cooper & Downing, 1998). ii. Cyclic multistage reactors with activated sludge (SBR) In case of cyclic reactors, the water inflow and outflow and sometimes only outflow is intermittent with the use of usually of one or more reactors. Post 1992, the development of new generation of cyclic reactors was done which were capable of integrated nutrient removal at smaller plants. This system treats wastewater in three or more activated sludge basins. Out of all the basins, one of them serves as a settling tank during a few hours. There is presence of no separate clarifier. In every few hours, one of the basins serves as a settling tank and flow directions are changed. 2. Attached growth system On the contrary, attached growth systems are used as a medium to attach and grow microbes. The biofilters are most common type example of attach growth system. These consist of a fixed base of gravel, plastic, ceramic etc. through which the sewage flows and creates a biofilm aiding microbes to thrive. Biofilters commonly are of two types (1) Trickling filters, where wastewater flows from the top to the bottom through porous biofilter filling. Such biofilters at low loads can remove not only biological oxygen demand (BOD) but also sustain nitrification and (2) Contact beds, where the filling in form of packets or loose plastic profiles is submerged in wastewater. Contact beds may be aerated, and then the removal efficiency is similar to the one obtained for trickling filters (removal of BOD and possible nitrification), or non-aerated, and then they can sustain denitrification (Boon, Hemfrey, Boon, & Brown, 1997; Chaudhary, Vigneswaran, Ngo, Shim, & Moon, 2003). 13.1.1.2.2 Anaerobic treatment

The wastewater treatment using anaerobic metabolism primarily involves several group microbes (Cha & Noike, 1997; Harper & Pohland, 1997). Generally, the complex wastes produced are stabilized via three mechanisms, namely hydrolysis, acid fermentation and methanogenesis (Ghosh, Misra, Dutta, & Choudhury, 1985; Tchobanoglous & Burton, 1990). Anaerobic treatment is more flexible and simple in terms of scale and place in which it is being implemented. Biobulk is a conventional anaerobic contact process with recirculation of sludge that is applicable for waste streams with high chemical oxygen demand (COD)/BOD concentrations and fats, grease and oils with a concentration of 150 mg/L and higher. This process is primarily applied in ice cream plants and food processing facilities with discharge effluents high in biodegradable fats and oils (Mes, Stams, Reith, & Zeeman, 2003).

Development in wastewater treatment plant design 315 13.1.1.3 Tertiary treatment The last and final part of the treatment involves tertiary treatment or also known as advanced treatment. This is achieved by unit operations and chemical unit processes that further remove BOD, nutrients, pathogens, toxic substances and parasites (Abdel-Raouf, Al-Homaidan, & Ibraheem, 2012). Chlorine has been the most common mean of disinfection for reducing the pathogens content in the wastewater. This process has been evidenced to be highly cost-effective; however, a higher concentration of chlorine has a harmful impact towards the aquatic life and wildlife.

13.1.2 Recent advances achieved in wastewater treatment plant This section describes the recent developments in WWTP and have been implemented to improve quality of the treatment, reduce environmental pollutions and enhance the quality of wastewater from both effluent and residual discharges from treatment plant. Hereby, we discuss the developments in details for each treatment section [primary, secondary (aerobic and anaerobic), and tertiary]. 13.1.2.1 Primary treatment Screens are seldom used in the present-day scenario, instead mechanical chain scrapers and extension arm scrapers are trend these days. For continuous and intermittent scraping, stepped and drum screens have been constructed in newer days. Combination of scraping and transport of screenings with dewatering and pressing are the best methods used. Optionally, flushing with water and then transport to enclosed containers are also done. Furthermore, primarily stainless steels are used for building screens. There exists a functional operation between the screen and the screening transport, which controls the level of wastewater sensors before the time or screen switches (Kurbiel, 1998). 13.1.2.2 Secondary treatment 13.1.2.2.1 Aerobic secondary treatment

As described in the previous sections; aerobic secondary growth system typically is of two types (1) Attached growth system and (2) Suspended growth system. We shall now discuss the advancements achieved in both of them individually. 1. Attached growth system In many sewage, plastic media are submerged within the aeration tank so that bacteria may attach and grow faster, these days attached growth is considered modern development. Submerged Aerated Filter, Submerged Aerated Fixed Film, Rotating Biological Contact Filter are some variants of attached growth process. Though it helps in the growth of bacteria, the major drawback of such system is excess growth of

316 Chapter 13 bacteria on the media surface, which requires regular media cleaning to avoid the airflow obstruction within the aeration tank. These twin problems were solved by the modern process called “moving bed bioreactor (MBBR)”. A circular plastic media of usually 22 mm diameter is introduced into the aeration tank and it provides large surface area for bacterial growth (Barwal & Chaudhary, 2014; Chaudhary et al., 2003; Randall, 1998). 2. Suspended growth system: A tabular representation has been given in Table 13.2. a. Cyclic activated sludge system (CASS) CASS as the name suggests is one of the most popular SBR processes employed to treat municipal wastewater and wastewater from a variety of industries including refineries and petrochemical plants.

Table 13.2: Comparisons of advanced attached biological treatment systems.

Parameter

Convention ASP

Treated Effluent quality

Meets specified discharge standards with additional filtration step Average

Ability to adjust to variable hydraulic and pollutant loading Pretreatment requirement

Ability to cope with ingress of oil Secondary clarifier requirement Complexity to operate and control Reliability and proven-ness of Technology Capital cost Operating cost Space requirement

Cyclic activated sludge system (CASS)

Integrated fixed film activated sludge (IFAS)

Membrane bioreactor (MBR)

Exceeds specified Exceeds specified Exceeds specified discharge standards without discharge discharge additional filtration step. standards without standards with Very good for recycle additional additional provided TDS level permits filtration step filtration step Very good Very good Very good

Fine screening for suspended impurities like hair and almost complete oil and grease removal

Suspended impurities, for example, oil and grease and TSS removal Average

Suspended impurities, for example, oil and grease and TSS removal Good

Suspended impurities, for example, oil and grease and TSS removal Average

Needed

Aeration Basin acts as clarifier Operator friendly

Needed Operator friendly

Average

Very good

Very good

Limited references in industrial applications

Low Low High

Low Low Low

High High Average

Very high Very high Low

Simple, but not operator friendly

TSS, Total suspended solids; TDS, total dissolved solids.

Poor and detrimental to membrane Clarifier is replaced by membrane filtration Requires skilled operators

Development in wastewater treatment plant design 317 b. Integrated fixed film activated sludge (IFAS) system There have been many industrial installations where two stage biological treatment comprising stone or plastic media trickling filter that is also known as biotower is followed by activated sludge process aeration tank, which is again followed by a secondary clarifier. There has been modification of the above configuration; whereas in newer industrial wastewater treatment system is fluidized media bioreactor called MBBR. This is in lieu of biotower followed by activated sludge process. This hybrid process of fluidized media and activated sludge process taking place in a single aeration tank is known as IFAS process. c. Membrane bioreactor (MBR) MBR is the latest technology for biological degradation of soluble organic impurities (Fig. 13.1). This technology has been in widespread usage for the treatment of domestic sewage. However, there has been limitation in case of industrial wastewater treatment. This system closely resembles the conventional activated sludge process. In both the processes, they have mixed liquor solids in suspension in an aeration tank. The only difference lies in the method of separation of bio-solids. In the MBR process, the bio-solids are separated by means of a polymeric membrane based on microfiltration or ultrafiltration unit, as against the gravity settling process in the secondary clarifier in conventional activated sludge process. A diagrammatic representation of aerobic wastewater treatment with MBR system is given in Fig. 13.1. 13.1.2.2.2 Anaerobic secondary treatment

Over the years anaerobic plants designs have advanced tremendously and now we shall have a detailed insight. One of the major applications of the high-rate anaerobic treatment is to un-couple the hydraulic retention time (HRT) and the solid retention time in the

Figure 13.1 Representation of aerobic secondary treatment process with membrane bioreactor (MBR).

318 Chapter 13

Figure 13.2 Representation of (A) up-flow anaerobic sludge bed (UASB) and (B) anaerobic fixed film reactor (AFFR).

reactor system. Short HRTs are more convenient for achieving high system loading rates at the same time with maintaining positive biomass. Therefore designing a reactor that can retain biomass within the reactor has been evolved hugely over the past two decades. The anaerobic treatment of wastewater depends upon characteristics of the wastewater, its treatment objectives as translated into desired effluent quality (Ozgun, 2013): Refer Fig. 13.2 for better a better understanding. Types of anaerobic treatment technology 1. Up-flow anaerobic sludge bed (UASB). 2. Expanded granular sludge bed (EGSB). 3. Anaerobic fixed film reactor (AFFR). 1. The up-flow anaerobic sludge bed (UASB) The concept of such a plant design was developed in late 1960s at Wageningen University in The Netherlands. The Dutch beet sugar firm, Centrale Suiker Maatschappij, developed the basic technology for wastewater treatment. The design and engineering of the settler that was simple yet efficient to effectively degasify the biomass and ensuring its retention in the reactor vessel caught a many eyes and was a key to commercialization step. The entry of the wastewaters is through the bottom of the reactor vessel through an inlet distribution system that passes upward through a dense anaerobic sludge bed. Soluble COD rich in methane is then converted to biomass and upward circulation of water, establishing well-settleable sludge. The specially constructed settler sections so that sludge particles not attached to the gas-bubbles would sink to the bottom establishing a return downward circulation provides an effective degasification. A continuous convection is created by the upward flow of

Development in wastewater treatment plant design 319 Spent wash from still house

Plate heat exchanger

Inlet tank Yeast settling tank

Inlet pit Under-ground tunnel

Lime tank

UASB digester

Buffer tank

Foam trap

Air Biogas holder

Lamella

secondary effluent treatment (for further treatment if required)

Figure 13.3 Representation of a basic anaerobic treatment plant design.

gas-containing sludge through the blanket combined with return downward flow of degassed sludge. Within the reactor, it ensures an effective contact of sludge and wastewater without aid of any energy consuming mechanical or hydraulic agitation. In relation to soluble organic solids passing through the sludge bed this unique design of the reactor allows a highly active biomass which hold responsible for very high loading rate (short HRT), which can readily be achieved (Chan, Chong, Law, & Hassell, 2009; Mes et al., 2003). A representational view is given in Fig. 13.3A. 2. The expanded granular sludge bed (EGSB) This design methodology extensively incorporates sludge granulation concept of UASB’s. The elimination of carrier material as a mechanism for biomass retention within the reactor is the main improvement of the EGSB system, trademarked ‘Biobed’ (by the company Biothane), compared to other types of anaerobic fluidized or expanded bed technologies. Applications for Biobed include wastewater from breweries, chemical plants, fermentation industries and pharmaceutical industries. Therefore this system is perceived either as a modified conventional fluidized base or as a high-rate UASB. This system is highly space efficient that requires just a small footprint size than a USAB system and widely operated at high COD loading (Driessen & Yspeert, 1999; Ozgun, 2013).

320 Chapter 13 3. Anaerobic fixed film reactors (AFFR) This plant design has been applied successfully in the treatment of high strength effluent treatment. It contains a mixed population of bacteria immobilized on the surfaces of support medium (Chua & Fung, 1996). This hybrid system was development to overcome the problems faced in UASB and Anaerobic Filter (AF) systems. Presence of dead zones and channeling in the lower part of the filter is very commonly noticed in an AF reactor. Likewise, when the wastewater contains a large amount of suspended solids, sludge washouts may occur due to problem in UASB systems. Therefore the hybrid built combines both fixed bed system (at the top of the reactor) with the UASB system. The filter zone in the hybrid reactor plays a key role for biomass retention as biological activity contributing to COD reduction (Chan et al., 2009; Elmitwalli, 2000; Ozgun, 2013). A representational view is given in Fig. 13.3B.

13.2 Tertiary treatment Ultraviolet (UV) light radiation recently has been a cost-effective and non-health hazardous alternative to chlorination/de-chlorination process. An economic survey has also reported that UV light radiation would be 12% less costly in terms of constructing and operating compared to chlorination/de-chlorination process. An added advantage of this is also that it does not have to deal with handling huge quantities of toxic materials like chlorine and sulfur dioxide which further reduces health risk to plant employees and the general public (De la Cruz, 2013).

13.3 Conclusion The chapter summarizes on the various aspects of wastewater treatment mainly via aerobic and anaerobic digestion. Globally industries are impregnated with the challenge imposed on them for treating the waste generated as a result of the multiple processes. Literature reports suggest that both aerobic and anaerobic treatment of wastewater are essential. A combinatorial strategy involving both the processes shall lead towards a more efficient treatment. Contextual to the requirements and possible solutions to the problems faced, the present chapter highlights the current methods and certain advancements.

References Abdel-Raouf, N., Al-Homaidan, A. A., & Ibraheem, I. B. M. (2012). Saudi Journal of Biological Sciences, 19, 257 275. Barwal, A., & Chaudhary, R. (2014). Reviews in Environmental Science and Biotechnology, 13, 285 299. Batstone, D. (2002). Water Science Technology, 45, 65 73. Boon, A. G., Hemfrey, J., Boon, K., & Brown, M. (1997). Water and Environment Journal, 11, 393 412. Bull, M. A., Steritt, R. M., & Lester, J. N. (1984). Chemical Engineering Research and Design, 62, 203 213.

Development in wastewater treatment plant design 321 Cha, G. C., & Noike, T. (1997). Water Science and Technology, 36, 247 253. Chan, Y. J., Chong, M. F., Law, C. L., & Hassell, D. G. (2009). Chemical Engineering Journal, 155, 1 18. Chaudhary, D. S., Vigneswaran, S., Ngo, H.-H., Shim, W. G., & Moon, H. (2003). Korean Journal of Chemical Engineering, 20, 1054. Chua, H., & Fung, J. P. C. (1996). Water Science and Technology, 33, 1 6. Cooper, P. F., & Downing, A. L. (1998). Water and Environment Journal, 12, 303 313. De la Cruz, N. (2013). Water Research, 47, 5836 5845. Driessen, W., & Yspeert, P. (1999). Water Science and Technology, 40, 221 228. Elmitwalli, T. A. (2000). Anaerobic treatment of domestic sewage at low temperature. Dissertation, Environmental Section of the Department of Agrotechnology and Food Sciences, Wageningen University, Wageningen, p. 120. Ghosh, A., Misra, S., Dutta, A. K., & Choudhury, A. (1985). Phytochemistry, 24, 1725 1727. Harper, S. R., & Pohland, F. G. (1997). Water Science and Technology, 36, 33 39. Koivunen, J., & Heinonen-Tanski, H. (2005). Water Research, 39, 4445 4453. Kurbiel, J. (1998). Development of wastewater treatment in Poland from perspective of practical implementation. In International symposium on the assessment disposal and treatment of rural wastes: The protection of freshwater resources. rivers, lakes and groundwater (pp. 34 35). Mes, T. Z. D., Stams, A. J. M., Reith, J. H., & Zeeman, G. (2003). Chapter 4. Methane production by anaerobic digestion of wastewater and solid wastes. In J. H. Reith, R. H. Wijffels, & H. Barten (Eds.), Biomethane and biohydrogen. Status and perspectives of biological methane and hydrogen production (pp. 58 94). Switzerland: Sustainable Sanitation and Water Management. Ozgun, H. (2013). Separation and Purification Technology, 118, 89 104. Randall, C. W. (1998). Water and Environment Journal, 12, 375 383. Tchobanoglous, G., & Burton, F. (1990). In M. Eddy (Ed.), Wastewater engineering: Treatment, disposal and reuse. Water resources and environmental engineering (Vol. 3). New York: McGraw-Hill. Zhang, T. C., Surampalli, R. Y., Tyagi, R. D., & Benergi, S. K. (2017). Chapter 14: Biological treatment of hazardous wastes. In J. W. C. Wong, R. D. Tyagi, & A. Pandey (Eds.), Current developments in biotechnology and bioengineering: Solid waste management (pp. 311 340). Elsevier.

CHAPTER 14

Engineering biomaterials for the bioremediation: advances in nanotechnological approaches for heavy metals removal from natural resources Magapu Solomon Sudhakar1, Aakriti Aggarwal2 and Mahesh Kumar Sah2 1

Applied Biotechnology Department, Sur College of Applied Sciences, Ministry of Higher Education, Sur, Sultanate of Oman, 2Department of Biotechnology, Dr. B. R. Ambedkar National Institute of Technology, Jalandhar, India

14.1 Introduction The recent years have witnessed an unprecedented growth in industrialization and urbanization across the globe. Although this has given comfort and solace to the growing population but has also brought substantial damage to the natural resources like water, soil, and air in the environment which has caused considerable loss of existing flora and fauna. Past decades have noticed tremendous rise in the activities of mining, electroplating, smelting, fertilizer, pesticides, tanneries, paper and electronic industries etc. and these have accounted for the release of large amounts of heavy metals and petroleum hydrocarbons into the natural ecosystem, which has been reported to disrupt and alter the physiological functions in biological systems. The continuous leaching of heavy metals by metal corrosion followed by atmospheric deposition and sediment resuspension to soil and ground water are the key factors causing ruthless contamination of the environment. In the predominant list of heavy metal pollutants, the most hazardous heavy metals such as arsenic (As), cadmium (Cd), lead (Pb), copper (Cu), chromium (Cr), nickel (Ni), zinc (Zn), aluminum (Al) and manganese (Mn) are known to be the major peril to the environment. These heavy metals enter the environment by both the natural and anthropogenic ways causing a relentless threat to human health as they can be amassed in biological food web. It is reported that heavy metals can cause damage to the kidneys, mental and central nervous functions, lungs, and other organs (Table 14.1). Heavy metals are considered highly toxic (Afroze & Sen, 2018) Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00014-6 © 2020 Elsevier Inc. All rights reserved.

323

324 Chapter 14 Table 14.1: Various kinds of heavy metals and their effect on human health with permissible limits based on U.S. Environmental Protection Agency (EPA) and the World Health Organisation reports (Kumar et al., 2017). S. no. Heavy metal 1 2 3 4 5 6 7

Arsenic (As)

Effect on human organs/health

Skin, lungs, brain, kidneys, liver, metabolic, cardiovascular, immunological, and endocrine Cadmium (Cd) Bones, liver, kidneys, lungs, testes, brain, cardiovascular, and immunological Chromium Skin, lungs, kidneys, liver, brain, pancreas, testes, gastrointestinal, (Cr) and reproductive Copper (Cu) Liver, brain, kidneys, cornea, lungs, gastrointestinal, immunological, and hematological Lead (Pb) Bones, liver, kidneys, lungs, spleen, brain, cardiovascular, immunological, hematological, and reproductive Mercury (Hg) Brain, lungs, kidneys, liver, cardiovascular, immunological, endocrine, and reproductive Manganese Central nervous system, lungs, and inhalation (Mn)

Permissible limits (mg/L) 0.02 0.06 0.05 0.1 0.1 0.01 0.26

and most of them have been reported to be carcinogenic (Cocaˆr¸ta˘ , Neam¸tu, & Re¸setar Deac, 2016). Similar hostile health impacts have been reported for other heavy metals and hence, U.S. Environmental Protection Agency and the World Health Organization have set a maximum tolerable limit for every heavy metal in various systems (Table 14.1) (Kumar, Kim, Bansal, Lazarides, & Kumar, 2017). Many technological methods are developed to remove heavy metals from various contaminated environmental media such as chemical ´ lvarez, 2006), ion exchange precipitation (Gonzalez-Munoz, Jesus, Rodrı´guez, Luque, & A (Verbych, Hilal, Sorokin, & Leaper, 2005), adsorption (Namasivayam, Sangeetha, & Gunasekaran, 2007), membrane filtration (Sudilovskiy, Kagramanov, Trushin, & Kolesnikov, 2007), electrochemical treatment (Tran, Leu, Chiu, & Lin, 2017), and so on. In fact, different technologies are blended for effective heavy metals removal (Blo¨cher et al., 2003; Mavrov, Erwe, Blo¨cher, & Chmiel, 2003). Among these, the most significantly used technique is the adsorption due to its low cost and simple operation. In recent years, porous metal-organic framework (MOF) materials have been reported to exhibit extreme dominance in removing hazardous substances from the environment due to its properties such as large tunable porosity, pore functionality, and various pore architecture characteristics. (Khan, Abedin, Hasan, & Jhung, 2013). Over the years, these nanomaterials have delivered a promising approach for removing heavy metals from natural resources (Buzea & Pacheco, 2017). Nanoscale sized materials is facilitated with some unique properties, such as a surface effect, small size effect, quantum effect, and macro quantum tunnel effect (Yinghua, Chengzhang, & Zirong, 2006). These properties contribute magnificently towards adsorption capacity and reactivity, both of which are favorable for

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the removal of heavy metal ions. For instance, a magnetic nanocomposite based on aluminum MOF (MIL-53) synthesized by Ricco et al. (2015) exhibited excellent removal capacity towards Pb(II). This paper reviews various nanomaterials and the engaged mechanisms for bioremediation. As these are commonly employable for heavy metals removal from water resources, the removal of same from soil with different strategies is discussed separately.

14.2 Bioremediation Bioremediation is utilizing biological agents such as bacteria, fungi, and other microorganisms and/or their products to cope with environmental issues (El-Kassas & Mohamed, 2014; Galdames et al., 2017; Sode et al., 2013). These agents clean up toxic, hazardous heavy metals from the environment by converting harmful materials get converted into harmless by-products along with water (Ekperusi & Aigbodion, 2015; Golodyaev, Kostenkov, & Oznobikhin, 2009). The two strategies to employ bioremediation, firstly either by enhancing the population of scavengers (e.g., “pollution eating” microbes) at the contaminated site and secondly by putting these agents and contaminated site exogenously (Huertas et al., 2008; Shankar et al., 2014). Thus, the bioremediation can be classified into two broad categories, in situ and ex situ. The in situ process treats the toxic materials at the site of contamination. This makes the process less expensive, reduced release of contaminants into the environment, and large area can be remediated. The ex situ method treats the waste after excavation. This makes the approach faster, easier to control and usable for treating wide range of contaminants in comparison to the in situ (Huertas et al., 2008).

14.3 Nanotechnology and bioremediation With the current progress of nanotechnology, the amalgamation of nanoparticles and biological entities effectively enhance measurement accuracy, bioremediation efficiency, and broader biochemical functions in environmental research. Nanoparticles based bioremediation has low risk of genetic leakage in the environment and can deliver supplementary functions and characters to the biochemical process. The concept of nanoremediation is an emerging field since 2009 and has been commercially applied in 44 clean-up sites around the world (Gholami-Shabani, Gholami-Shabani, Shams-Ghahfarokhi, & Razzaghi-Abyaneh, 2018). The nanoparticles are able to remediate many contaminants without any disadvantages and limitations because of their high specificity. In addition, the nanomaterials can be manufactured by “green” approaches (Mishra et al., 2014). Green technology is the broadly acknowledged technique for bioremediation for their non-toxic effect, clean and eco-friendly style. While there are numerous approaches for the production of nanoparticles like sol-gel method, chemical synthesis but biologically synthesis of

326 Chapter 14 nanoparticles is utmost suitable and eco-friendly (Mishra et al., 2014). Nanoparticles manufactured by green nanotechnology approach use living organisms, plants and microbes. Microbes are mainly used for the commercial production of nanoparticles due to their high tolerance, reproduction power and rapid decontamination. The unification of nanoparticles with diversified shape and size affecting its properties is a significant aspect of nano-biotechnology. The particles generated biologically have higher catalytic reactivity and greater specific surface area (Riddin, Gericke, & Whiteley, 2010). Biosynthesized nanoparticles do not aggregate due to the presence of capping agent secreted by specific microorganism. Extracellular biosynthesis of nanoparticles have drawn attention because of low cost requirement and no downstream processing (Mishra et al., 2014). The secondary metabolites and extracellular components existing in cell free extract bring out the redox reaction for biosynthesis of particles after addition of precursor molecule. By varying biological and physical parameters, the configuration of particles can also be varied as depicted (Fig. 14.1).

14.3.1 Nanomaterials used for removing pollutants 14.3.1.1 Carbon-based nanomaterials The carbon-based nanomaterials were initially used by the electronics industry due to their unique optical, electronic, vibrational, mechanical and thermal properties (Popov, 2004). In the recent times, a plethora of reports have been published on the applications of carbon nanotubes (CNTs) for the removal of heavy metals from the wastewater (Gupta et al., 2016). Primarily, the CNTs are categorized into two types, the single-walled carbon nanotubes and the multi-walled carbon nanotubes (Martel, Schmidt, Shea, Hertel, & Avouris, 1998). Carbon nanotubes display high superiority in treating heavy metal wastewater, mainly due to their large specific surface area, high adsorption capacity, and fast adsorption kinetics (Lu et al., 2016). It is reported to have excellent adsorption effects towards Mn(VII), TI(I), Cu(II), Pb(II), and Cr(VI) (Kabbashi et al., 2009; Pu et al., 2013; Tang et al., 2012; Yadav & Srivastava, 2017). The possible adsorption active sites of carbon nanotubes are mainly comprised of the outside surface, interstitial channels, internal sites, and external groove sites. To improve the adsorption capacity of CNTs towards heavy metals, functional groups, such as COOH, NH, OH, etc. are generally introduced onto the surface of CNTs by means of chemical modification, heat treatment or endohedral filling (Kumar, Khan, & Haq, 2014). Though using CNTs to remove heavy metals from wastewater has many gains, it still has a few disadvantages. Firstly, the high cost of CNTs impede their commercial use. Moreover, it is usually difficult to separate the CNTs from wastewater after the adsorption and this increases the treatment costs and the risk of secondary pollution. This necessitates the synthesis of CNTs by green methods. An ecofriendly approach by Hakim et al. describes the production of CNTs from the graphite flakes. These graphite flakes were obtained from coconut shell wastes, and then used in

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Figure 14.1 The schematic representation of different methods for nanoparticles development in different forms and further modification to equipped it for efficient mechanism for heavy metals removal.

one-step water assisted synthesis of CNTs. The nanotubes thus formed, were used for the remediation of lead (Hakim, Zaim, Yulizar, Nurcahyo, & Surya, 2018). Li et al. developed CNT/calcium alginate complex (CNT/CA). The CNTs were immobilized onto CA and further used for studying Cu adsorption. The results demonstrated high Cu removal efficiency, even at a lower pH (Li et al., 2012). Furthermore, graphene oxide (GO) and reduced graphene oxide (r-GO) are two kinds of graphene-based nanomaterials which can be used to remove heavy metals from waste water.

328 Chapter 14

Figure 14.2 Conversion of graphene to its reduced graphene oxide (r-GO) form for decorating with Au nanoparticles (Au NPs) capable of heavy metals adsorption. The side extension in the carbon shell indicates the presence of oxygen containing functional groups (Gnanaprakasam et al., 2016).

GO is the oxidation product of graphene and it contains miscellaneous oxygen-containing functional groups, such as hydroxyl, carboxyl, epoxide, and carbonyl functional groups, which craft it; making it a promising material to eliminate heavy metals (Fig. 14.2). A study, highlighting the use of r-GO supported Au nanoparticles (rGO-Au NPs) modified electrode was conducted. The reducing agent used for synthesizing these nanoparticles was Abelmoschus esculentus vegetable extract. This bioremediation process simultaneously detected Cd(II), Pb(II), Cu(II), and Hg(II) ions. The bacterium species of Pseudomonas aeruginosa, Rizobium gallium, Staphylococus aureus, and Bacillus subtilis were further used as scavengers for the heavy metal ions owing to their tremendous adsorption property (Gnanaprakasam, Jeena, Premnath, & Selvaraju, 2016). 14.3.1.2 Hydroxyapatite nanomaterials Hydroxyapatite, Ca10(PO4)6(OH)2, has high affinity for divalent heavy metals ions. The high calcium substances present in the egg shells aids the development of hydroxyapatite nanoparticles. Banana is one of the most loved sustenance worldwide and the husk is dumped as rubbish. This is made out of biopolymers, for example, cellulose, hemicelluloses, gelatin, lignin and proteins and could be effectively utilized for the union of nanoparticles. Hydroxyapatite nanoparticles have been combined utilizing banana husk. Gelatin present in the husk assumes a noteworthy job in the surface adjustment of the nanoparticles. Hydroxyapatite is a unique inorganic compound and has been used to remove various heavy metals such as nickel (Zamani, Salahi, & Mobasherpour, 2013), lead (Mavropoulos et al., 2002; Zamani et al., 2013), cadmium (Lusvardi, Malavasi, Menabue, & Saladini, 2002), cooper (Corami, Mignardi, & Ferrini, 2007), zinc (Skwarek, 2014), and arsenic (Nakahira, Okajima, Honma, Yoshioka, & Tanaka, 2006).

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14.3.1.3 Metallic nanoparticles Nanoparticles have two key properties that make them especially alluring as sorbents. On a mass premise, they have a lot of bigger surface territories than mass particles. It can likewise be functionalized with different synthetic gatherings to expand their proclivity towards target mixes. Magnetic nanoparticles of Fe3O4 functionalized with ammonium oleate coated on the surface of Pseudomonas delafieldii. The microbial when subjected to an external magnetic field, desulfurize organic sulfur from fossil fuel (dibenzothiophene) (Shan, Xing, Zhang, & Liu, 2005). Table 14.2 highlights the different types of nanoparticles and application theirof for effective removal of heavy metals. 14.3.1.4 Biogenic uraninite nanoparticles Uraninite Nanoparticles were first formed by reducing U(VI), with the help of iron reducing bacterium Geobacter metallireducens strain GS15. Nanoparticles with size ,3 nm were Table 14.2: Nanoparticles synthesized with biological agents for heavy metals removal from natural resources. Nanoparticles Epoxypolysaccharide NPs, EPS-605

Remediation of pollutants and related strategy

Pb(II), Cu(II), and Cd(II) by adsorption. The adsorption ability is affected by pH, temperature, and initial adsorbate concentration Covellite nanocrystals, Synthesized using biologically generated CuS sulfide (by sulfate reducing bacteria, SRB) for waste water treatment Magnetic, Fe3O4 NPs Algae extracts (Padina pavonica, Sargassum acinarium) reduce ferric chloride, useful for Pb remediation Iron, Fe NPs Polyphenols in eucalyptus leaves, tea leaves extract and other natural polyphenols rich sources utilized for the formation of Fe NP, which are further used for groundwater remediation Pt-Au-PDA/rGO hybrid Pt-Au dendrimer on the surface of polydopamine wrapped reduced graphene oxide for the remediation of 4-nitrophenol Glycerophosphodiesterase (GpdQ) Ser127Ala mutantenzyme based mutant of Enterobacter PAMAM dendrimeraerogenes, immobilized on PAMAM modified magnetite dendrimer-modified magnetite NPs nanoparticles for remediating organophosphate pesticide

References Li et al. (2017)

da Costa et al. (2013)

El-Kassas, Mohamed, and Gharib (2016) Kuppusamy et al., (2017), Truskewycz, Shukla, and Ball (2016), and Wang, Lin, Chen, Megharaj, and Naidu (2014) Ye et al. (2016)

Daumann, James, Ollis, Schenk, and Gahan (2014)

NPs, Nanoparticles; r-GO, reduced graphene oxide; PAMAM, polyamidoamine.

330 Chapter 14 produced using pure cultures of Desulfosporosinus spp. In situ U(VI) reduction, using Thiobacillus denitrificans and Acidithiobacillus ferrooxidans is being investigated in fieldscale tests at a number of contaminated U.S. Department of Energy nuclear legacy sites and has produced a lot of positive results for the bioremediation of subsurface U(VI) contamination (Bargar, 2006). 14.3.1.5 Dendrimers Dendritic polymeric nanoparticles are delicate and used for removing natural and inorganic solutes, microbes, and harmful metal particles from water. They are highly branched polymer compounds, have different functional groups where they can react with different functional entities. Chitosan-polypropylene imine dendrimer has been known to remove two textile reactive dyes, Reactive Black 5 (RB5) and Reactive Red 198 (RR198). This biopolymer shows strong adsorption for these dyes (Sadeghi-Kiakhani, Arami, & Gharanjig, 2013). Various studies on Polyamidoamine starbust dendrimers show high concentration of nitrogen ligands on both the internal and external surfaces of dendrimer molecules. These molecules thus serve as uranyl ion (UO2)21 “sponges” in environmental cleanups (Ottaviani et al., 2000). 14.3.1.6 Polymeric nanocomposites Natural polymers like chitosan, cyclodextrin, and artificial polymers like polystyrene, poly (1-vinylimidazole etc. (Bhaumik, Maity, Srinivasu, & Onyango, 2011; Chen et al., 2015; Rajesh Kumar, Singh, & Singh, 2016; Reddy & Lee, 2013; Takafuji, Ide, Ihara, & Xu, 2004; Wu, Tseng, & Juang, 2010; Wu et al., 2013) were used with magnetic composites for heavy metal ion removal. Lately, many studies show the use of magnetic nanoparticles in the polymer matrix for very effective remediation process. Coating the magnetic nanoparticles with polymers leads to decreased toxicity, inhibition of aggregation, and increased storage life. Fe3O4, Fe2O3, CoFe2O4, NiFe2O4, and ZnFe2O4 are some of the magnetic particles used with polymers (Cheng et al., 2005; Covaliu et al., 2011; Lin et al., 2012; Luo, Tian, Yang, Zhang, & Yan, 2013; Qin, Li, Chen, & Jiang, 2009; Singh, Srivastava, Dutta, & Dutta, 2011; Singh, Srivastava, Kalita, & Malhotra, 2012). Various studies on polymeric nanocomposites reveal two kinds of structures—magnetic corepolymer shell and homogeneous magnetic cores dispersed in polymer matrix (Reddy & Lee, 2013). In one of the studies, amine functionalized magnetic nanoparticles were modified by chitosan, and the resulting complex was used to eliminate metal ions from water. The results indicated a reversible interaction between heavy metal ions and chitosan. Desorption of metal ions from chitosan was performed in weak acidic conditions, assisted by ultrasound radiation (Liu, Hu, Fang, Zhang, & Zhang, 2008). In another instance, carboxymethyl-β-cyclodextrin polymer modified Fe3O4 (CDpoly-MNPs) nanoparticles were

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fabricated for the selective removal of Cd(II), Pb(II), and Ni(II) ions from solutions. Qi and Xu (2004) have assessed the sorption of Pb(II) onto chitosan nanoparticles (40 100 nm) made by ionic gelation of chitosan and tripolyphosphate (Qi & Xu, 2004). 14.3.1.7 Nanozymes Nanozymes, also known as next generation artificial enzymes, are nanoparticle based enzyme mimics. Under optimum physiological conditions, they can catalyze the biochemical reactions in a way, similar to that of natural enzymes (Gao, Fan, & Yan, 2017). Low cost and high stability are the important features, in addition to their mimicking action. Nanozymes usually lack an active site, only a specific substrate can bind and undergo a chemical reaction. They are found to be a powerful, cost effective and simple method for detecting and degrading pollutants such as dyes, lignin containing wastes and organic compounds, etc. Table 14.3 discusses a few nanozymes, the enzymes they mimic and pollutants treated by them. 14.3.1.8 Microorganisms Various microorganisms are known to function in the bioremediation process with high efficiency. Environmental studies of the waste water at the copper mine in Amazon Brazilian region, showed extracellular formation of Cu nanoparticles, using the fungal strains of Hypocrea lixii and Trichoderma koningiopsis (Salvadori, Ando, do Nascimento, & Correˆa, 2014; Salvadori, Lepre, Ando, do Nascimento, & Correˆa; 2013). A recent study highlights degradation of para- nitrophenol to amino phenol in a short time using biosynthesized gold nanoparticles by Trichoderma viride as heterogeneous catalyst (Ai and Jiang, 2013; Mishra et al., 2014).

Table 14.3: Nanoparticles mimicking and functioning as enzyme for the bioremediation. Nanozymes (nanocomposite)

Enzyme mimicked

Pollutants remediated

References

CeO2/γ-Fe2O3

Phosphatase

Janoˇs et al. (2015)

Magnetic NPs (Fe3O4-MNPs) CH-Cu nanozyme (Cu with cysteinehistidine dipeptide

Peroxidases

Organophosphorus pesticide- parathion methyl, as well as the chemical warfare agents, Soman and VX Organic pollutants, such as methylene blue, phenol, and rhodamine Phenols

Laccase

Sharma, Dangi, and Shukla (2018) Wang et al. (2019)

332 Chapter 14

14.3.2 Bioremediation of soil Soil being an indispensable natural resource, plays a pivotal role in the economic growth. Therefore soil health becomes a growing global concern as the farmers depend upon agricultural produce. The soil gets contaminated due to the industrial waste, metals from mines and house hold discharges. Remediation of soil had long been a laborious task by excavation which was followed by incineration and then landfilling, but this takes substantial time and produces hazardous gases and by products. The bioremediation process of soil through the approach of “Green Technology”, does not cause damage to the environment and the amount of energy consumed by the chemical processes employed is reasonably curtailed (Kidwai & Mohan, 2005). 14.3.2.1 Phytoremediation Plants play a major role in the remediation process of the soil as the heavy metal contaminants are extracted by the plants. The transport channels of plant cells play a vital role in the transportation and accumulation of heavy metals into the vacuoles of the plant. This natural bioremediation process was further investigated using CeO2 and ZnO nanoparticles. These nanoparticles increase the root and shoot growth in edible plants such as soybean, wheat, corn and alfalfa. These experiments suggested the enhanced efficiency of phytoremediation using nanotechnology (Lo´pez-Moreno et al., 2010). Nano TiO2 has been known to increase the growth of spinach, when administered to the seeds or sprayed onto the leaves, thereby promoting the adsorption of nitrate and accelerating the transformation of inorganic nitrogen into organic form (Gao, Teresa, Zhao, Han, & Qiu, 2009; Jang, Dong, Kim, & Kim, 2001). 14.3.2.2 Microbial bioremediation The inhabiting microbes in the soil have the coherent potential of converting the higher contaminates into by products or nanoparticles of the corresponding salts or ions in the soil. These microbes have a good tolerance to the heavy metal contaminations in the soil. These biosynthesized nanoparticles could be extracted from the microbes and employed in industrial remediation process or could increase the soil and plant growth activity (Mishra et al., 2014). Polycyclic aromatic hydrocarbons (PAH) are the products of the incomplete combustion of fossil fuels which has become a major toxic organic contaminant sneaking into terrestrial and aquatic ecosystems (Gibson, 1984). Enzyme manganese peroxide (MnP) which is synthesized by Anthracophyllum discolor, a white rot fungus from Chile, was immobilized on nanoclay which was obtained from volcanic soil and its capacity to destroy PAHs was studied with the free enzyme.(Acevedo et al., 2010) The three enzymes namely, lignin

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peroxidase, laccase, and MnP which are synthesized from white rot fungi play a significant role in degradation of PAHs (Collins, Kotterman, Field, & Dobson, 1996). 14.3.2.3 Nanomaterials based bioremediation Nanoparticles can likewise be utilized as biocatalysts for reductive dechlorination. De Windt, Aelterman, and Verstraertem (2005) revealed palladium Pd(0) nanoparticles can be stored on the cell divider and inside the cytoplasm of Shewanella oneidensis and charged with H radicals by including various substrates, for example, hydrogen, acetic acid derivation and format as electron givers in bioreductive tests containing Pd(II). At the point when these charged Pd(0) kept in S. oneidensis cells get in contact with chlorinated exacerbates; the H 1 radical on the Pd(0) can chemically respond with PCP bringing about the expulsion of the chlorine atom from the chlorinated mixes (Windt, Aelterman, & Verstraete, 2005). Nanoparticles can be additionally used to immobilize microbial cells that can debase or biorecover explicit synthetic compounds. Magnetic nanoparticles (Fe3O4) were functionalized with ammonium oleate, covered on the outside of Pseudomonas delafieldii. By applying an outside magnetic field to these microbial cells, these magnetic nanoparticle-covered cells were exhibited to desulfurize natural sulfur from petroleum product (i.e., dibenzothiophene) (Shan et al., 2005). Nanoparticles of Euphorbia macroclada is proposed for evacuating and detoxification of substantial metals (particularly Pb, Cd, Cu, and Zn), from dirtied conditions (Mohsenzadeh & Rad, 2011).

14.4 Conclusion This paper reviews different nanomaterials, both customized and nanostructured, and considers their proficiency in expelling lethal overwhelming metals from wastewater effluents. As of now, there exists an extraordinary assortment of sorbents for expelling these metals, but there is no all-inclusive sorbent so far because of the explicitness of conditions and impacts of different factors. Materials based on nanotechnology, and having remarkable physical, chemical and mechanical properties, can turn into a promising option in contrast to conventional sorbents, both autonomously and as a component of half-breed materials for progressively proficient expulsion of substantial metal contaminants from waste water effluents. In this work, a progression of nanomaterials including carbon-based nanomaterials, metallic and non-metallic nanoparticles, nanozymes, polymeric nanocomposites, and dendrimers are examined in detail. The main deterrent to utilizing the previously mentioned nanomaterials as adsorbents is their high cost. Nonetheless, with the expansion in their yield on a mechanical scale with further advancement of nanotechnology, the cost of these materials will diminish. There are still a few bottlenecks that need to be defeated to utilize these nanomaterials in water

334 Chapter 14 treatment. It is generally hard to isolate the nanomaterials from the water quickly and proficiently due to their nanoscale size. The proposition of nanocomposites is by all accounts a promising way to deal with taking care of these issues. Just as the working expenses of nanomaterials ought to be streamlined for the economy, the generation of these nanomaterials should meet the pre-requisites of green science. With the expanding utilization of nanomaterials in waste water treatment, their effects and toxicities towards both the earth and individuals ought to be mulled over. Like the studies on the conduct of nanomaterials on human wellbeing, the compatibility of these materials with the earth should be additionally examined. Furthermore, the remediation capacities of these nanomaterials are for the most part examined in invigorated water with the general basic segments. The reports on their utility in actual wastewater sites requires more studies and experiments. Finally, more efforts are required from our side to diminish the waste, which in turn will itself lead to improved remediation outcomes.

Acknowledgement This work was supported by Third phase of Technical Education Quality Improvement Programme (TEQIP Phase-III), a program by MHRD (Govt. of India). The authors are also thankful to the affiliated Institutes for providing infrastructural backing.

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CHAPTER 15

Algal
bacterial symbiosis and its application in wastewater treatment Inigo Johnson, Sudeeptha Girijan, Binay Kumar Tripathy, Mohammad Abubakar Sithik Ali and Mathava Kumar Environmental and Water Resource Engineering Division, Department of Civil Engineering, IIT Madras, Chennai, India

15.1 Introduction Algae and bacteria are the two most common groups of organisms in freshwater and marine habitats. Algae is a general term to depict organisms that can fix inorganic carbon through photosynthesis and carry out the metabolic processes by obtaining energy from light. They can be prokaryotic (cyanobacteria and diatoms), eukaryotic (microalgae and macroalgae), single, and multicellular. Multicellular algae lack developmental features like a root, shoot, and stem that are generally seen in higher plants. Bacteria are unicellular prokaryotes which are considered to be the first life forms on earth. Bacteria feed on organic carbon sources unlike algae and obtain energy by breaking down complex organics into simple forms. Cyanobacteria are bacteria that have obtained pigments which can trap energy from light and help in photosynthesis. Though they belong to the bacterial family, in this book chapter, the word “bacteria” will be used to refer to bacteria other than cyanobacteria and the word algae will be used to include cyanobacteria along with microalgae and macroalgae. Algae contribute ,1 Gt C of total biomass of earth (B550 Gt C), whereas bacteria contribute to around 70 Gt C (Bar-On, Phillips, & Milo, 2018). The remainder is mostly covered by plants (B450 Gt C), animals, and archaea. The coexistence of a myriad of organisms in a single habitat have paved the way for the organisms to interact with each other with mechanisms that were fine-tuned through evolution. The huge contribution by bacteria after plants signify their availability for interaction with various hosts that occupy the earth. The interaction of bacteria with photosynthetic organisms is well known. The exchange of carbon, nitrogen, sulfur, and other nutrients between photosynthetic algae and bacteria forms the basis of the nutrient cycle in the marine system. Algae supply fixed carbon and oxygen to bacteria, whereas bacteria supply carbon dioxide and nitrogen to

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00015-8 © 2020 Elsevier Inc. All rights reserved.

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342 Chapter 15 algae. The interactions are more complex involving the exchange of bioactive compounds like transcriptional regulators, growth factors, and toxins. This symbiotic relation between bacteria and algae has been exploited for wastewater treatment applications. They have been used for biochemical oxygen demand (BOD) and nutrient removal, disinfection, and in many cases for refractory compounds removal. In addition, they have also been used for carbon sequestration. The major advantage of using algal
bacterial symbiotic systems is the reduced requirement for aeration and a more efficient nutrient removal. The algal biomass produced also acts as a raw material for fuel production and resource recovery, further reducing the economic burden on wastewater treatment. However, the large area requirement, low Hydraulic retention time (HRTs), effluent quality deterioration due to algal biomass flush-out, etc. limit their application.

15.2 The symbiotic process Algal growth is influenced by its microenvironment which is contributed by the nutrients and other microorganisms. The habitat around the algae, along with the interacting microorganisms, is called the phycosphere. Algal
bacterial interactions can either be stimulatory or antagonistic. The term symbiosis was coined by a German botanist, Heinrich Anton de Bary, to describe the interactions between two different organisms that exist in proximity. When the symbiotic interaction is advantageous to both the organisms, it is called mutualism. If only one organism benefits leaving the other organism unaffected, it is called commensalism. If one organism benefits and the other organism is harmed, it is called parasitism. If one species is harmed without benefiting the other organism, it is called amensalism. The type of dependence on a symbiotic relationship can be classified as obligate and facultative. If the interaction is essential for at least one of the interacting species to survive, it is called obligate interaction. If the interaction is beneficial, but it is not compulsorily a necessary interaction, it is called facultative interaction. The interactions of Oscillatoria sp. and bacteria (Flavobacterium, Xanthomonas, and Pseudomonas) isolated from the Oscillatoria filaments were studied. It was observed that all the bacteria from the Oscillatoria microcosm stimulated the growth of Oscillatoria. On the other hand, the bacteria were stimulatory, antagonistic, or had no effects on the growth of other algal species. Certain algae (Chlorella and Euglena) were able to overcome the inhibitory effects of the bacteria which later became stimulatory to the algal growth. During the experiments with heat inactivated bacteria, the stimulatory and antagonistic effects were not observed, leading to the conclusion that the interactions are specific and are caused by agents from viable bacteria (Delucca & McCracken, 1977). The algae were also able to dominate their microenvironment by being able to produce toxins and other molecules leading to selective enrichment of stimulatory prokaryotes around themselves. The complex interactions between algae and bacteria were speculated to be due to the simplification of

Algal
bacterial symbiosis and its application in wastewater treatment 343 nitrogenous compounds, release of growth factors, adsorption of nutrients, and evolution of CO2 by the bacteria in the vicinity of the algae leading to growth enhancement. Oxygen is produced during algal photosynthesis which acts as an electron acceptor to heterotrophic organisms leading to their growth enhancement. This is called photosynthetic oxygenation and is a major reason for the enhanced performance of algal
bacterial symbiotic systems. The major challenge in a mixed culture system is the competition for nutrients until a steady state is achieved. When Pseudomonas and Flavobacterium were coexisting with an alga (Ankistrodesmus braunii), the rapidly growing Pseudomonas depleted all the phosphorus in the system, thereby decreasing the growth of algae to a great extent. The slow growth rate of Flavobacterium also reduced algal growth, but not to the extent of Pseudomonas. An increase in phosphate level in the media led to the increase in algal growth and the system reached steady state when the algal concentration was appropriate to provide enough dissolved organic carbon to the bacteria, thereby sustaining the algal
bacterial consortium (Mayfield & Innis, 1977). Other than extracellular organic compounds secreted by algae, the exchange of proteins is also suspected to be involved during algal
bacterial interaction (Baker & Herson, 1978). The proteins can either be toxic or stimulatory depending on the competition for nutrients and space in the growth medium (Kayser, 1979). Therefore the interactions between algae and bacteria can be considered as a complex exchange of biochemicals and biomolecules.

15.2.1 Exchange of information in the form of bioactive compounds for symbiosis Bioactive compounds are biomolecules/biochemicals released by organisms in a very minor quantity and can modulate metabolic processes. Their effects include antibiotic activity, enzyme activation or inhibition, receptor activation or inhibition, gene expression regulation, etc. In lower eukaryotes and prokaryotes, cell to cell communication is mediated by the release of bioactive compounds (small molecules, phytohormones, peptides, phytochemicals, quorum-sensing molecules, polyphenols, organosulfur compounds, etc.). Some bacteria belonging to the Cytophaga
Flavobacterium
Bacteroides complex are known to induce morphogenesis-inducing activity on algae. The germ-free axenic cultures of algae were found to differ in morphogenesis when compared with the algae in a phycosphere where the morphogenesis-inducing bacteria (MIB) were present. Some algae like Monostroma oxyspermum were able to produce spores only in the presence of MIB (Matsuo, Suzuki, & Kasai, 2003). There were several studies where bacterial lipids (sulfonolipids, lysophosphatidylethanolamines, capnione, etc.) were able to mediate developmental changes in algae by acting as molecular cues. An amino acid derivative called thalluscin was found to initiate thallus development in algae (Nishizawa, Iyenaga, & Kurisaki, 2007). A soluble chondroitin lyase secreted by Vibrio fischeri called EroS was able to induce mating in the surrounding eukaryotic organisms including algae (Woznica &

344 Chapter 15 King, 2018). The interactions between diatoms and surrounding bacteria were found to be very specific. Sulfitobacter sp. was found to produce indole-3-acetic acid (IAA), a plant hormone which promoted cell division in a diatom. The production of IAA was in turn found to be mediated by tryptophan released by the diatom (Amin, Hmelo, & Van Tol, 2015). Symbiosis between the diatom and Sulfitobacter extended with the exchange or organosulfur and nitrogen compounds. Organosulfur compounds were found to improve the photosynthetic activity of diatoms and thereby increase the dissolved organic carbon on which the bacteria can feed. 15.2.1.1 Quorum sensing Quorum sensing (QS) is a stimulus in response to cell population density. Bacteria can secrete extracellular bioactive compounds that can act as molecular cues to regulate structural, functional, and metabolic activities of nearby microorganisms. QS is regarded as the key mediator of structuring the community, formation of biofilms, virulence, swarming motility, antibiotic resistance, and secondary metabolite production. Microbes are known to sense host signals and respond accordingly in order to determine host compatibility for symbiosis or even to produce the virulence factors for pathogenesis (Freestone & Lyte, 2008). This can be extended to algal
bacterial symbiosis. The general mechanism of QS is portrayed in Fig. 15.1. Bacterial signaling molecules involved in QS are mostly acyl homoserine lactones (AHLs) and aromatic autoinducer signals (AIs). Gram-negative bacteria have a LuxR system which responds to AHL autoinducer molecules. Gram-positive bacteria have a LuxS/AI-2 system which depends on peptides for interspecies communication. A separate AI-3 system mediates interkingdom signaling in bacteria. Algae and bacteria can sense extracellular signaling molecules (QS molecules) and initiate responses of their own. The response of algae to bacterial QS molecules will depend on the function of the QS molecule. Since induction of host pathogenesis is the most common function of QS molecules, some algae have developed defense mechanism against pathogenesis-inducing QS molecules. An increase in the proteins responsive for defense against pathogens was observed during microarray analysis of plant and algae proteomes. Some algae secrete AHL-mimicking substances that confuse the bacteria, thereby hindering the QS response. Marine algae Delisea pulchra secretes halogenated furanones that are structurally related to bacterial AHLs and have an antagonistic effect on AHLs (Manefield, de Nys, & Naresh, 1999; Manefield, Rasmussen, & Henzter, 2002). The biofilm formation by harmful bacteria like Pseudomonas aeruginosa was also found to be inhibited by halogenated furanones (Rice, Givskov, & Høiby, 2002). The phycosphere of D. pulchra was found to be structured by secretion of halogenated furanones. When exposed to dimethylsulfoniopropionate (DMSP), an organosulfur metabolite synthesized by the phytoplanktons, R. pomeroyi, an α-proteobacter, secreted higher levels of a QS molecule N-(3-oxotetradecanoyl)-L-homoserine lactone. The bacteria underwent metabolic changes in

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bacterial symbiosis and its application in wastewater treatment 345

Figure 15.1 Quorum sensing (QS) signal transduction by bacteria through autoinducers (AI) and inhibition of QS signals by algae.

anticipation of phytoplankton-derived nutrients (Johnson, Kido Soule, & Kujawinski, 2016). DMSP was also a significant source of sulfur and carbon. DMSP was acting as a chemoattractant to various phytoplanktons and bacteria (Seymour, Simo, Ahmed, & Stocker, 2010). A comprehensive review on QS in algal
bacterial interactions has been done by Zhou, Lyu, and Richlen (2016).

15.2.2 Exudates that can inhibit the microbes in the vicinity Algae and bacteria can be involved in detrimental interactions. Space and nutrient limitations can lead to competition between the organisms in the phycosphere which secrete compounds that can be limiting the growth of other organisms in the vicinity. Algal and bacterial biomass are a rich source of carbon, macro-, and micronutrients. Therefore during nutrient limitation, they can be overlooked as a nutrient source and the organisms that secrete hydrolytic enzymes can digest the cellular components of competing organisms.

346 Chapter 15 The bacterial community around an algae is controlled by the toxins produced by the algae (Sison-Mangus, Jiang, Tran, & Kudela, 2014) and, similarly, compounds secreted by bacteria can control algal growth. Fatty acid compounds with inhibitory effects on bacteria were produced by Chlamydomonas reinhardii, Chlorella vulgaris, and S. quadricauda (Jørgensen, 1962). Alteromonos sp. produces 2-n-pentyl 4-quinolinol, which can regulate the growth of diatoms and cyanobacteria. At lower concentrations (,10 nM), the algae produced more mucus around its cell wall allowing the colonization of bacteria. At higher concentrations ( . 100 nM), the compound was found to inhibit the growth of algae (Long, Qureshi, Faulkner, & Azam, 2003). In a similar case, it was observed that bacteria which are in a symbiotic relationship with algae can also transform into opportunistic pathogens (Seyedsayamdost, Case, Kolter, & Clardy, 2011).

15.2.3 Exudates that can stimulate the microbes in the vicinity Algal exudates such as polysaccharides and glycoproteins are nutrient sources and contribute to the sustenance of bacterial growth. Filtered algal exudates were found to increase the growth of bacteria (including indicator bacteria) and the growth of bacteria was directly proportional to the dissolved organic matter in the exudate. The exudates were able to support the growth of bacterium that required both simple and complex organic matter (Byappanahalli, Shively, & Nevers, 2003). The exudates were generally high-molecularweight acyl polysaccharides which were rich in monosaccharides like rhamnose, fucose, arabinose, xylose, mannose, glucose, galactose, and ribose (Aluwihare & Repeta, 1999). It was also observed that some bacteria rely on reduced sulfur from exogenous sources like phytoplanktons. DMSP, the algal osmolyte, and methionine act as reduced sulfur sources for marine alpha proteobacteria. Cyanobacteria are known to fix atmospheric nitrogen and make it available for other eukaryotes that are not capable of nitrogen fixation. A special case of symbiosis was observed for Atelocyanobacterium thalassa, a unicellular prokaryotic algae which does not possess genes for Photosystem II, ribulose bisphosphate carboxylase/ oxygenase, and tricarboxylic acid cycle. It only had genes for nitrogen fixation and therefore it survives by forming a symbiotic relation with haptophytes. The haptophytes provide the cyanobacterium with fixed carbon and the cyanobacterium provide the haptophytes with fixed nitrogen (Thompson, Foster, & Krupke, 2012). Iron is an important micronutrient for algal growth. It has its role in electron transfer, nitrogen fixation, oxygen transport, DNA synthesis, and other metabolic processes. The bioavailability of iron is challenged by its ability to form colloidal particles (oxohydride polymers and ferric hydroxide), thereby making them inaccessible to the microorganisms. To tackle this problem, microorganisms, especially bacteria and cyanobacteria produce lowmolecular-weight molecules that bind and transport iron called siderophores. Eukaryotic phytoplankton don’t produce siderophores, thereby making them rely on bacterial and

Algal
bacterial symbiosis and its application in wastewater treatment 347 cyanobacterial siderophores. Soluble siderophores
Fe complexes generated by bacterial siderophores are cleaved by photooxidation, thereby recycling the siderophores. The chemistry of siderophores is briefly reviewed by Dertz and Raymond (2003). The released soluble Fe(III) iron is assimilated by both algae and bacteria. The algae thus readily assimilate iron which would have not been bioavailable in the absence of the bacterium. Algae compensates the siderophores acquisition from bacteria by providing it with fixed organic carbon which is assimilated by bacteria (Amin, Green, & Hart, 2009). The siderophores generally contain either an hydroxymate, a catecholate, or an α-hydroxy carboxylate moiety. Hydroxymate siderophores are photostable while catecholate siderophores are susceptible to photooxidation in their complexed and uncomplexed forms. α-Hydroxy carboxylate siderophores are photostable when they are free and become susceptible to photooxidation when they are bound to iron (Barbeau, Rue, & Trick, 2003). Thus the release of α-hydroxy carboxylate siderophores by bacteria increased the uptake of iron by .20-fold compared with other siderophores. Some examples of siderophores are vibrioferrin, xanthoferrin, enterobactin, bacillibactin, amonabactin, desferrioxamin B, and rhizoferrin. 15.2.3.1 Cofactor auxotrophy Vitamins are the most commonly required growth factors for all living organisms because they are cofactors for various metabolically active enzymes. Thiamine monophosphate, a precursor of active vitamin B1 (thyamine pyrophosphate), is phosphorylated by bacterial kinases to the active pyrophosphate form which is an important cofactor to the enzyme transketolase in the Calvin
Benson cycle for CO2 fixation (Kluger & Tittmann, 2008; San˜udo-Wilhelmy, Go´mez-Consarnau, Suffridge, & Webb, 2011). Algal species like Ostreococcus and Micromonas are auxotrophic to thiamine due to a missing gene in thiamine biosynthesis pathway, thereby stopping de novo synthesis of the vitamin. They are not able to salvage the vitamin from closely related compounds (thiamine diphosphate, 4-methyl 5-thiazoleethanol, 4-amino 5-hydroxymethyl 2-methyl pyrimidine, etc.) against expectations (Paerl, Bertrand, & Allen, 2015; Higgins, Gennity, & Fitzgerald, 2018). However, when cocultured with Pseudoalteromonas sp. the algae showed improved growth due to utilization of thiamine which was made available from thiamine diphosphate due to the phosphatase activity of the bacteria. Biotin (vitamin B7) binds to CO2 and therefore is an important cofactor to enzymes that require the transfer of CO2 (carboxylases, decarboxylases, and transcarboxylases) (Waldrop, Holden, & Maurice, 2012). Most of the algal species are capable of synthesizing biotin de novo. Almost all vitamin B7 auxotrophs are auxotrophic to either vitamin B12 or vitamin B1 or both. Vitamin B12 helps in inorganic carbon assimilation through methyl malonyl CoA mutase and also helps in DNA synthesis through the enzyme methionine synthase. Methionine synthase-encoding genes are of two types: metH and metE. Algae containing metE are

348 Chapter 15 capable of encoding methionine synthase even in the absence of vitamin B12, whereas metH-containing algae encode the enzyme that requires vitamin B12. Another interesting observation is that C. reinhardtii contains both the genes for Methionine synthase and preferentially encodes vitamin B12-dependent methionine synthase. Since the gene pool for encoding vitamin B12 is not present in most of the algae, vitamin B12 is obtained from bacteria through symbiosis. Bacteria was found to be associated tightly to the muciferous layer of algal cells feeding on the photosynthetic products of algae (Croft, Lawrence, & Raux-Deery, 2005). The symbiotic growth of bacteria and algae is regulated. The growth of algal
bacterial coculture was in steady state and the equilibrium was disturbed either by adding external vitamin B12 or by adding an external carbon source (Kazamia, Czesnick, & Nguyen, 2012). Thereby bacteria were able to assist the growth of vitamin auxotrophs by either salvaging or supplying the vitamins, and algae were able to assist the growth of bacteria by providing photosynthates, which are mostly carbonaceous polymers. The interactions of algae and bacteria in a symbiotic system are portrayed in Fig. 15.2.

Figure 15.2 The interactions in an algal
bacterial symbiotic system (QS, Quorum Sensing Signal; AI, autoinducers).

Algal
bacterial symbiosis and its application in wastewater treatment 349

15.2.4 Factors affecting symbiotic systems 15.2.4.1 Dissolved oxygen Dissolved oxygen (DO) within the reactor is a critical parameter affecting the symbiotic system in terms of nutrient and organic matter removal and also in terms of interactions between the algae and bacteria. Individually both the systems are dependent on DO. Aerobic heterotrophic bacteria use oxygen as the electron acceptor and its deficiency can decrease the organic content removal. Meanwhile, algal systems are adversely affected by a high oxygen concentration due to photosynthesis inhibition (Tang, Zuo, & Tian, 2016). Moreover, algae need a minimal amount of DO for respiration. Therefore in a symbiotic system, an increase in DO over a certain limit leads to a higher sludge:algae ratio. This negatively affects the stability of the symbiotic system. The organic matter removal is not affected much since the higher aeration allows for better growth of bacteria. However, the removal of ammonia nitrogen and total phosphate through assimilation decreases with higher aeration rate due to low algal biomass growth (Tang et al., 2016). Moreover, the higher aeration causes higher shear stresses, negatively affecting the formation of sludge
algae flocs (Tang et al., 2016). 15.2.4.2 Carbon dioxide This is an essential component for the photosynthesis in a symbiotic system and a higher concentration leads to a better algal growth. The use of flue gas as a CO2 source was more beneficial to the algal pond system than pure CO2 in terms of assimilation into biomass, organic matter, and total phosphate removal, and lipid content in the biomass (Posadas, Morales, del, & Gomez, 2015). 15.2.4.3 Light Light is one of the most essential components for photosynthesis and inefficient light penetration is one of the major issues frequently encountered in algal ponds. The intensity of light has a positive effect on the growth of algae and was also observed to affect the type of biomass in the reactor. However, the effect on overall chemical oxygen demand (COD) removal was not prominent. As algal growth is enhanced, nutrient removal is also improved. Lower light intensities promote the growth of algae with higher motility as they can better adjust to the change by moving along the depth (Pearson, Mara, Mills, & Smallman, 1987; Meng, Xi, & Liu, 2019). Although light is not a major factor affecting the metabolism or growth of heterotrophs, high light intensity was found to inhibit certain types of bacteria, such as the nitrite-oxidizing bacteria. Moreover, light intensities changed the biomass composition also, especially those of algae, shifting the biomass balance toward polyunsaturated fatty acids. This was hypothesized to be due to the effect of light on the biomolecules responsible for lipid synthesis (Meng et al., 2019). Light intensities up to 600

350 Chapter 15 μmol/m2/s were not inhibitive to the algal
bacterial consortia (Yang, Gou, & Fang, 2018a). Green light was found to be more effective on microalgal growth in an algal
bacterial consortium. On studying the effect of the wavelength of light on algal
bacterial consortia, it was found that at lower light intensities (250
500 μmol/m2/s), red and blue lights resulted in higher oxygen evolution rates than green and white lights. When the light intensity was further increased (500
2000 μmol/m2/s), the rate of oxygen evolution started decreasing with time for lights of blue and red wavelength, whereas oxygen evolution from green and while light stayed stable and was higher than that of red and blue lights. Therefore at higher intensities red and blue wavelength lights were found to inhibit the photosynthetic oxygenation resulting in reduced performance of algal
bacterial symbiotic systems (Kang, Kim, & Heo, 2019). Green light wavelength is more suitable for algal
bacterial symbiotic systems. The frequency of the light to dark cycle also affects the algal
bacterial symbiotic system. The light:dark ratio is directly corelated with algal: bacterial ratio. Also increasing the dark period improved COD reduction and increasing the light period improved the nutrient reduction (Lee, Lee, & Ko, 2015). Thus the intensity, wavelength, and duration of light must be optimized for an algal
bacterial symbiotic system to work efficiently. 15.2.4.4 Hydraulic retention time HRT is a parameter under consideration in a reactor utilizing the symbiotic system like algal ponds. The usual HRTs maintained in an algal pond range from 4 to 7 days. A high HRT is favored in a symbiotic system since longer residence times allow for more stable algae
bacteria flocs to be formed. This shifts the algal concentration from the supernatant to the flocs, which is found advantageous for nitrogen removal (Medina & Neis, 2007). In addition, a higher HRT allows the establishment of slow-growing microbes. A high retention time promoted the growth of cyanobacteria and heterotrophic nitrifiers which in turn improved the settling properties and nitrification in the algal pond (Cromar & Fallowfield, 1997). A very low HRT of less than 1 day may lead to the washout of biomass further leading to decreased performance. 15.2.4.5 Initial algae:bacteria ratio The ratio of initial algal and bacterial concentrations introduced into the system usually ranges from 1:1 to 1:3. A higher ratio has an advantageous effect on the symbiotic interactions between the species. This was verified by Ji, Jiang, and Zhang (2018) who reported a higher concentration of QS sensors at higher algae:biomass ratios. In general, a higher ratio improves the biomass growth in the system which in turn results in better COD and nutrient removal. However, the percentage improvement in removal efficiencies with the algae:bacteria ratio depended on the consortium under consideration (Ji et al., 2018). This could be attributed to the better symbiotic relationship between the algal and bacterial species and the corresponding increases in efficiency of mass transfer.

Algal
bacterial symbiosis and its application in wastewater treatment 351 15.2.4.6 Substrate concentrations One of the major factors governing the activity of biomass in any system is the carbon: nitrogen:phosphorus (C:N:P) ratio of the substrate. Biomass growth in symbiotic systems is usually limited by the phosphate concentrations and the feed usually have low P concentrations which affect the performance of the system drastically. The favorable C:N ratio for aerobic heterotrophic systems varies between 10 and 20, whereas those for algal systems are between 12 and 15. A high C:N:P favors COD removal and biomass growth. Organic loading rate also has an enhancing effect on the performance via microbial growth. It was observed that an increase in organic loading increased nutrient removal, whereas an increased C:N:P ratio did not cause much variation in the nitrogen removal (Cromar & Fallowfield, 1997). Considering the nitrogen source, low ammonia content limits algal growth, which in turn affects the phosphate and nitrate removal in the system (Ji et al., 2018). However, high concentrations of ammonia (5
10 mM) can cause ammonia toxicity to algae and bacteria (Konig, Pearson, & Silva, 1987). In addition, sulfides are also found to be toxic at concentrations higher than 0.055 mM to algae and 0.28 mM to bacteria (Ku¨ster, Dorusch, & Altenburger, 2005). Since the toxicity varies with each species, a shift in the microbial community structure and decrease in overall growth is observed with varying ammonia and sulfide concentrations. 15.2.4.7 pH and temperature pH is more appropriately a consequence in symbiotic systems than a factor. Photosynthesis and uptake of CO2 result in a high pH, usually around 8
10 in symbiotic systems. However, pH compounds the effects of some other key parameters, such as the amount of inorganic carbon, solubility of ammonia, etc. (Azov & Shelef, 1987). In this regard, the effect of pH is seen to be critical as the presence of undissociated ammonia and sulfide is pH dependant (Konig et al., 1987; Pearson et al., 1987). Moreover, a very high pH can inhibit the photosynthesis process (Azov & Shelef, 1987). Since all the processes involved in organic matter removal are enzyme-assisted, a higher temperature in the range of diurnal variations (25
30 C) generally leads to faster removal (Matamoros, Gutie´rrez, & Ferrer, 2015). It also favors the growth of biomass. However, low temperatures adversely affected algal growth and were observed to shift the algae:bacteria ratio by around ten times (Matamoros et al., 2015).

15.3 Applications in wastewater treatment 15.3.1 Types of reactors Treatment of wastewater with algal
bacterial symbiotic systems has been carried out in a myriad of reactor systems. These reactor systems fall under two categories,that is, pond

352 Chapter 15 systems and photobioreactors. The pond system includes high-rate algal ponds (HRAP), maturation ponds, etc. The reactor systems include batch photobioreactors, sequential batch photobioreactors, continuous photobioreactors, tubular photobioreactors, flat plate photobioreactors, and membrane photobioreactors. Maturation ponds (or stabilization ponds) were used as tertiary treatment systems for polishing the effluent. It was observed that algal
bacterial consortia in a stabilization pond inactivated virus particles more effectively than that of pure cultures of algae and bacteria separately (Sobsey & Cooper, 1973). Maturation ponds with algal
bacterial consortium can also remove trace organic compounds due to the photosynthetic oxygenation of algae and heterotrophic activity of bacteria. HRAP, on the other hand are used for the enrichment of algal biomass since they help in higher light penetration due to the reduced depth (,50 cm). Recently, algal
bacterial consortia have been deployed in HRAPs and have been used for enhanced organics and nutrients reduction from wastewater along with the enrichment of algal biomass for biofuel production (Ferrero, de Godos, & Rodrı´guez, 2012; Oh, Choi, & Kim, 2014). Also algal-associated bacteria were found to improve the settleability of the algae consortia resulting in improved harvesting of algal biomass (Lee, Cho, & Ramanan, 2013). Algal
bacterial systems can also be operated in a closed and controlled environment called photobioreactors. Photobioreactors prevent external bacterial and fungal contaminations which are major limiting factors for pond systems. Also photobioreactor models help us understand the effect of various factors on the algal
bacterial consortium by growing them in controlled environments. The reactors are designed in such a way to promote light diffusion and improve gas transfer and homogenous biomass distribution. Various configurations of photobioreactors are utilized for the treatment of wastewater with algal
bacterial consortia. Tubular photobioreactors are designed to simulate plug-flow in the reactor to improve mass transfer and light transfer. The tubes can be arranged horizontally or vertically depending on the space constraints. Horizontal tubular photobioreactors have better mixing and higher mass transfer abilities than vertical tubular photobioreactors. To improve mass transfer in vertical photobioreactors, air-lift and bubble column techniques are used. The hydrodynamics of vertical photobioreactors play an important role in algal biomass production. Higher shear may affect the biomass growth in the reactor, therefore the airflow-velocity and bubble sparging rates must be optimized to ensure proper mixing and at the same time apply only a small shear on the biomass. A superficial gas flow velocity up to 0.04 m/s is sufficient to suspend the microalgae and prevent their aggregation. Gas flow velocities higher than 0.55 m/s were detrimental to the growth of algae (Contreras, Garcı´a, Molina, & Merchuk, 1998). Reactors can be operated in batch and continuous modes. Continuous operation can be performed in cyclic tubular photobioreactors or with the use of membranes in the

Algal
bacterial symbiosis and its application in wastewater treatment 353 photobioreactors. Individual algal systems generally foul the membranes of membrane photobioreactors due to the production of extracellular polymer substances (EPS). In algal
bacterial symbiotic consortia, the EPS produced by algae are readily taken up by bacteria as a carbon source, thereby reducing membrane fouling. Also bacteria improve the setting quality of the sludge reducing membrane fouling even further. The biomass can either be grown suspended or can be grown attached to a substrate. Usually algal
bacteria symbiotic systems in a suspended system are used for organic, nitrogen, and phosphate removal. On the other hand, attached growth systems or granular systems are better in handling metals, emerging contaminant (EC), and hazardous compounds. Immobilized algae and algae combined with bacteria are used for the biosorption of heavy metals and toxic organic compounds. The algae can be immobilized in polysaccharide materials such as carrageenan, agar, alginate, synthetic matrices like acrylamide, polyurethane polyvinyl resins, chitosan, and polyvinyl foams so that harvesting of the biomass also becomes easier. Immobilization of microalgae and suspended bacteria shows more benefits for wastewater treatment and biofuel production. One major disadvantage in the suspended system is harvesting of algal biomass after treatment. However, there are flocculent bacteria which can be incorporated into symbiosis with algae which may help in solid
liquid separation after treatment. On the other hand, solid
liquid separation (i.e., algal biomass removal) is relatively better and more efficient in attached growth system. Nevertheless, biodiesel production will be based on lipid content and may depend on factors independent of the type of reactor system. The opting of any reactor is a paramount task in the case of any wastewater treatment because it incurs operation and maintenance costs based on the oxygen and light energy requirements alongside the required pollutant loading rate.

15.3.2 Nutrients removal Algae require inorganic carbon, nitrogen, phosphorus, along with the other macronutrients (such as sodium, magnesium and calcium) and trace metals for growth and maintenance purposes. Nitrogen is the second most abundant element in algal biomass (around 10%) (Kratz & Myers, 2006). Almost 30% of algal biomass is made of proteins for which nitrogen is necessary. It can be taken in the form of nitrates, nitrites, and ammonia. Some algae can fix atmospheric nitrogen (Allen & Arnon, 1955). Phosphorus is required in the range of 0.3%
3% (Zeng, Danquah, Chen, & Lu, 2011). Phosphorus can be obtained in the form of orthophosphates, polyphosphates, pyrophosphates, metaphosphates, and organic phosphates. Wastewater laden with nitrogen and phosphorus can be used as a source of nutrients for algal growth. On the other hand, excess nutrients can cause eutrophication which restricts the discharge of wastewater into water bodies. The synergistic combination of algae and bacteria can efficiently remove carbon as well as

354 Chapter 15 nutrients from wastewater and can enhance the metabolism of each other due to symbiosis. Many studies have shown that coculture of algae and bacteria enhances the nitrogen utilization capacity in algae. Table 15.1 shows the algal
bacterial symbiotic systems for the removal of nutrients from wastewater. However, the synergistic interaction depends on the types of species. For instance, the addition of bacterial species, that is, Bacillus licheniformis, has improved the NH41 consumption in C. vulgaris in a bacterial
algal mixed culture treatment system (Hu, Qi, Zhao, & Chen, 2018). Coculture of a C. vulgaris and B. licheniformis system showed better nitrogen removal than a Microsystis aeruginosa and B. licheniformis system (Ji et al., 2018). The algae and bacteria proportion for better nitrogen removal was observed to be 1:3 and C. vulgaris was able to remove 72% and 23% of N and P, respectively. Similarly, a combined system of algae and bacteria in a ratio of 1:3 showed more efficient removal of N and P from the supernatant of ozone-treated sludge when compared with individual systems (Lei, Tian, & Zhang, 2018). On the other hand, an algae to sludge (bacteria) ratio of 5:1 was found to be suitable for total nitrogen (TN) and total phosphorus (TP) removal from domestic wastewater and the corresponding removal Table 15.1: COD and nutrient removal with different algal
bacteria consortia. Wastewater Synthetic

Algae Scenedesmus quadricauda

Bacteria

Conditions

Removal efficiencies

NH41—50

NH41

HRT 5 1 day

Synthetic

Chlorella sp.

From nitrogen enriched activated sludge Activated sludge

Domestic waste water







HRT 5 10 days

Municipal wastewater

Scenedesmus sp.

Bacteria

Pretreated municipal wastewater Syntheticallymade municipal wastewater

Wastewater borne microalgae Chlorella vulgaris

Activated sludge

12h:60 h LD, followed by 12h:12 h LD cycles NH41—39.4 mg/L HRT—10 day Algae:sludge 5:1 Bacteria:algae 5 0.5 HRT 5 2 days

Leachate




Activated sludge




mg/L HRT 5 1 day

NH41—20 mg/L

Reference

Karya et al. (2013) NH41 5 99.2% Wang et al. TP 5 83.9% (2016) COD 5 87.3 C, N, and P Posadas removals of et al. 91%,70%, and 85% (2013) 92.3% COD, 95.8% Lee et al. TN, and 98.1% TP (2016) removal 5 100%

N removal 5 91% P removal 5 93%

Su et al. (2012)

N removal 5 99.8% TP removal 5 100% COD removal 5 90
95 78% in 6 days

Mujtaba et al. (2018) Sniffen et al. (2016)

LD, Light-dark; HRT, hydraulic retention time; COD, chemical oxygen demand; TP, total phosphorus; TN, total nitrogen.

Algal
bacterial symbiosis and its application in wastewater treatment 355 was 91% and 93%, respectively, within 10 days (Su, Mennerich, & Urban, 2012). The bacterial system increased the growth rate of algae from 0.1852 to 0.2182 day21; in addition, TN and TP removal of 63% and 53%, respectively, were observed after 10 days. Selection of algae and bacteria species and other environmental factors like light, pH, DO, etc., greatly affect the nutrient removal. For instance, an algal
bacterial system for municipal wastewater treatment conducted in a two-phase system [12:60 h light-dark (LD) cycle followed by 12:12 h cycle] resulted in better nutrient removal and lipid production (Lee, Oh, & Oh, 2016) (Table 15.1). However, it was reported that biomass recycling had no significant effect on nutrient removal (Ashok, Shriwastav, & Bose, 2014). In the same study, the TN and TP removal was 97% (initial of 110 mg/L) and 92% (initial of 25 mg/L), respectively, after 2 days in an algal
bacterial system. Comparison of an algal
bacterial consortium with an axenic algal system with regards to nitrogen removal shows that the combined system can operate even under HRT of 1 day resulting in low operation and maintenance costs (Jia & Yuan, 2016). A C. vulgaris and B. licheniformis system was used for a study conducted to achieve NH41 and TP removal in wastewater for a treatment period of 6 days. It was found that the combined algal
bacteria consortium showed better removal than individual wastewater treatment using algae and bacteria (Liang, Liu, & Ge, 2013). The combined system achieved 78% NH41 and 92% TP removal. On the other hand, the single algae system achieved only 29% and 55% of NH41 and TP removal, respectively. The NH41 in the bacterial system was merely 1% and TP removal was observed to be 78%. In addition, when pH was adjusted to 7, NH41 and TP removal increased to 86% and 93%, respectively. The symbiosis of C. vulgaris with bacteria from activated sludge achieved 99% and 84% of N and P removal, respectively (Wang, Liu, & Zhao, 2016). This study shows that activated sludge quality could be improved with algae along with nutrient recycling in an algal
bacteria symbiotic system. Similarly, coculturing C. vulgaris and activated sludge resulted in better TN and TP removal (99.8% and 100%, respectively, after 2 days) (Mujtaba, Rizwan, Kim, & Lee, 2018). The use of synthetic textile wastewater as media for microbes in a packed bed reactor for treating saline wastewater for combined carbon, nitrogen, and PO42 was investigated (Babatsouli, Fodelianakis, & Paranychianakis, 2015). A combination of algal species, that is, Picochlorumatomus and marine bacterial species, were used in this study. Nitrogen and PO42 removal up to 95% were obtained after 4
5 h of operation and, in addition, an increase in COD resulted in more nutrient removal. On the other hand, a high concentration of ammonia (free ammonia) can decrease nitrogen removal efficiency as reported for leachate removal with the algal
bacterial symbiotic process (Sniffen, Sales, & Olson, 2016). The treatment could achieve removal rate of 9.18 mg N/L/day (initial ammonia of 80 mg/L) in a semibatch tank reactor system. Removal of nitrogen and TP in an open algal
bacterial biofilm reactor shows promising results and TN and TP removal were 70% and 85%, respectively (Posadas, Garcı´a-Encina, & Soltau, 2013). It was seen that nutrient

356 Chapter 15 removal in an algal
bacterial system was twice that of a solo-bacterial system. The combination of algae Scenedesmus sp. with nitrifying bacteria shows that 81%
85% of ammonia was removed due to nitrifier rather than algal uptake and algae assisted the growth of nitrifier by producing O2 in the open photobioreactor (Karya, van der Steen, & Lens, 2013). Similarly, combined algal
bacterial treatment of wastewater achieved 80% TN removal through nitrification and an additional 13% through algal uptake (Wang, Yang, & Ergas, 2015). The algae accounted for 74% of the required O2 supply resulting in a cost reduction. In addition, nitrification efficiency increased with the increase in light intensity. The coimmobilization of Pseudomonas putida and C. vulgaris showed 100% removal of both TN and TP in a batch system after 2 days (Shen, Gao, & Li, 2017). Moreover, a continuous reactor employing the same algae and bacteria achieved 96% and 89% of TN (initial 50 mg/L) and TP (initial 50 mg/L), respectively, after 1 day. However, the inhibition of phosphate uptake in algae was reported in the presence of pseudomonas H4 bacteria (Chlorella sorokiniana) and therefore the consortium may not be effective in phosphate removal from wastewater (Chen, Zhao, & Wang, 2017).

15.3.3 Metal removal Metal has a toxic effect on many microorganisms, plants, animals, and humans. Many industrial effluents contain metals at higher concentrations causing danger to ecosystems. However, algae show a high tolerance to metal and have a higher capacity for metal uptake. In fact, the growth of algae requires metals in trace amounts, including iron, zinc, manganese, molybdenum, copper, and cobalt. The use of algae for metal removal is carried out using various different methods such as surface sorption, which includes adsorption, complexation, and ion exchange, and bioaccumulation where algal cells uptake metals (Cuellar-Bermudez, Aleman-Nava, & Chandra, 2017). Table 15.2 lists the algal
bacterial symbiotic systems for the removal of metals from different wastewaters. Algal
bacteria biofilm is capable of treating acid mine drainage wastewater containing metals such as Pb, Zn, and Cu (Abinandan, Subashchandrabose, Venkateswarlu, & Megharaj, 2018). Acid mine drainage wastewater was treated with a microalgae (Ulothrix sp.) and bacteria consortium in a photorotating biological contractor. The treatment resulted in 50% of Cu and Ni removal, 40%
45% of Mn removal, and 35% Zn removal in 10 days at pH 3.5 (Orandi, Lewis, & Moheimani, 2012). Algae in combination with anaerobic bacteria have been grown in a high-rate pond system to remove heavy metals like selenium along with nitrate from agricultural waste water (Gerhardt, Green, & Newman, 1991). The combined algae
anaerobic bacteria system removed 94%
100% of selenium. The symbiosis of algae and bacteria was applied in biosorption of Cu at pH 5 (initial Cu-20 mg/L) and the biosorption capacity was found to be 8.5 mg/g (Mun˜oz, Alvarez, & Mun˜oz, 2006). The bacteria were able to enhance the biosorption capacity of algal biomass.

Algal
bacterial symbiosis and its application in wastewater treatment 357 Table 15.2: Metal removal with algal
bacterial consortia. Wastewater Algae

Bacteria

Conditions

Industrial wastewater

Rhodococcus sp. Ac-1267, Kibdelosporangium aridum 754

Initial conc. (mg/L) Cu—0.035 Ni—0.21 Mn—0.19 10 days at pH 3.5

Acid mine drainage

Chlorella sp. ES13, Chlorella sp. ES-30, and Scenedesmus obliquus ES-55 Ulothrix sp.

Agricultural wastewater




Synthetic wastewater

C. sorokiniana




Selenatereducing bacteria Ralstonia basilensis

Initial conc. 5 200
400 μg/L Biosorption Initial Cu— 20 mg/L pH 5

Removal efficiencies Cu 5 62%,

Reference Safonova et al. (2004)

Ni 5 62%, Mn—70%, 50% of Cu and Ni Orandi et al. removal, 40%
45% (2012) of Mn removal and 35% Zn removal Gerhardt et al. 94%
100% (1991) Adsorption capacity Cu 5 8.5 mg/g

˜oz et al. Mun (2006)

Algal
bacterial symbiosis for remediation of industrial wastewater containing different metal and toxic organics was carried out in a wastewater pond system (Safonova, Kvitko, & Iankevitch, 2004). The symbiotic pond system was able to remove 62% Cu, 62% Ni, 90% Zn, 70% Mn, and 64% Fe from the industrial wastewater. The increased metal removal in a combined algal
bacterial system was attributed to synergistic effects rather than uptake by algal biomass which means bacteria might be able to increase the metal biosorption capacity of algae (Lei et al., 2018).

15.3.4 Organic matter removal Domestic and industrial wastewater contain organic compounds that contribute to the COD. The most common organic compounds found in wastewater are mono- and polysaccharides, proteins, lipids, organic acids, aromatic hydrocarbons, alcohols, etc. Those organic compounds that can be oxidized biologically contribute to the BOD. The remaining organics that are not biologically degradable (nbCOD) fall into refractory and recalcitrant compounds that are very hard to be degraded by biological systems. Algal
bacterial consortia have been studied extensively for organics removal in synthetic and real wastewater systems. Algae are generally autotrophic aerobes and utilize only inorganic carbon as the major carbon source. Therefore algal treatment systems are not effective for the treatment of wastewater with high organic content. Some algae (Chlorella sp., Desmosesmes sp., Spirullina sp., etc.) are capable of utilizing both organic and inorganic

358 Chapter 15 carbon and hence are called mixotrophs (Neilson & Lewin, 2010). Also some of the refractory organic compounds can be removed by algal consortia by accumulating the compound within the cell. Some of the accumulated refractory organic compounds can be degraded by the action of intracellular oxidizing and hydrolyzing enzymes (dioxygenase, cytochrome P-450 monooxygenase, epoxide hydrolases, etc.) (Semple & Cain, 1996; Semple, Cain, & Schmidt, 1999). On the other hand, bacteria are heterotrophic and can assimilate organic carbon in aerobic and anaerobic conditions. Bacterial systems can easily oxidize biodegradable compounds and accumulate refractory compounds similar to algae. Bacterial systems require an external supply of oxygen to carry out the biological degradation. When algae and bacteria exist in symbiotic relationship with each other, photosynthetic oxygenation by algae increases the DO in the water. This oxygen can then be utilized by bacteria for organic carbon mineralization. Therefore algae and bacteria in a combined system performed better with respect to organic matter removal compared with individual algal and bacterial systems. Algal
bacterial systems are able to degrade hazardous contaminants like polyaromatic hydrocarbon (PAH), phenolics, and organic solvents(Mun˜oz & Guieysse, 2006). The ratio of algae:bacteria in an algal
bacterial system plays an important role in the removal of organics. A lower ratio (1:2) was preferred for nitrogen and nutrient removal, whereas a higher ratio (1:10) was preferred for the removal of organic content (Mujtaba & Lee, 2017). The effectiveness of an algal
bacterial symbiotic system is also based on the stability of the interaction between the algae and the bacteria. The presence of higher amounts of dissolved organic carbon results in higher bacterial growth and lower algal growth which then results in destabilization of the symbiotic system. During the treatment of municipal wastewater, with an algal
bacterial consortium, a dissolved organic carbon content above 231 mg/L inhibited the growth of algae and destabilized the symbiotic system (He, Mao, & Lu¨, 2013). Photosynthetic oxygenation was important to bacteria for the mineralization of dissolved organic carbon in an algal
bacterial symbiotic system. When P. putida, a heterotrophic bacterium, was cultivated in a wastewater containing glucose, it was able to mineralize only 50% of glucose without an external oxygen supply. When it was cocultured with C. vulgaris, a eukaryotic microalga, P. putida was able to mineralize 100% of glucose without external aeration. Therefore photosynthetic oxygenation in an algal
bacterial consortium was more effective in mineralizing glucose without the requirement for external aeration (Praveen & Lon, 2015). Another important factor which is not explored much is the oxidation state of organic contaminant. When providing sources of various carbon oxidation states like methane, methanol, and glucose (24, 22, and 0, respectively), without external aeration, algal
bacterial consortia were able to degrade glucose more effectively than methane and methanol. The order of degradation was glucose . methanol . methane. The reason was cited to be the requirement of more oxygen demand for lower carbon oxidation state carbon sources like methane than that for higher carbon oxidation state molecules like

Algal
bacterial symbiosis and its application in wastewater treatment 359 glucose. Also the degradation of lower oxidation state molecules improved with bicarbonate addition since bicarbonate improves photosynthetic oxygenation (Bahr, Stams, & De La Rosa, 2011).

15.3.5 Emerging contaminants removal ECs are a new class of compounds at trace concentrations whose presence has been established in the environment. These contaminants are poorly removed in conventional treatment systems and end up being discharged into water bodies. Symbiotic processes are a sustainable and cost-friendly option considering the other effective treatment options for ECs such as advanced oxidation. Individual studies on EC removal by pure algal cultures or heterotrophic bacteria have been successfully conducted (Wang, Liu, & Kang, 2017). However, symbiotic systems are better suited for removal of ECs since they contain a variety of microbial species and are therefore more adaptable to stresses (Wang et al., 2017). While the bacterial species can effect removal through biosorption and biodegradation, algal species can also bioaccumulate the compounds. The presence of functional groups on the cell surface of microalgae and bacteria promote biosorption and thereby removal (Hom-Diaz, Jae´n-Gil, & Bello-Laserna, 2017). In addition to this, symbiotic systems are also rich in intracellular and extracellular enzyme activity. Algae in symbiotic systems can also induce some photodegradation by the release of enzymes (Matamoros et al., 2015). Many researchers have focused on the removal of ECs by algal
bacterial symbiotic systems with differing results (Table 15.3). EC removal by symbiotic systems is dependent on many factors. Sorption is dependent on the hydrophilicity, structure, functional groups, etc. of ECs. Compounds with high Kow values were found in larger concentrations in the biosolids, whereas the musk fragrances with high Henry’s constant were removed through volatilization (Matamoros et al., 2015). Compounds such as acetaminophen, ciprofloxacin, and tetracycline were found to undergo photodegradation (Bai & Acharya, 2017; de Godos, Mun˜oz, & Guieysse, 2012; Hom-Diaz et al., 2017). Removal efficiencies of ECs decreased with low influent concentrations as they are not sufficient to cause stresses in algal cells (Hom-Diaz et al., 2017; Norvill, Shilton, & Guieysse, 2016). This leads to a lack of production of oxidative enzymes and radicals as a response. The increase in HRT had varying effect on the removal of ECs. It was observed that when HRT was increased from 8 to 12 days, naproxen, lorazepam, and hydrochlorothiazole showed substantial increases in removal, whereas salicylic acid decreased from 100% to 33%. The rest of the investigated compounds did not show much variation (Hom-Diaz et al., 2017). Another of the major factors to be considered in a symbiotic system is the effect of the high diurnal variations in the pH and DO concentrations within the system (Norvill et al., 2016).

Table 15.3: Removal efficiencies of emerging contaminants (ECs) investigated with different algal
bacterial consortia. Operating Wastewater Reactor type conditions Urban wastewater

Consortium

HRAP

HRT 5 4 and 8
days; 0.47 m3; 1.54 m2; 0.3 m depth

Multitubular PBR

HRT 5 8 and 12 days; 1200 L




HRAP

14 L; HRT 5 7 days; 0.07 m2, 10 WPAR/m2; 25 6 2 C HRT 5 4 days; 500 lux; 30 6 2 C; 12
12 lightdark cycle (LD) HRT 5 15 days; 36.25 L; 100 μE/m2/s; 12
12 LD cycle

Chlorella vulgaris 1 heterotrophs

Stirred tank PBR

Algal pond

Compounds

Removal (%)

Caffeine, acetaminophen, ibuprofen, methyl dihydrojasmonate, oxybenzone, ketoprofen, hydrocinnamic acid, 5-methyl benzotriazole, naproxen, carbamazepine, galaxolide, benzothiazole, diclofenac, methyl paraben, benzotriazole, tonalide, OH-benzothiazole, tributyl phosphate, tris (2-chloroethyl) phosphate, triclosan, cashmeran, octylphenol, diazinon, celestolide, atrazine, bisphenol A, and 2,4-D Acetaminophen, ibuprofen, ketoprofen, naproxen, salicylic acid, antibiotics, atenolol, lorazepam, alprazolam, paroxetine, hydrochlorothiazole, and diatazem Tetracycline

98, 99, 99, 99, 99, 95, 89, 62, 92, 75, 84, 90, 63, 89, 95, 79, 53, 85, 85, 32.

Reference 99, 95, 97, 78, 82, 86, 93, 63,

Matamoros et al. (2015)

99, 90, 35, 69, 33, 100, Hom-Diaz .80, 57, 87, 93, 84, 77 et al. (2017)

69

de Godos et al. (2012)

Paracetamol, p-aminophenol, Chlorella sp., Pseudomonas aeuroginosa, ketoprofen, and salycilic acid Pseudominas sp., and Stenotrophomonas

100, 100, 98, and 95

Ismail et al. (2017)

Estrone, 17β-estradiol, and 17αAnabena cylindrica, ethinylestradiol Chlorococcus, Spirullina platensis, Chlorella, Scenedesmus quadricauda, and Anabena var.

80, 90, and 90

Shi, Wang, Rousseau and Lens (2010)

Algal
bacterial symbiosis and its application in wastewater treatment 361 The presence of ECs will also affect the symbiotic system. However, a lack of knowledge on the exact degradation mechanism restricts us from understanding many of these effects. A tetracycline concentration of 0.8 mg/L was observed to cause deflocculation of the algal biomass (C. vulgaris) (de Godos et al., 2012). An influent concentration of paracetamol, aspirin, and ketoprofen in the ratio 0.5/0.5/0.5 mM at 3 days HRT showed less than 20% biomass settling even after an hour. An accumulation of a nonbiodegradable metabolite, p-aminophenol, was also observed (Ismail, Essam, & Ragab, 2017). These effects are also significant considering the harvest of biomass for fuel production. In addition, EC removal through any biodegradation mechanism will cause the rise of antibiotic-resistant genes. The intermediates formed during the process are also a matter of interest. Since the symbiotic systems are complex metabolic pathways, further study on the intermediates is necessary to elucidate the actual mechanism of biodegradation of ECs.

15.3.6 Removal of refractory compounds The removal of PAHs and other hazardous pollutants from wastewater is a challenging issue. Researches were carried out with successful removal of PAH and other refractory compounds by algal-bacterial consortia (Table 15.4). The use of algal
bacterial Table 15.4: Removal efficiencies of polyaromatic hydrocarbons (PAHs) and other refractory compounds investigated with different algal
bacterial consortia. Reactor Wastewater type

Operating conditions

Coke wastewater

200 mg/L biomass, 25 C, 120 μmol/ m2/s, 2% CO2, 96 h 2500 Luxcontinuous lighting, 21
25 mL

Batch system

Batch system

Batch test

155 mL

Batch test

10 days

Consortium

Removal Compound (%)

Scendesmus quadricauda 1 activated sludge

Phenol

Chlorella sorokiniana 1 Ralstonia basilensis, Acinetobacter haemolyticus, Pseudomonas migulae, Sphingomona syanoikuyae Chlorella sorokiniana, Chlorella vulgaris, Scenedesmus obliquus, Microcystis aeuroginosa, wild Bolivian microalga 1 Ralstonia basilensis Selenastrum. capricornatum 1 M.A1-PYR

Salicylate, 100, 89, and 15 phenol, and phanthrene

HRAP, High-rate algal ponds; PBR, packed-bed reactor.

Salicylate

Pyrene

100

Rates in mg/L/ h 5 18.5, 14, 5.2, 1.3, and 17.5 100

Reference Ryu et al. (2017)

Luo et al. (2014)

362 Chapter 15 consortiums is especially beneficial for the removal of hazardous compounds like PAHs and aromatic hydrocarbons compounds with high volatility which may escape degradation if mechanical aeration is provided (Mun˜oz & Guieysse, 2006). Bacterial consortia could degrade PAHs and other aromatic compounds better than algal consortia (Hatti-Kaul, Nugier-Chauvin, Patin, & Mattiasson, 2003; Luo, Chen, & Lin, 2014). An algal culture seeded with activated sludge could achieve complete removal of phenol (430 mg/L) within 4 days from coke wastewater, whereas the same algal culture without seeded activated sludge could achieve only 6% removal of phenol from 40% coke wastewater (Ryu, Kim, Han, & Yang, 2017). Algae were seen to be inhibited by the presence of toxic compounds. An investigation studying the removal of salicylate by algal
bacterial consortia showed that various bacterial strains had a different response to salicylate concentrations. Among the various microalgal strains studied, M. aeruginosa had the least inhibitory response (Mun˜oz, Ko¨llner, Guieysse, & Mattiasson, 2003). This is a major factor to be considered when selecting appropriate strains for refractory compound removal. Mixed algal
bacterial cultures had added advantages of less aeration requirement, biomass production, etc. Degradation studies by a mixed algal
bacterial consortium showed greater than 85% removal for salicylate and phenol, whereas phenanthrene showed only 15% removal. The role of photosynthesis and synergistic relationships on removal efficiencies was demonstrated by a decrease in removals under anaerobic, dark, abiotic, and single culture (either algae or bacteria) conditions (Hatti-Kaul et al., 2003). The removal of most PAHs happened under aerobic conditions. The low oxygen levels observed during salicylate degradation indicate the prominent role of heterotrophs in the actual degradation of toxic compounds (Mun˜oz et al., 2003). The degradation of PAHs could also be attributed to the exudates and metabolites released by algal or bacterial species. Degradation of pyrene by an algal
bacterial consortium showed that the phenolic acids produced after pyrene degradation improved the algal growth which in turn increased the bacterial biomass and thereby pyrene removal (Luo et al., 2014). However, studies have also observed that the exudates from bacteria and algae did not affect the removal of the target compound (Mun˜oz et al., 2003). This calls for better criteria in the selection of suitable strains for specific compound removal.

15.4 Energy generation Microalgae biomass is a possible replacement for the current nonrenewable energy resources. Microalgae can accumulate lipids and help in biodiesel generation, also they require less space than plants and other sources of biomass. Another important criterion is that the nutrients required for algal growth can be provided by wastewater that was previously treated to remove dissolved organic carbon and that is rich in nutrients. The growth of algae and the generation of algal biomass depends on the microbiome

Algal
bacterial symbiosis and its application in wastewater treatment 363 surrounding the algae. Therefore the phycosphere has a major effect on algal growth and energy generation from algae.

15.4.1 Algal biohydrogen production Hydrogen has a very high energy density (120 MJ/kg). Hydrogen is the cleanest energy carrier among the fuels since its combustion product is water. Green algae can generate hydrogen gas under certain conditions by fixing solar energy. During photosynthesis the photolysis of water results in H2 production. The protons and electrons generated during photolysis of water are cycled to Fe-hydrogenases for effective hydrogen production. For algae to generate biohydrogen, they must be grown in a sulfur-deprived condition. During sulfur deprivation, the photosystems are inactivated leading to the disruption of photosynthate synthesis. This cell becomes oxygen-deprived and this helps the oxygensensitive Fe-hydrogenases to become active. Fe-hydrogenases are linked to the electron transport chain through ferredoxin which is activated by light energy. Also in dark anaerobic conditions, algae can concert carbohydrates into organic acids and hydrogen gas. Biohydrogen production through algal systems improved significantly when they were grown with bacterial symbionts. The bacteria increased oxygen uptake, thereby speeding up the rate of anaerobiosis which in turn improved dark anaerobic biohydrogen production in algae. Algal
bacterial consortia improved hydrogen production by 14 times when compared with axenic algal systems (Fakhimi & Tavakoli, 2019; Lakatos, Dea´k, & Vass, 2014; Wu, Li, Yu, & Wang, 2012).

15.4.2 Algal lipid production Algal lipids are used to generate biodiesel, a sustainable alternative to fossil fuels. Through photosynthesis, algae can convert CO2 and water into organic matter like carbohydrates and lipids. During ideal conditions, algae will produce carbohydrates, whereas when an external stress by limitation of nutrients arises, the algae is known to accumulate lipids instead of carbohydrates. These lipids can be converted into fatty acyl methyl esters through transesterification reactions and can be used as fuel because of their excellent energy density. Algal lipid bioaccumulation was found to increase when it was grown in symbiosis with bacteria. The activity of enzymes responsible for lipid accumulation (acyl Co-A carboxylase, diacyl glycerol acyl transferase, and phospholipid diacyl glycerol transferase) was regulated along with the increase in algal growth and biomass when grown along with bacteria. An almost 20 times increase in lipid productivity was observed by growing algae in symbiosis with bacteria compared with nitrogen limitation conditions (Xu, Cheng, & Wang, 2018). The bacteria were providing fixed nitrogen to algae, thereby improving growth in the algae. The increased lipid accumulation (around 60%) in algae accompanied a decrease in protein content in the algae when cultivated along with the bacteria under

364 Chapter 15 nitrogen depletion conditions. This may be attributed to a lower metabolic load on algae due to the symbiosis with bacteria which could have resulted in a reduction of protein content in the algae. Furthermore, because of the limited nitrogen, the biosynthetic pathways facilitate conversion of organic carbon into lipids and carbohydrates rather than proteins (Xu et al., 2018; Yen, Hu, & Chen, 2013). Mixotrophic algae grown under heterotrophic conditions was also found to accumulate lipids when cultivated along with bacteria (Leyva, Bashan, Mendoza, & de-Bashan, 2014).

15.4.3 Microbial fuel cell reactor using algal
bacteria interaction Algae can interact better with aerobic bacteria for synergistic relationships when oxygen supplied by algae are used by bacteria for their growth. However, algae cannot directly interact with anaerobic bacteria. On the other hand, the transfer of ions such as nitrates, nitrites, sulfates, etc. between photosynthetic and anaerobic microbes could be achieved with a microbial fuel cell (MFC). MFCs are bioelectrochemical systems which can produce electricity from the metabolic activities of microbes and the anaerobic oxidation of organic matter present in wastewater (Hwang, Church, Lee, & Park, 2016; Commault, Laczka, & Siboni, 2017). An MFC consists of an anode and a cathode connected through a cation exchange membrane for the transfer of ions. In a MFC employing algae, the microbes in the anode oxidize organic matter and produce CO2 and protons/electrons; thereby the produced CO2 is utilized by the algae as a carbon source (He, Zhou, & Yang, 2014). In addition, the liberated oxygen may also act as an electron acceptor. The biochemical reaction in an algal MFC is illustrated in Eqs. (15.1)
(15.3). C6 H12 O6 1 6H2 O-Bacterial biomass 1 6CO2 1 24H1 1 24e2

(15.1)

Light 1 nCO2 1 nH2 O-Algal biomass 1 ðCH2 OÞn

(15.2)

O2 1 H1 1 4e2 -2H2 O

(15.3)

An MFC can efficiently treat wastewater and produce electricity resulting in it being a costeffective and sustainable technique. A photomicrobial fuel cell (PFC) incorporating the synergistic relationship between C. vulgaris and electrochemically active bacteria achieved 99.6%, 87.6%, and 69.8% removal of carbon, nitrogen, and phosphorus, respectively (Zhang, Noori, & Angelidaki, 2011). The PFC achieved a power density of 68 mW/m2 and biomass of 0.56 g/L. An algal biofilm MFC achieved 96%, 91.5%, and 80% TN, TP, and COD removal, respectively, and a power density of 52.33 mW/m2 (Yang, Pei, & Hou, 2018b). In addition, the MFC in a continuous field trial obtained 0.094kWh/m3 of wastewater. The combination of algal MFC with activated sludge as an anode resulted in a highest voltage of 0.89 V and a power density of 1.78 W/m2 (Rashid, Cui, Muhammad, & Han, 2013). MFC using nitrogen-fixing bacteria and C. vulgaris produced 34.2 mW/m2 of

Algal
bacterial symbiosis and its application in wastewater treatment 365 electricity (Commault et al., 2017). It was observed that with the algae the electricity generated reduced to 15.2 mW/m2.

15.5 Conclusion and future directions Algal
bacterial symbiosis systems have a great potential in treating wastewater as an alternate to conventional treatment processes. Algal systems were capable of removing nutrients effectively, but their incompetence in organic carbon removal is a drawback during treatment. On the other hand, bacterial systems were capable of removing organic carbon efficiently, but were not capable of the complete removal of nutrients. Subsequently, bacterial systems also incur more operational cost due to aeration requirements. Also ECs and metals were efficiently bioadsorbed in algae and were degraded in negligible amounts by bacteria. Therefore combined algal
bacterial symbiotic systems would bridge the gap and eliminate the shortcomings of individual algal and bacterial systems. They require lesser aeration (photosynthetic aeration) and can involve the efficient recycling of nutrients and organic carbon, thereby removing the pollutants from wastewater in a complete and cost-effective manner. A myriad of factors affect the algal
bacterial symbiotic systems. Since algae and bacteria can adapt to the environmental changes quickly, algal
bacterial symbiotic systems can operate at an extended range of operating conditions like pH, temperature, DO, etc. Although the advantages outgrow any individual system of treatment, the use of algal
bacterial systems has not found its feet in wastewater treatment at the industrial scale. This can be attributed to the settleability of algae in such a system and the secretion of algal and bacterial toxins which even in trace amounts can be detrimental. Therefore the selection of algal and bacterial strains with good host compatibility and higher pollutant removal rates is necessary. Bacteria that could improve the settleability in algae must be used to improve solid
liquid separation in the reactor. Also maintaining a photobioreactor within the operating boundaries is very energy intensive. Pond systems and sequential batch systems were found to be more effective for wastewater treatment than photobioreactors in terms of operating cost. The capability of algal
bacterial symbiotic systems to treat wastewater along with energy generation makes them a viable option to operate the reactor in an energy neutral way, thereby reducing operating costs. Biohydrogen and biodiesel are the major pathways for energy generation from algae. The algal
bacterial sludge remaining after biodiesel/biohydrogen production can then be digested to generate biogas and to recover valuable by-products. The future of wastewater treatment with algal
bacterial symbiotic systems is commendable. The interactions should be explored more to understand the basis of interaction between the species and therefore selection of interacting strains with molecular level compatibility is necessary. Another important criterion is the toxicity of effluent after treatment which remains a parameter that is neglected in most of the treatment systems.

366 Chapter 15 Toxicity analysis can be done during the selection of interacting species and pretreatment to ensure safety during disposal or reuse of the effluent. Biotechnological advances have enabled the molecular level modifications of bacterial and algal strains to suit specific purposes. Genetic engineering and recombinant DNA technologies can be applied in algal
bacterial systems to enhance the specific removal of refractory compounds and ECs. Gene knockout and silencing techniques can be used to prevent the consortia from producing toxins that may hamper the effluent quality. The design of photobioreactor systems can be enhanced to exploit sunlight for algal growth in field-scale applications. Algae and bacteria can be grown on carrier systems that can aid in the higher removal of refractory and toxic organics, improve solid
liquid separation, enhance mixing, and improve light availability to the algae. Mixed algal
bacterial cultures with organisms acclimatized from the source of the contaminated wastewater have not been explored to a good extent. Such systems can help in increased productivity and pollutant removal rates.

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CHAPTER 16

Role of plant growth promoting rhizobacteria in mitigation of heavy metals toxicity to Oryza sativa L. Vishnu Kumar1, Gayatri Singh1, Rajveer Singh Chauhan2 and Geetgovind Sinam1 1

Plant Ecology & Climate Change Science Division, CSIR-National Botanical Research Institute, Lucknow, India, 2Department of Botany, Deen Dayal Upadhyaya Gorakhpur University, Gorakhpur, India

16.1 Introduction The use of microorganisms for human welfare dates way back to 7000 BCE (Vitorino & Bessa, 2017). In recent times, microbes are being widely used in the upliftment of plant growth and yield. Apart from biological control, microbes also play a key role in the functioning of plant growth by changing their physiology and metabolism. Microbes also improve nutrient uptake to plants by associating mutually, for example, mycorrhiza, algae, cyanobacteria, and plant growth promoting rhizobia (PGPR). PGPR are the rhizosphere bacteria, which can ameliorate plant growth. These bacteria are also reported to play a crucial role in availing of mineral nutrient by facilitating the recycling of plant nutrients and thereby reducing the use of chemical fertilization (Cakmakci, Do¨nmez, & Erdo˘gan, 2007). Hence these bacteria can be used as potential biofertilizers in organic farming and sustainable agriculture ecosystems. Extensive literature is available on the positive effects of PGPR in controlling plant pathogens and the enhancement of nutrient uptake from soil (Stefan et al., 2010; Zhang, Dashti, Hynes, & Smith, 1996). Rice (Oryza sativa L.) is the most consumed cereal in the world and is the staple food for more than two billion people in Asia and for millions in Africa and Latin America. One fourth of global food energy is supplied by rice. Amongst the various rice-growing countries of the world, India has the largest area under rice crop and is ranked second in production, after China. The future demand in rice production has to come from the same or even a reduced land area and hence there is a quantum requirement to increase the rice productivity (yield/ha). Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00016-X © 2020 Elsevier Inc. All rights reserved.

373

374 Chapter 16 In this industrial era, strategies of rice production management mainly focus on the use of chemical fertilizers and pesticides to enhance the yield of the crop. The persistent and injudicious use of these chemicals has detrimental effects on nontarget microorganisms of the soil and can cause undesirable changes in the environment. Furthermore, the use of fertilizers and pesticides are too expensive for the poor farmers of Asia where 90% of the world’s rice is grown (Badawi, 2004). The strategy of adopting a biological approach by using PGPR as an alternative strategy may hold the solution in overcoming the environmental hazards posed by the persistent use of synthetic chemicals. Moreover, the biological approach has a great potential in supplying essential nutrients, and simultaneously acting as a biocontrol of plant pathogens which eventually can lead to sustainable rice production. Among the different PGPR genera, the genus Azospirillum are free-living, and are known to promote the growth and yield of numerous plant species of agronomic importance. The use of rice Azospirillum as inoculants in rice production management for enhancing the growth and yield of rice plant has been reviewed by several works (Barbieri, Bernardi, Galli, & Zanetti, 1988; Hegazi, Monib, Amer, & Shokr, 1983; Kapulnik et al., 1981; Rao & Rajamamohan Rao, 1983) (Table. 16.1). The bacterial genera Azospirillum, Pseudomonas, and Methylobacterium have often been found in the rhizosphere of paddy fields; this association of bacteria with rice that is found in natural conditions is why it is used as an agricultural bioinoculant all over the world. However, the application of Azospirillum as an agent for the biocontrol of soil-borne plant pathogens is still yet to be confirmed as they lack direct suppressive chemicals or hydrolytic enzymes which could control plant pathogens. However, in recent years, literature has been available that contradicts this mechanism and is now being investigated. PGPR are able to infiltrate the endodermis barrier, entering the root cortex and from there into the vascular system, and subsequently thrive as endophytes in the stem, leaves, and roots. They are able to colonize forming an intimate association between the bacteria and the host plants without harming the plant (Compant et al., 2005). The present understanding how PGPR act as a biocontrol are mediated by the following mechanisms: by establishment of an ecological niche inside the host; competition for substrate; production of inhibitory allelochemicals; and induction of systemic resistance in the host plant to a wide range of pathogens and abiotic stresses (Be´chet et al., 2012; Gad et al., 2014; Zhang et al., 1996)

16.2 Different genera of plant growth promoting rhizobacteria Bacteria that colonize plant roots and promote growth are referred to as PGPR and constitute only 1% to 2% of the total population found in the rhizosphere. PGPR are Gramnegative rhizosphere bacteria, which can ameliorate plant growth. The class of bacteria containing PGPR comprises the following bacterial species: Pseudomonas, Azospirillum,

Role of plant growth promoting rhizobacteria 375 Table 16.1: List of plant growth promoting rhizobacteria (PGPR) reported for rice plant growth promotion under heavy metal stress. Relationship with plant

S. No. PGPR species

Metal

1.

Azospirillum sp.

Pb, Cd, and Cr

2.

Azoarcus sp.

Cr and Cd Endophytic

Leptochloa sp., Sorghum sp., and O. sativa

3.

Azotobacter sp.

Cu and Pb

Rhizospheric

Z. mays and T. aestivum

4.

Bacillus polymyxa

Rhizospheric

T. aestivum

5.

Burkholderia sp.

Endophytic

O. sativa

6.

Cyanobacteria

Rhizospheric

O. sativa and T. aestivum

7.

Gluconacetobacter diazotrophicus

Endophytic

Sorghum sp., Saccharum officinarum

8.

Herbaspirillum sp. Pb

Endophytic

O. sativa, Sorghum sp., S. officinarum, and Z. mays

9.

Klebsiella pneumoniae

Cd and Hg

Endophytic

O. sativa and T. aestivum

10.

Pseudomonas aeruginosa

Cd

Zn, Cd, Ni, Cr, and Pb Pb, Cd, and Ni

Rhizospheric

Plant name

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Arora, Sharma, and Monti (2016), Boddey, Baldani, Baldani, and ¨bereiner (1986), de Salamone, Do ¨bereiner, Urquiaga, and Boddey Do (1996), Ding, Wu, You, Huang, and Cao (2017), and Malik et al. (1997) Ding et al. (2017), Egener, Hurek, and Reinhold-Hurek (1999), Hurek, Handley, Reinhold-Hurek, and Piche´ (2002), and Stein, Hayen-Schneg, and Fendrik (1997) Mrkovacki and Milic (2001), Pandey, Sharma, and Palni (1998), and Rizvi and Khan (2018) Omar, Hamouda, and Mahrous (1996) ¨bereiner Baldani, Baldani, and Do (2000) and Nwuche, Ujam, Ihedioha, and Chime (2017) Abo-Shady (2007), Hashem (2001), and Obreht, Kerby, Gantar, and Rowell (1993) Boddey, Polidoro, Resende, Alves, and Urquiaga (2001), Isopi, Fabbri, Del Gallo, and Puppi (1995), and Sevilla, Burris, Gunapala, and Kennedy (2001) Govarthanan et al. (2016), James et al. (2002), James, Olivares, ¨bereiner (1997), and Baldani, and Do Pimentel et al. (1991) Dong et al. (2003), Gontia-Mishra, Sapre, Sharma, and Tiwari (2016), and Pramanik, Mitra, Sarkar, Soren, and Maiti (2017) Nath, Deb, Sharma, and Pandey (2014)

O. sativa

Azotobacter, Klebsiella, Enterobacter, Alcaligenes, Arthrobacter, Burkholderia, Bacillus, and Serratia (Table 16.2) (Glick, 1995; Gray & Smith, 2005; Okon & Labandera-Gonzalez, 1994). Almost all, if not all PGPRs are reported to be beneficial to plants, enhancing plant growth and improving yield production by mediating the uptake of nutrients. Furthermore, the predominant PGPR belong to the genera of Bacillus and Pseudomonas spp. (Podile & Kishore, 2007). PGPR are found to colonize the rhizosphere, the root surface (rhizoplane),

376 Chapter 16 Table 16.2: List of different genera of beneficial plant growth promoting rhizobacteria (PGPR). Different classes of PGPR Alphaproteobacteria

Betaproteobacteria Gammaproteobacteria Actinobacteria

Fermicutes

Azorhizobium caulinodans Azospirillum amazonense A. halopraeferens A. irakense A. lipoferum

Burkholderia cepacia

Klebsiella planticola

Bacillus cereus

B. gladioli

Kluyvera ascorbata

B. graminis B. vietnamensis Hydrogenophaga Pseudoflava

K. cryocrescens Citrobacter freundii Pseudomonas aeruginosa

A. brasilense Phyllobacterium rubiacearum

Curtobacterium flaccumfaciens Rathayibacter rathayi Streptomyces griseoviridis

B. coagulans B. lateroporus B. licheniformis B. macerans

P. aureofaciens P. corrugata

B. megaterium B. mycoides

P. fluorescens P. marginalis P. putida P. rubrilineans Enterobacter agglomerans E. cloacae Stenotrophomonas sp. Azotobacter chroococcum Erwinia herbicola Flavimonas oryzihabitans Serratia marcescens

B. pasteurii B. polymyxa B. pumilus B. sphaericus B. subtilis

or the root itself. The population of PGPR in the rhizosphere is 10 100 times more than the bulk soil, as the zone is rich in nutrients, amino acids, and sugars, providing a rich source of energy and nutrients for bacteria. The general characteristics of an ideal PGPR are as follows: high competence; high saprophytic capability; enhanced plant growth; easy mass multiplication; capable of performing broad actions; capable of controlling plant pathogens; not showing adverse side effects; and capable of existing symbiotically with other microbes (Jeyarajan & Nakkeeran, 2000). Apart from these roles they also facilitate the mobility of heavy metals in the soil and thus enhance the bioremediation of soil from heavy metals, such as arsenic, cadmium, mercury, and lead, by increasing the solubility of the heavy metals. Investigations into the application of PGPR strains in amending the toxicity of heavy metals by decreasing the bioavailability to plants, improving soil texture, neutralizing harmful chemicals, and controlling pests are well documented (Denton, 2007). The interaction that exists between the rhizobacteria and the plants growing therein are categorized into three basic interactions: neutral, negative, or positive (Whipps, 2001). Most of these rhizobacteria associate with plants commensally in which the bacteria establish an innocuous interaction with the host plants. In negative interactions, the pathogenic

Role of plant growth promoting rhizobacteria 377 rhizobacteria produces phytotoxic substances such as hydrogen cyanide or ethylene, thus negatively influencing the growth and physiology of the plants. In positive interaction, some PGPR exert positive affect to plant growth; the interaction may be by direct mechanisms such as the solubilization of nutrients, nitrogen fixation, production of growth regulators; or by indirect mechanisms such as the stimulation of mycorrhizae development, competitive exclusion of pathogens, or removal of phytotoxic substances. PGPR may also be categorized based upon their degree of association with the plant root cells; they are classified into extracellular plant growth promoting rhizobacteria and intracellular plant growth promoting rhizobacteria. If the populations of the PGPR exist in the rhizosphere on the rhizoplane, these groups of PGPRs are called extracellular plant growth promoting rhizobacteria, while intracellular plant growth promoting rhizobacteria are located inside the specialized nodular structures of root cells. The bacterial genera such as Agrobacterium, Arthrobacter, Allorhizobium, Azorhizobium, Azotobacter, Azospirillum, Bradyrhizobium, Bacillus, Burkholderia, Caulobacter, Chromobacterium, Erwinia, Flavobacterium, Micrococcus, Mesorhizobium, Pseudomonas, and Serratia belong to extracellular plant growth promoting rhizobacteria (Gray & Smith, 2005). The intracellular plant growth promoting rhizobacteria include the endophytes and Frankia species, and these PGPRs can symbiotically fix atmospheric N2.

16.3 Plant growth promoting rhizobacteria role in heavy metals dynamics in the soil A major percentage of the nutrients for plants are taken up from the soil. The nutrients are then channeled from the root to other parts of the plant. Excessive inputs of heavy metals in the environment by various activities have become a major concern worldwide. In the past few decades, rapid industrialization with no consideration of the implications on the environment has been imparting detrimental effects on the soil as well as on crop productivity by heavy metals pollution (Shahid et al., 2015). Pollution of soil by heavy metals and other toxic chemicals has led to alterations in soil texture, pH of soil, availability of different elements, and accumulation of heavy metals, causing direct and/or indirect reduction of plant growth, and thereby adversely affecting various physiological and molecular activities of plants (Hassan, Bano, & Naz, 2017; Panuccio, Sorgona`, Rizzo, & Cacco, 2009). Heavy metals such as Zn, Cu, Mo, Mn, Co, and Fe are essential for crucial biological processes and the developmental pathways of plant (Salla, Hardaway, & Sneddon, 2011; Shahid et al., 2015). However, the toxic metals As, Pb, Cd, and Hg are not required for normal growth and development of plants, and they reduce crop productivity when their concentration rises beyond critical threshold values (Pierart, Shahid, SejalonDelmas, & Dumat, 2015; Xiong et al., 2014). The abundance of toxic elements found in soil are in the order Pb . Cr . As . Zn . Cd . Cu . Hg. These toxic elements cause

378 Chapter 16 morphological abnormalities and metabolic disorders that can lead to reductions in yield or even senecence in plants (Amari, Ghnaya, & Abdelly, 2017). The heavy metals in soil exist in an exchangeable state, carbonate state, iron and manganese oxidation state, organic state, and residual state (Chang, Chen, Wu, Zhou, & Cheng, 2018). The exchangeable state, which constitutes only a small fraction of the total metal, is mainly adsorbed on the surface of the soil. The major fraction of metals in soil exists in the carbonate form; under acidic conditions it is readily converted to the exchangeable form. The heavy metals in iron and manganese oxidation state are adsorbed in the soil with iron and manganese oxide. This state of heavy metals can be easily dissociated to other forms. Heavy metals bonded to organic complexes are said to be in the organic state and some of them are easily soluble and are available to plants. The remaining fraction and the main form of heavy metals is the residual state. The heavy metals in the residual state are very stable and this fraction is not available for plant uptake. Microorganisms play a central role in the cycling of elements in nature (Fig. 16.1). PGPR living around or in the plant roots are efficient in transforming, mobilizing, and solubilizing the nutrients in the rhizosphere compared with those from bulk soils (Hayat, Ali, Amara, Khalid, & Ahmed, 2010). There are numerous populations of PGPR which are able to solubilize P or K from the rhizospheric soil and make it available for plant uptake (Sperberg, 1958). Different bacterial species of genera Pseudomonas, Agrobacterium, Bacillus, Rhizobium, and Flavobacterium are able to solubilize inorganic phosphate compounds, such as tricalcium phosphate, hydroxyapatite, and rock phosphate (Goldstein, 1986). Some groups of bacteria have also been known to dissolve insoluble potassium, silicon, and aluminum from mineral rock (Aleksandrov, Blagodyr, & Iiiev, 1967). PGPR produce siderophores, a low molecular mass iron chelator, which solubilize iron from minerals or organic compounds in the rhizosphere. Furthermore, siderophores also form stable complexes with other heavy metals that have the potential to cause environmental risk, such as Al, Cd, Cu, Ga, In, Pb, and Zn (Neubauer, Furrer, Kayser, & Schulin, 2000). PGPR accumulate nutrients and heavy metals present in soil by binding them as cations to the cell surface in a passive process. In rice, there are numerous PGPR species which avail the beneficial elements to plants as well as expel out the toxicants from the rhizospheric region. Azospirillum and other PGPR strains have been reported to increase the oxygenic environment in the rhizosphere, which helps the rice plant in the formation of Fe plaque in its root; The Fe plaque formed at the root of the rice root helps to oxidize As(III) into the less toxic and less mobile As(V), reducing the metalloid uptake by the plant. The PGPR Methylobacterium oryzae strain CMBM20 and Burkholderia sp. strain CMBM40, isolated from rice (O. sativa) tissues can immobilize Ni and Cd in soil and prevent it from being taken up by plants (Madhaiyan, Poonguzhali, & Sa, 2007).

Role of plant growth promoting rhizobacteria 379

Figure 16.1 Mechanism of nutrient dynamics and pathogen biocontrol by plant growth promoting rhizobacteria (PGPR) in the rhizosphere of a rice (Oryza. sativa L.) field. Source: Adapted from Tabak, H. H., Lens, P., van Hullebusch, E. D., & Dejonghe, W. (2005). Developments in bioremediation of soils and sediments polluted with metals and radionuclides—1. Microbial processes and mechanisms affecting bioremediation of metal contamination and influencing metal toxicity and transport. Reviews in Environmental Science and Biotechnology, 4(3), 115 156.

16.4 Plant growth promoting rhizobacteria’s role in controlling pathogens in rice Apart from the role of PGPR in the remediation of heavy metals and plant growth and development, they also play an important role in controlling disease by competing with soil-borne disease-causing microbes. Various diseases which are most common to rice plants, such as blight disease caused by bacteria and blast disease caused by fungi, damage the plants and also reduce the crop yield. Rice blast disease, caused by the foliar fungal

380 Chapter 16 pathogen Magnaporthe oryzae, occurs in more than 85 countries and destroys crops. It is estimated that each year this disease destroys approximately 30% of the global rice production (Nalley, Tsiboe, Durand-Morat, Shew, & Thoma, 2016). Unfortunately, to date there is no effective means to provide lasting, adequate control of the pathogen. Among the PGPR, Pseudomonas species have been well-studied for the biocontrol of pathogens as they produce a number of antimicrobial secondary metabolites (Silby, Winstanley, Godfrey, Levy, & Jackson, 2011), such as phenazines (Thomashow & Weller, 1988), hydrogen cyanide (Rudrappa, Splaine, Biedrzycki, & Bais, 2008; Voisard, Keel, Haas, & De`fago, 1989), 2,4-diacetylphloroglucinol (Raaijmakers, Weller, & Thomashow, 1997), pyrrolnitrin (Howell & Stipanovic, 1979), pyoluteorin (Howell & Stipanovic, 1980), cyclic lipopeptides and tensin (Nielsen, Thrane, Christophersen, Anthoni, & Sørensen, 2000), and viscosinamide (Nielsen, Christophersen, Anthoni, & Sørensen, 1999). Various PGPR that are present in the rhizospheric zone of the rice compete with the pathogenic bacteria and other pathogens by the mechanism of allelopathy. The most common PGPR present in the rhizospheric zone of rice are Azospirillum oryzae, Bacillus sp., Enterococcus sp., Clostridium sp., and Pseudomonas sp. In a study to analyze the effect of the bacterial endophyte Azospirillum on disease resistance in a host rice plant, the Azospirillum strain provided resistance against rice blast disease causing fungi, Magnaporthe oryzae, and a rice blight causing virulent bacterial pathogen, Xanthomonas oryzae (Yasuda, Isawa, Shinozaki, Minamisawa, & Nakashita, 2009). The role of the bacteria strains PBZ1 and B510 in controlling pathogens in the rice plant were studies. It was reported that neither salicylic acid (SA) accumulation nor expression of pathogenesisrelated genes was induced by interaction with this bacterium. The authors indicated that the strain B510 was able to induce disease resistance in rice by activating a novel type of resistance mechanism independent of SA-mediated defense signaling. Sivamani, Anuratha, and Gnanamanickam (1987) reported that the strains of P. fluorescens were shown to inhibit the growth of Xanthomonas oryzae pv. oryzae. Gad et al. (2014), in their experiment on the effect of bacterial strain P. aeuroginosa in controlling aggregate sheath spot caused by the fungi Rhizoctonia oryzae-sativae, showed that the PGPR could inhibit the mycelia growth by 35% and also promoted the growth of the plant.

16.5 Plant growth promoting rhizobacteria in remediation of the environment Heavy metal stress has become a major concern in various spheres of ecosystems worldwide. Nowadays extensive industrialization and injudicious use of resources imparts detrimental effects on the environment, particularly on soil where the build-up of heavy metals has increased manifold (Shahid et al., 2015). Plants acquire their essential nutrient elements from the soil. However, along with these metals other highly toxic heavy metals

Role of plant growth promoting rhizobacteria 381 like arsenic, cadmium, chromium, lead, and mercury are also taken up by the plant if the soil is contaminated by these metals (Pierart et al., 2015; Xiong et al., 2014). To withstand heavy metal stress and metal toxicity, plants have evolved numerous defense mechanisms, viz., reduced heavy metal uptake, sequestration of metal into vacuoles, binding to phytochelatins/metallothioneins, and activation of various antioxidants (Shahid et al., 2015). In addition to the abovementioned tolerance mechanisms against heavy metals, plant also mutually associate with beneficial soil microbes, which enhances the capability of plants to cope with heavy metal stress. The functioning of PGPR in heavy metal polluted soil can be affected by both the micropartner (plant-associated bacteria) and the host plant. Therefore the adaptation capabilities of both the partners of the associative plant as well as the bioremediation potential of the PGPR are of importance to minimize the detrimental effect of heavy metal pollution. To overcome the metal stress, PGPR have evolved a number of mechanisms by which they able to tolerate the uptake of heavy metals. Such mechanisms include pumping toxic metal ions out of the cell, the accumulation and sequestration of the metal ions inside the cell, transforming the toxic metal to less toxic forms (Wani, Khan, & Zaidi, 2008), and adsorption/desorption of metals (Mamaril, Paner, & Alpante, 1997; Nies, 2000) (Fig. 16.1). As PGPR, like other bacteria, have plasmids in their cell, which carry resistant genes for various heavy metals. This trait of PGPR makes it an ideal microorganism for the remediation of heavy metal polluted fields. So far resistance genes have been identified from PGPR for the heavy metals zinc, cadmium, and cobalt. These genes are believed to code for the membrane efflux proteins and are probably of ATPases types. PGPR are also reported to provide some tolerance to plants from arsenic, chromium, and cadmium (Roane & Pepper, 2000). The use of PGPR has so far been limited to growthpromoting agents in agronomic practices; substantial emphasis is now being placed on them in order to exploit their bioremediation potential as well. PGPR could be used as an alternate strategy for sustainable and ecologically prudent agricultural production. The major rice-growing area of India is infested with arsenic pollution in its soil. The rice plant growing in these areas, besides taking up essential nutrients from the soil, also takes up heavy metals like arsenic, cadmium, etc. The consumption of rice grown in such areas is becoming a major source of arsenic for humans (Sohn, 2014). To ameliorate the negative effects of heavy metals to plants, PGPR could be harnessed. Numerous PGPR are known to exhibit plant growth promoting traits under heavy metal stress. They impart favorable effects on plants either by direct or indirect mechanisms, such as biofilm formation, siderophores, exopolysaccharide, and phytohormones production (Tiwari, Lata, Chauhan, & Nautiyal, 2016; Tiwari, Prasad, Chauhan, & Lata, 2017). The remediation of heavy metals by PGPR does not involve any transgenic modifications, therefore it is ethically and societally acceptable remedial technique. The investigation of heavy metal tolerance in plants through microbial remediation has been going on for many years; there is still considerable interest in plant microbe metal

382 Chapter 16 association studies due to their direct effects on enhanced biomass production and heavy metal tolerance (Glick, 2003; Hansda & Kumar, 2017; Taj & Rajkumar, 2016). In rice, there are numerous PGPR species which avail the beneficial elements to plants as well as expel the toxicants from the rhizospheric region. A number of researchers have used Azospirillum and other PGPR strains to reduce the accumulation of As in rice plants, as it increases the oxygenic environment in the rhizosphere and increases Fe plaque formation, thus helping to oxidize As(III) to (V) in soils (Jia, Huang, Chen, & Zhu, 2014; Liu et al., 2006). Iron plaques render As less available to plants for assimilation in the immediate vicinity of the root. The PGPR Pantoea sp. strain EA106 has been reported to significantly increase Fe plaque formation, thereby attenuating As uptake by rice roots (Lakshmanan et al., 2015). In another study, a strain of Pseudomonas maltophilio is reported to transform the mobile and toxic form of chromium (Cr(VI)) to the nontoxic and immobile form (Cr(III)) and also minimized the mobility of other toxic ions, such as Hg21, Pb21 and Cd21 (Blake et al., 1993; Park, Keyhan, & Matin, 1999). In most cases PGPR help to supply beneficial elements to the plants and detoxify or metabolize the toxic elements into the nontoxic forms by various mechanisms, which are biosorption, bioaccumulation, chemisorption of metals, biomineralization, biotransformation, and bioleaching (Fig. 16.1). 1. Biosorption and bioaccumulation: biosorption is a process of remediation in which active/inactive microbial biomass binds to and concentrates heavy metals from any aqueous media. Various microbes remove heavy metals and other toxic elements from contaminated sites by associating with them. In bioaccumulation, solutes are transported from the outside the cell of PGPR through the cellular membrane into the cell cytoplasm, where the metal is sequestered. Biosorption describes the association of soluble substances with the cell surface. PGPR are capable of absorbing heavy metals in far greater amounts compared to the charge density of the cell surface. Scientists have demonstrated that charged functional groups serve as nucleation sites for the biosorption of various metal-bearing precipitates. There are three possible mechanisms through which microbes can biosorb heavy metals from contaminated sites: (1) sorption on the bacterial cell surface site; (2) additional surface complexation and precipitation of actinides; and (3) precipitation of actinides with bacterial cell lysates (Tabak, Lens, van Hullebusch, & Dejonghe, 2005). The Gram-negative bacteria have a lesser ability to absorb metals, possibly because they lack organic phosphate groups on their cell surface. Bacteria and Archea have a crystalline proteinaceous surface layer on their cell surface structures which is called the S-layer. The lack of S-layer is attributed for the decreased sorption ability of Gram-negative bacteria. 2. Biotransformation: biotransformation is the process by which one chemical species of an element or compound changes into another (oxidative or reduced form) by the enzymes secreted by PGPR. Metal-reducing PGPRs can reduce a wide variety of the

Role of plant growth promoting rhizobacteria 383 multivalent metals, both toxic and nontoxic, present in the soil. Direct enzymatic reduction involves the use of oxidized forms of metals as electron acceptors. The oxidized forms of some metals are highly soluble in aqueous media and are generally very mobile, while the reduced species are highly insoluble and often form precipitates rendering the metal unavailable for plant uptake. Certain PGPR are known to reduce the highly soluble Cr(IV) to less soluble Cr(III) ions (Tabak et al., 2005). 3. Chemisorption of metals: chemisorption is a kind of adsorption in which the adsorbed substance is held by chemical bonds. In other words we can say that the chemical binding of an atom or molecule (the adsorbate) to a solid surface (the adsorbentor substrate) is termed chemisorption. Many microbes which belong to PGPR, such as E. coli, Citrobacter sp., and others, play a crucial role in metal homeostasis as well as availability to plants. PGPR remove toxic heavy metals from contaminated environments by bonding with or by converting them into nontoxic forms. Immobilized cells of a Citrobacter sp. have been reported to removed Np and Pu from solution using an established technique that uses biologically produced phosphate ligands (Pi) for metal phosphate bioprecipitation (Macaskie & Basnakova, 1998). 4. Biomineralization: the process by which living organisms produce minerals inside their tissue to harden or stiffen it is called biomineralization. Based on the organism involved, the biomineralization process occurs through two mechanisms, namely a passive mechanism where it is biologically induced or microbially controlled, and active biomineralization. Most of the minerals present in soil, including carbonates, phosphates, sulfates and sulfides, arsenates, silica, chlorides, fluorides, oxides, hydroxides, and Fe-Mn-oxides, are known to be biomineralized by PGPR present in the rhizosphere (Skinner, 2005). Besides iron oxides, phosphates are the main class which together represent about 25% of the biogenic minerals usually formed in a biologically controlled fashion, except for struvite and brushite, which are formed by the biologically induced mechanism (Konhauser & Riding, 2012; Weiner & Dove, 2003). 5. Bioleaching: bioleaching is the extraction of metals from their ores using living organisms. In other words, bioleaching is a process described as the use of microorganisms to transform elements, so that the elements can be extracted from a material when water is filtered through it. Some of the commonly used microorganisms for bioleaching are Thiobacillus thiooxidans, T. ferrooxidans, Bacillus licheniformis, B. luteus, B. megaterium, B. polymyxa, and Leptospirillum ferrooxidans. The process of bioleaching occurs through two reaction processes: direct and indirect leaching. In direct leaching, bacteria and metal are in direct physical contact and oxidation of minerals takes place by the enzymes released by the PGPR. In indirect leaching, the metals are not in direct contact with PGPR but are oxidized by the leaching agents produced by microorganisms. Bioleaching has emerged as an ecologically sustainable technology in recent years for recovering metals from soil.

384 Chapter 16

16.6 Conclusion and future prospect Globally most of the total cultivated and irrigated agricultural area is contaminated with different heavy metals and metalloids including arsenic, zinc, cadmium, selenium, mercury, etc. and this contamination is increasing rapidly every year. Due to this reason, much of the cultivatable area will be affected by heavy metals/metalloids in the coming years. Heavy metals/metalloids have detrimental effects on plants, such as oxidative damage to plants and their photosynthetic machinery, growth retardation, and ultimately yield loss. However, the rhizosphere of plants have a diverse community of microbes known as PGPR that have the potency to cope with the problem of heavy metal stress. PGPRs can bind to heavy metals and form minerals. The use of PGPR has so far been limited to growth-promoting agents in agronomic practices, but a substantial emphasis is now being placed on them in order to exploit their bioremediation potential as well. PGPR could be used as an alternate strategy for sustainable and ecologically prudent agricultural production.

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Role of plant growth promoting rhizobacteria 387 Neubauer, U., Furrer, G., Kayser, A., & Schulin, R. (2000). Siderophores, NTA, and citrate: Potential soil amendments to enhance heavy metal mobility in phytoremediation. International Journal of Phytoremediation, 2(4), 353 368. Nielsen, T. H., Christophersen, C., Anthoni, U., & Sørensen, J. (1999). Viscosinamide, a new cyclic depsipeptide with surfactant and antifungal properties produced by Pseudomonas fluorescens DR54. Journal of Applied Microbiology, 87(1), 80 90. Nielsen, T. H., Thrane, C., Christophersen, C., Anthoni, U., & Sørensen, J. (2000). Structure, production characteristics and fungal antagonism of tensin—a new antifungal cyclic lipopeptide from Pseudomonas fluorescens strain 96.578. Journal of Applied Microbiology, 89(6), 992 1001. Nies, D. H. (2000). Microbial heavy-metal resistance. Applied Microbiology and Biotechnology, 51, 451 460. Nwuche, C. O., Ujam, O. T., Ihedioha, J. N., & Chime, C. C. (2017). An “ex situ” microbial process for the removal of heavy metals from polluted soil: A case study of Ada rice field, Adani, Enugu State, Nigeria. Bioremediation Journal, 21(3-4), 128 137. Obreht, Z., Kerby, N. W., Gantar, M., & Rowell, P. (1993). Effects of root-associated N2-fixing cyanobacteria on the growth and nitrogen content of wheat (Triticum vulgare L.) seedlings. Biology and Fertility of Soils, 15(1), 68 72. Okon, Y., & Labandera-Gonzalez, C. A. (1994). Agronomic applications of Azospirillum. Improving plant productivity with rhizosphere bacteria (pp. 274 278). Adelaide: Commonwealth Scientific and Industrial Research Organization. Omar, M. N. A., Hamouda, A. M., & Mahrous, N. M. (1996). Evaluating the efficiency of inoculating some diazotrophs on yield and protein content of 3 wheat cultivars under graded levels of nitrogen fertilization. Annals of Agricultural Science, Ain-Shams University (Egypt), 41(2), 579 590. Pandey, A., Sharma, E., & Palni, L. M. S. (1998). Influence of bacterial inoculation on maize in upland farming systems of the Sikkim Himalaya. Soil Biology and Biochemistry, 30(3), 379 384. Panuccio, M. R., Sorgona`, A., Rizzo, M., & Cacco, G. (2009). Cadmium adsorption on vermiculite, zeolite and pumice: Batch experimental studies. Journal of Environmental Management, 90(1), 364 374. Park, C. H., Keyhan, M., & Matin, A. (1999). Purification and characterization of chromate reductase in Pseudomonas putida. In: 99th General meeting of the American Society for Microbiology (pp. 536 554). Chicago, IL, USA. Pierart, A., Shahid, M., Sejalon-Delmas, N., & Dumat, C. (2015). Antimony bioavailability: Knowledge and research perspectives for sustainable agricultures. Journal of Hazardous Materials, 289, 219 234. Pimentel, J. P., Olivares, F., Pitard, R. M., Urquiaga, S., Akiba, F., & Do¨bereiner, J. (1991). Dinitrogen fixation and infection of grass leaves by Pseudomonas rubrisubalbicans and Herbaspirillum seropedicae. Nitrogen fixation (pp. 225 229). Dordrecht: Springer. Podile, A. R., & Kishore, G. K. (2007). Plant growth-promoting rhizobacteria. Plant-associated bacteria (pp. 195 230). Dordrecht: Springer. Pramanik, K., Mitra, S., Sarkar, A., Soren, T., & Maiti, T. K. (2017). Characterization of cadmium-resistant Klebsiella pneumoniae MCC 3091 promoted rice seedling growth by alleviating phytotoxicity of cadmium. Environmental Science and Pollution Research, 24(31), 24419 24437. Raaijmakers, J. M., Weller, D. M., & Thomashow, L. S. (1997). Frequency of antibiotic-producing Pseudomonas spp. in natural environments. Applied and Environmental Microbiology, 63(3), 881 887. Rao, J. L. N., & Rao, V. R. (1983). Nitrogenase activity in the rice rhizosphere soil as affected by Azospirillum inoculation and fertilizer nitrogen under upland conditions. Current Science (India), 52, 686 688. Rizvi, A., & Khan, M. S. (2018). Heavy metal induced oxidative damage and root morphology alterations of maize (Zea mays L.) plants and stress mitigation by metal tolerant nitrogen fixing Azotobacter chroococcum. Ecotoxicology and Environmental Safety, 157, 9 20. Roane, T. M., & Pepper, I. L. (2000). Microorganisms and metal pollution. In R. M. Maier, I. L. Pepper, & C. B. Gerba (Eds.), Environmental microbiology (p. 55). London: Academic Press. Rudrappa, T., Splaine, R. E., Biedrzycki, M. L., & Bais, H. P. (2008). Cyanogenic pseudomonads influence multitrophic interactions in the rhizosphere. PLoS One, 3(4), e2073.

388 Chapter 16 Salla, V., Hardaway, C. J., & Sneddon, J. (2011). Preliminary investigation of Spartina alterniflora for phytoextraction of selected heavy metals in soils from Southwest Louisiana. Microchemical Journal, 97(2), 207 212. Sevilla, M., Burris, R. H., Gunapala, N., & Kennedy, C. (2001). Comparison of benefit to sugarcane plant growth and 15N2 incorporation following inoculation of sterile plants with Acetobacter diazotrophicus wildtype and nif mutant strains. Molecular Plant Microbe Interactions, 14(3), 358 366. Shahid, M., Khalid, S., Abbas, G., Shahid, N., Nadeem, M., Sabir, M., . . . Dumat, C. (2015). Heavy metal stress and crop productivity. Crop production and global environmental issues (pp. 1 25). Cham: Springer. Silby, M. W., Winstanley, C., Godfrey, S. A., Levy, S. B., & Jackson, R. W. (2011). Pseudomonas genomes: Diverse and adaptable. FEMS Microbiology Reviews, 35(4), 652 680. Sivamani, E., Anuratha, C. S., & Gnanamanickam, S. S. (1987). Toxicity of Pseudomonas fluorescens towards bacterial plant pathogens of banana (Pseudomonas solanacearum) and rice (Xanthomonas campestris pv. oryzae). Current Science (Bangalore), 56(12), 547 548. Skinner, H. C. W. (2005). Biominerals. Mineralogical Magazine, 69(5), 621 641. Sohn, E. (2014). The toxic side of rice. Nature, 514(7524), S62 S63. Sperberg, J. I. (1958). The incidence of apatite-solubilizing organisms in the rhizosphere and soil. Australian Journal of Agricultural and Resource Economics, 9, 778. Stefan, M., Dunca, S., Olteanu, Z., Oprica, L., Ungureanu, E., Hritcu, L., . . . Cojocaru, D. (2010). Soybean (Glycine max [L] Merr.) inoculation with Bacillus pumilus Rs3 promotes plant growth and increases seed protein yield: Relevance for environmentally-friendly agricultural applications. Carpathian Journal of Earth and Environmental Sciences, 5(1), 131 138. Stein, T., Hayen-Schneg, N., & Fendrik, I. (1997). Contribution of BNF by Azoarcus sp. BH72 in Sorghum vulgare. Soil Biology and Biochemistry, 29(5-6), 969 971. Tabak, H. H., Lens, P., van Hullebusch, E. D., & Dejonghe, W. (2005). Developments in bioremediation of soils and sediments polluted with metals and radionuclides—1. Microbial processes and mechanisms affecting bioremediation of metal contamination and influencing metal toxicity and transport. Reviews in Environmental Science and Bio/Technology, 4(3), 115 156. Taj, Z. Z., & Rajkumar, M. (2016). Perspectives of plant growth-promoting actinomycetes in heavy metal phytoremediation. In S. Gopalakrishnan, A. Sathya, & Rajendran (Eds.), Plant growth promoting actinobacteria (pp. 213 231). Singapore: Springer. Thomashow, L. S., & Weller, D. M. (1988). Role of a phenazine antibiotic from Pseudomonas fluorescens in biological control of Gaeumannomyces graminis var. tritici. Journal of Bacteriology, 170(8), 3499 3508. Tiwari, S., Lata, C., Chauhan, P. S., & Nautiyal, C. S. (2016). Pseudomonas putida attunes morphophysiological, biochemical and molecular responses in Cicer arietinum L. during drought stress and recovery. Plant Physiology and Biochemistry, 99, 108 117. Tiwari, S., Prasad, V., Chauhan, P. S., & Lata, C. (2017). Bacillus amyloliquefaciens confers tolerance to various abiotic stresses and modulates plant response to phytohormones through osmoprotection and gene expression regulation in rice. Frontiers in Plant Science, 8, 1510. Vitorino, L. C., & Bessa, L. A. (2017). Technological microbiology: Development and applications. Frontiers in Microbiology, 8, 827. Voisard, C., Keel, C., Haas, D., & De`fago, G. (1989). Cyanide production by Pseudomonas fluorescens helps suppress black root rot of tobacco under gnotobiotic conditions. The EMBO Journal, 8(2), 351 358. Wani, P. A., Khan, M. S., & Zaidi, A. (2008). Chromium-reducing and plant growth-promoting Mesorhizobium improves chickpea growth in chromium-amended soil. Biotechnology Letters, 30(1), 159 163. Weiner, S., & Dove, P. M. (2003). An overview of biomineralization processes and the problem of the vital effect. Reviews in Mineralogy and Geochemistry, 54(1), 1 29. Whipps, J. M. (2001). Microbial interactions and biocontrol in the rhizosphere. Journal of Experimental Botany, 52(1), 487 511.

Role of plant growth promoting rhizobacteria 389 Xiong, T., Leveque, T., Shahid, M., Foucault, Y., Mombo, S., & Dumat, C. (2014). Lead and cadmium phytoavailability and human bioaccessibility for vegetables exposed to soil or atmospheric pollution by process ultrafine particles. Journal of Environmental Quality, 43(5), 1593 1600. Yasuda, M., Isawa, T., Shinozaki, S., Minamisawa, K., & Nakashita, H. (2009). Effects of colonization of a bacterial endophyte, Azospirillum sp. B510, on disease resistance in rice. Bioscience, Biotechnology, and Biochemistry, 73(12), 2595 2599. Zhang, F., Dashti, N., Hynes, R. K., & Smith, D. L. (1996). Plant growth promoting rhizobacteria and soybean [Glycine max (L.) Merr.] nodulation and nitrogen fixation at suboptimal root zone temperatures. Annals of Botany, 77(5), 453 460.

Further reading Beattie, G. A. (2007). Plant-associated bacteria: Survey, molecular phylogeny, genomics and recent advances. Plant-associated bacteria (pp. 1 56). Dordrecht: Springer. Elliott, L. F., & Lynch, J. M. (1995). The international workshop on establishment of microbial inocula in soils: Cooperative research project on biological resource management of the Organization for Economic Cooperation and Development (OECD). American Journal of Alternative Agriculture, 10(2), 50 73. Gill, S. S., & Tuteja, N. (2010). Reactive oxygen species and antioxidant machinery in abiotic stress tolerance in crop plants. Plant Physiology and Biochemistry, 48(12), 909 930. Hider, R. C., & Kong, X. (2010). Chemistry and biology of siderophores. Natural Product reports, 27(5), 637 657. Holguin, G., & Patten, C. L. (1999). Biochemical and genetic mechanisms used by plant growth promoting bacteria. World Scientific. Iqbal, M., Ahmad, A., Ansari, M. K. A., Qureshi, M. I., Aref, I. M., Khan, P. R., . . . Hakeem, K. R. (2014). Improving the phytoextraction capacity of plants to scavenge metal (loid)-contaminated sites. Environmental Reviews, 23(1), 44 65. John, S. G., Ruggiero, C. E., Hersman, L. E., Tung, C. S., & Neu, M. P. (2001). Siderophore mediated plutonium accumulation by Microbacterium flavescens (JG-9). Environmental Science and technology, 35(14), 2942 2948. Kao, P. H., Huang, C. C., & Hseu, Z. Y. (2006). Response of microbial activities to heavy metals in a neutral loamy soil treated with biosolid. Chemosphere, 64(1), 63 70. Kiss, T., & Farkas, E. (1998). Metal-binding ability of desferrioxamine B. Journal of Inclusion Phenomena and Molecular Recognition in Chemistry, 32(2 3), 385 403. Li, Y., Pang, H. D., He, L. Y., Wang, Q., & Sheng, X. F. (2017). Cd immobilization and reduced tissue Cd accumulation of rice (Oryza sativa wuyun-23) in the presence of heavy metal-resistant bacteria. Ecotoxicology and Environmental Safety, 138, 56 63. Merzaeva, O. V., & Shirokikh, I. G. (2006). Colonization of plant rhizosphere by actinomycetes of different genera. Microbiology, 75(2), 226 230. Mishra, M., Sahu, R. K., Sahu, S. K., & Padhy, R. N. (2005). Effect of vermicomposted municipal solid wastes on growth, yield and heavy metal contents of rice (Oryza sativa). Fresenius Environmental Bulletin, 14(7), 584 590. Nabulo, G., Black, C. R., & Young, S. D. (2011). Trace metal uptake by tropical vegetables grown on soil amended with urban sewage sludge. Environmental Pollution, 159(2), 368 376. Oves, M., Khan, M. S., & Zaidi, A. (2013). Chromium reducing and plant growth promoting novel strain Pseudomonas aeruginosa OSG41 enhance chickpea growth in chromium amended soils. European Journal of Soil Biology, 56, 72 83. Pourrut, B., Jean, S., Silvestre, J., & Pinelli, E. (2011). Lead-induced DNA damage in Viciafaba root cells: Potential involvement of oxidative stress. Mutation Research/Genetic Toxicology and Environmental Mutagenesis, 726(2), 123 128.

390 Chapter 16 Rajkumar, M., Ae, N., Prasad, M. N. V., & Freitas, H. (2010). Potential of siderophore-producing bacteria for improving heavy metal phytoextraction. Trends in Biotechnology, 28(3), 142 149. Sabaratnam, S., & Traquair, J. A. (2002). Formulation of a Streptomyces biocontrol agent for the suppression of Rhizoctonia damping-off in tomato transplants. Biological Control, 23(3), 245 253. Sahoo, R. K., Ansari, M. W., Pradhan, M., Dangar, T. K., Mohanty, S., & Tuteja, N. (2014). A novel Azotobacter vinellandii (SRI Az 3) functions in salinity stress tolerance in rice. Plant Signaling and Behavior, 9(7), 511 523. Umrania, V. V. (2006). Bioremediation of toxic heavy metals using acidothermophilic autotrophes. Bioresource Technology, 97(10), 1237 1242. Velusamy, P., & Gnanamanickam, S. S. (2003). Identification of 2,4-diacetylphloroglucinol production by plantassociated bacteria and its role in suppression of rice bacterial blight in India. Current Science, 85(9), 1270 1273. Velusamy, P., Immanuel, J. E., Gnanamanickam, S. S., & Thomashow, L. (2006). Biological control of rice bacterial blight by plant-associated bacteria producing 2,4-diacetylphloroglucinol. Canadian Journal of Microbiology, 52(1), 56 65. Zhuang, X., Chen, J., Shim, H., & Bai, Z. (2007). New advances in plant growth-promoting rhizobacteria for bioremediation. Environment International, 33(3), 406 413.

CHAPTER 17

Study of transport models for arsenic removal using nanofiltration process: recent perspectives Robin Marlar Rajendran, Sangeeta Garg and Shailendra Bajpai Department of Chemical Engineering, Dr B R Ambedkar National Institute of Technology, Jalandhar, India

17.1 Introduction Arsenic is the one of the pollutants which causes contamination to groundwater. The Comprehensive Environmental Response, Compensation and Liability Act, ranked arsenic in first place for toxicity among the heavy metals (ATSDR, 2010). Regular consumption of arsenic-contaminated water for drinking purposes may cause serious health hazards to humans. Many countries have reported contamination due to arsenic, especially South Asian countries, including Bangladesh and India. A higher concentration of arsenic in groundwater is found in many parts of West Bengal, Bihar, Punjab, and many other states of India. The Ministry of Drinking Water and Sanitation, Government of India reported that 17 districts of Uttar Pradesh, India have been affected by arsenic contamination recently (Sabha, n.d.). Arsenic-affected states of India with respect to number of habitations are shown in Fig. 17.1. The treatment of arsenic-contaminated water is an emerging challenge for researchers.

17.1.1 Sources Arsenic is available in the form of a metalloid and as a chemical compound in environment (Shih, 2005). Anthropogenic and natural sources are the two main sources, as classified in Fig. 17.2 (Oliveira, Gonc¸alves, Oliveira, & Guilherme, 2008; Xia et al., 2007). Anthropogenic or human-made sources are due to industry, agricultural applications, mining activity, etc. Mineral intrusion into groundwater, volcanic ash eruptions, weathering conditions, and biological conditions are the natural sources. Arsenic can enter into groundwater through either of the two sources listed above. Both sources can lead to Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00017-1 © 2020 Elsevier Inc. All rights reserved.

391

392 Chapter 17 Number of Habitations state wise in India 12,000

10,928

No of Habitations 10,000 8000 6000 4518 4000 2000 0

871 Assam

Bihar

102

102

Jharkhand

Punjab

748 Uttar Pradesh West Bengal

Figure 17.1 State-wise arsenic-affected habitations (Sabha, n.d.). Arsenic source Natural conditions

Anthropogenic Industries (petroleum, ceramic, etc.)

Agricultural applications (herbicide, pesticide, etc.)

Mining activity (physical forces such as grinding, crushing, pulvarization, etc.)

Mineralized groundwater

Volcanic ash

Weathering conditions

Biological activity

Figure 17.2 Sources of arsenic (Oliveira et al., 2008; Xia et al., 2007).

arsenic entering and spreading through the environment and increasing the concentration of arsenic in the environment, especially in groundwater (Shih, 2005). 17.1.1.1 Anthropogenic sources Anthropogenic activities facilitate the mobilization of arsenic into the environment. Commercial wastes, industrial processes (including the mining industry, petroleum industry, ceramic, coal ash, pesticides, etc.), and wood preservatives are the major causes of arsenic contamination in many cities and industrial areas (Dong, Ma, Gress, Harris, & Li, 2014). Industries are manufacturing arsenic-containing products, such as calcium arsenate [Ca3(AsO4)2], lead arsenate (PbHAsO4), magnesium arsenate [Mg3(AsO4)2], sodium arsenite (NaAsO2), and Paris green [Cu(C2H3O2)2
3Cu(AsO2)2]. Some of these arsenic products are used for aquatic weed control which easily leads to the intrusion of arsenic into groundwater (Peryea & Creger, 1994).

Study of transport models for arsenic removal using nanofiltration process 393 17.1.1.2 Natural sources The natural cycle is contributing one third of the arsenic (Smith, Naidu, Alston, & Donald, 1998). Mineralaquifer interaction over a long period of time results in arsenic and other minerals entering into the groundwater (Choong, Chuah, Robiah, Gregory Koay, & Azni, 2007). Many arsenic-bearing minerals are available in nature such as arsenopyrite, chalcopyrite, galena, and pyrite. The minerals in which the arsenic concentration is above 10% by weight are detailed in Table 17.1. An interesting fact is that the pure metal form of arsenic is rarely found in nature. Arsenic can easily dissociate from arsenic-bearing materials and enter groundwater. High-temperature natural processes such as volcanic activity and hydrothermal processes also cause arsenic emissions into the environment (Sarkar & Paul, 2017).

17.1.2 Health effects More than 100 million people have been affected by arsenicosis worldwide (Jha, Mishra, Damodaran, Sharma, & Kumar, 2017). Arsenic exposure may affect liver, lungs, kidney, nasal passages, skin, etc., (Harisha, Hosamani, Keri, Nataraj, & Aminabhavi, 2010). Diseases caused by the consumption of arsenic-contaminated water are called arsenicosis. Consuming arsenic-free water at the initial stage of arsenicosis could help to get rid of the later stages. Four stages of arsenicosis and its effects on human health is mentioned in Table 17.2. The fourth stage of arsenicosis can lead to cancer and other diseases.

17.1.3 Permissible limit The permissible limit for arsenic in drinking water is 0.01 mg/L according to the Central Pollution Control Board, India. The World Health Organization (WHO) lowered the Table 17.1: Most common arsenic-bearing ore minerals. Chemical S. no. Ore mineral formula

Occurrence

Composition

1.

Arsenopyrite FeAsS

Metamorphic rocks

2.

Chalcopyrite CuFeS2

3.

Galena

PbS

4.

Pyrite

FeS2

5.

Realgar

As4S4

Sedimentary region and mafic igneous rocks Hydrothermal veins and metamorphic deposits Hydrothermal veins as an accessory mineral Hydrothermal veins and volcanic sublimations

Compound of iron, arsenic, and sulfur Compound of copper, ferrous, and sulfate Compound of lead and sulfur Compound of iron sulfide, cobalt, nickel, silver, and gold Compound of sulfur

Source: Mindat.org - Mines, Minerals and More, n.d. Retrieved July 4, 2018, from https://www.mindat.org/.

394 Chapter 17 Table 17.2: Four stages of arsenicosis in human health. S. no.

Stages

Effects on human health

1.

Preclinical

2.

Clinical

3.

Complications

4.

Malignancy

a. b. a. b. c. a. b. a. b.

No major symptoms Arsenic can be detected in the human body Skin issues Swelling of hands and feet 510 years of treatment needed to get cure Enlargement of liver, kidneys, and spleen Conjunctives and diabetes Cancers or tumors Tongue or bladder cancer

Source: Choong, T. S. Y., Chuah, T. G., Robiah, Y., Gregory Koay, F. L., & Azni, I. (2007). Arsenic toxicity, health hazards and removal techniques from water: An overview. Desalination, 217(13), 139166. https://doi.org/10.1016/j. desal.2007.01.015.

Japan Hungary China Countries

Europian unit Mexican Vietnam French Germany WHO United States India 0

0.01

0.02

0.03

0.04

0.05

0.06

Arsenic concentration (mg/L) Permissible limit of arsenic in drinking water

Figure 17.3 Permissible limit of arsenic in drinking water in various countries (Abedin et al., 2002; Choong et al., 2007; Gorchev & Ozolins, 2011; Nimick, Moore, Dalby, & Savka, 1998; Urase, Oh, & Yamamoto, 1998; Xia et al., 2007).

permissible limit value for arsenic from 0.05 to 0.01 mg/L in 1993 due to the toxic nature of arsenic (Abedin, Cotter-Howells, & Meharg, 2002). China’s permissible limit of arsenic in drinking water was 0.05 mg/L. Later, The National Construction Ministry of China changed the permissible limit to 0.01 mg/L (Xia et al., 2007). Countries like Germany, Australia, Japan, Hungary, French, the United States, etc. also have reduced the permissible limit of arsenic in drinking water to either 0.01 mg/L or below (Choong et al., 2007). Fig. 17.3 compares the permissible limit of arsenic in drinking water in various countries. To summarize, arsenic can be considered as a toxic heavy metal which has to be removed by treatment, especially in drinking water.

Study of transport models for arsenic removal using nanofiltration process 395

17.2 Chemistry of arsenic Arsenic is a metalloid (can act as a metal and nonmetal), that is, insoluble in water and soluble in oxidizing acids. For example, if you drop arsenic rock in a glass of pure water it will not dissolve easily. If we allow organic carbon and microorganisms in the right proportions, arsenic minerals can be released into water. Arsenic can exist in various forms in the environment (23, 0, 13, and 15) (Thirunavukkarasu, Viraraghavan, Subramanian, & Tanjore, 2002). The main elemental arsenic species come in two forms, that is, As(III)/arsenite (metal) and As(V)/arsenate (nonmetal). As(III) is more dangerous than As(V) (Kumar, Chaudhari, Khilar, & Mahajan, 2004; Smith et al., 1998). The structures of As(III) and As(V) is shown in Fig. 17.4; these are the dominating arsenic species in groundwater. Physicochemical characteristics of arsenic species are mentioned in Table 17.3.

17.3 Methods of arsenic removal from water/wastewater Heavy metals can be removed from aqueous solution by various practices such as adsorption (Bajpai et al., 2012; Giles, Mohapatra, Issa, Anand, & Singh, 2011), bioremediation (Gorchev & Ozolins, 2011; Urase et al., 1998), chemical precipitation (Harper & Kingham, 1992), coagulation and flocculation (Han, Runnells, Zimbron, & Wickramasinghe, 2002; Song et al., 2006), electrodialysis (Ortega et al., 2017), ion exchange (Kim & Benjamin, 2004), membrane distillation (Criscuoli & Figoli, 2018), membrane separation process (Uddin, Mozumder, Figoli, Islam, & Drioli, 2007), and phytoremediation (Gorchev & Ozolins, 2011; Urase et al., 1998). Combinations of techniques with oxidation have been tried by researchers and it has been suggested that the oxidation process (converting As(III) into As(V)) helps to improve the overall efficiency of rejection (Teychene et al., 2014). It also helps in the detoxification of drinking water. Sen et al. (2010) combined oxidation and nanofiltration (NF) for the removal of arsenic. Increasing the oxidant concentration, improved the efficiency from 60% to 98% which shows the significance (Sen et al., 2010). HO

HO

HO

As

HO

HO

As

HO

Arsenite/As(III)

Arsenate/As(V)

Figure 17.4 Structure of As(III) and As(V).

O

396 Chapter 17 Table 17.3: Basic chemical properties of arsenic. S. no. Properties

Toxic semimetallic element

1. 2.

33 Group 15; located in VA (directly below phosphorous)

Atomic no. Periodic table

3.

Predominant oxidation states a. Environment b. Drinking water

As (-III), As (0), As(III), and As(V) As(III) and As(V)

4.

Stable forms of arsenic Arsenates As(V) Arsenites As(III)

5.

Under aerobic/oxidizing conditions Anaerobic/reducing conditions

Arsenic species

Molar weight (g/mol) Hydrodynamic radius (nm)

HAsO22 4

139.9 125.9 140 125

(arsenate ion) H3AsO3 (arsenious acid) H 2 AsO2 4 (arsenate ion) H2 AsO2 3 (arsenite ion)

0.78 0.24 0.59 0.24

Source: Data from Arsenic in drinking-water. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans, 84, 41267. https://doi.org/10.1016/j.kjms.2011.05.002; Ergican, E., Gecol, H., & Fuchs, A. (2005). The effect of co-occurring inorganic solutes on the removal of arsenic (V) from water using cationic surfactant micelles and an ultrafiltration membrane. Desalination, 181 (13), 926. https://doi.org/10.1016/j.desal.2005.02.011; Harisha, R. S., Hosamani, K. M., Keri, R. S., Nataraj, S. K., & Aminabhavi, T. M. (2010). Arsenic removal from drinking water using thin film composite nanofiltration membrane. Desalination, 252 (13), 7580. https://doi.org/10.1016/j.desal.2009.10.022; Nguyen, C. M., Bang, S., Cho, J., & Kim, K. W. (2009). Performance and mechanism of arsenic removal from water by a nanofiltration membrane. Desalination, 245(13), 8294. https://doi.org/10.1016/j. desal.2008.04.047; Urase, T., Oh, J. I., & Yamamoto, K. (1998). Effect of pH on rejection of different species of arsenic by nanofiltration. Desalination, 117(13), 1118. https://doi.org/10.1016/S0011-9164(98)00062-9; Vrijenhoek, E. M., & Waypa, J. J. (2000). Arsenic removal from drinking water by a “loose” nanofiltration membrane. Desalination, 130(3), 265277. https://doi.org/10.1016/ S0011-9164(00)00091-6; Xia, S., Dong, B., Zhang, Q., Xu, B., Gao, N., & Causseranda, C. (2007). Study of arsenic removal by nanofiltration and its application in China. Desalination, 204(13 SPEC. ISS.), 374379. https://doi.org/10.1016/j.desal.2006.04.035.

17.3.1 Membrane technology Membrane

Pore size

Molecular weight cutoff (MWCO)

Pressure range (kPa)

MF UF NF RO

0.11 μm 0.010.1 μm 110 nm Nonporous

.1000 kDa .10 kDa .200 .100

100400 200700 6001000 14007000

MF, Microfiltration; UF, ultrafiltration; NF, nanofiltration; RO, reverse osmosis.

Membrane separation process is a widely used technique for arsenic removal without a phase change. Membrane separation is a filtration process through a thin-film membrane layer which is a synthetic material. The driving force for the membrane separation process can be a pressure, concentration, chemical potential, and electrical force (Mulder, 1996). The available membrane technologies are microfiltration (MF), ultrafiltration (UF), NF, and reverse osmosis (RO) (Uddin et al., 2007). Pore size, molecular weight cutoff (MWCO), and pressure range of MF, UF, NF, and RO are compared in Fig. 17.5.

Study of transport models for arsenic removal using nanofiltration process 397 Membrane

Pore size

MWCO

Pressure range (kPa)

MF

0.1–1 µm

>1000 kDa

100–400

UF

0.01–0.1 µm

>10 kDa

200–700

NF

1–10 nm

>200

600–1000

RO

Nonporous

>100

1400–7000

Figure 17.5 Comparison of membrane techniques [pore size, molecular weight cutoff (MWCO), and pressure range] (Yoon, 2016). Source: Data from Yoon, S.-H. (2016). Classification of membranes according to pore size, OnlineMBR y. Retrieved from , http://onlinembr.info/membrane-process/classification-of-membranesaccording-to-pore-size/ . .

17.4 Nanofiltration of arsenic NF is a proficient technique to treat arsenic-contaminated water. Separation of ions is based upon diffusion, convection, and electromigration (Figoli et al., 2010). Moreover, the NF process requires low transmembrane pressure compared to RO and therefore cost reduction is the biggest advantage of it (Fang & Deng, 2014a, 2014b). Criscuoli and Figoli (2018) compared RO, NF, and membrane distillation (MD) processes for arsenic removal and found that rejection efficiency is in the order of NF . RO . MD (as per flux). This implies that the optimized best process for arsenic removal is NF.

17.4.1 Modeling of nanofiltration membranes for arsenic removal The purpose of modeling is to optimize the process performance and to predict the characteristics of the membrane. The most extensively used model for NF membranes is based on the extended Nernst Plank (ENP) equation as shown in Eq. (17.1) (De´on, Dutournie´, Fievet, Limousy, & Bourseau, 2013). Ji 5 2 Di;P

dci zi Fci Di;P dϕ 1 Ki;c ci V 2 3 dx dx Rg T

(17.1)

Pal et al. (2012) adapted the Donnan steric pore model (DSPM) for arsenic rejection using three different polyamide NF membranes (NF1, NF20, and NF2). Filtration was combined

398 Chapter 17 with preoxidation of arsenite (As(III)) to get complete conversion into arsenate (As(V)) by using KMnO4 as an oxidant. Oxidation of arsenite in the presence of KMnO4 takes place per the following reactions: 1 2 KMnO2 4 -K 1 MnO4

(17.2)

2 51 1 MnO2 1 2H2 O As31 1 MnO2 4 1 4H 2 -As

(17.3)

The DSPM model assumes that the membrane contains identical cylinderical pores in series. Hindrance factors were included for the correction of convection and diffusion terms. The osmotic pressure difference (Δπ) and salvation energy barrier (ΔWi ) are neglected in this model. The transport of ions is described in terms of pore size, membrane thickness, and membrane charge density. Real groundwater was taken from the Bengal Delta Basin, India (Pal et al., 2012). This modeling study was derived only for a binary system. However, in real groundwater both coions and counterions are present, therefore following a binary system may mean the simulated results are misleading. Okhovat and Mousavi (2012) modeled the NF process performance using genetic programming (GP) for three different heavy metals such as arsenic, chromium, and cadmium removal. This model assumed that rejection of arsenic is the function of two independent variables: feed concentration and transmembrane pressure. The function was normalized to match the unit levels of individual factors. The termination criterion for the normalized feed concentration (x1) and transmembrane pressure (x2) was (Okhovat & Mousavi, 2012): cos (sin cos cos cos( 2 log (cosx2 2 (x1 x2 ÞÞÞ 3 log (sin cos cos x2 Þ 3 cos x1 3  (cos (sinx2 1 cos(2x1 2 cos(x2 sin(sinx2 3 logx1  ðx1 1 x2 ÞÞÞÞÞ 2 cosx1 2 x 1^2ÞÞ (17.4) The R2 value obtained was 0.98 which confirms that experimental data fitted well with the GP model. However, the GP model fails to evaluate membrane parameters such as pore size, membrane thickness, etc. This modeling study did not consider other important parameters such as feed flowrate, pH, etc. Fang and Deng (2014a) studied arsenic rejection in dead-end filtration using two NF membranes (TFC-SR2 and NF270). The authors applied various models separately to model the rejection of arsenite and arsenate. Modeling of the arsenite (As(III)) rejection was done using the SpieglerKedem steric hindrance model (SK-SHM). Model fitting was very poor for all the experimental studies because experimental rejection was lower than the calculated rejection values. The authors explained the reason for this poor fitting is due the adsorption of As(III) to the polyamide membrane which produces an H2 bond from the As (III) molecule and N from the polyamide membrane. SK-SHM is considered only steric

Study of transport models for arsenic removal using nanofiltration process 399 hindrance in the model not adsorption. Another reason is that the neutral As(III) molecule is weakly bounded with shells. It can detach from the hydration layer that facilitates transport through the membrane. As(V) rejection was modeled in terms of mass transfer coefficient, concentration at membrane surface (Cm), and membrane charge density (Xd) using DSPM coupled with concentration polarization (CP). The mass transfer coefficient was calculated using Eq. (17.5) for the stirring speed experiments (Fang & Deng, 2014a). Use of film theory enhanced the model prediction of As(V) rejection. However, the modeling results were not good for As(III) removal.   ωr  v stirrer 0:567 (17.5) Di =rstirrer k 5 0:23 DN v Fang and Deng (2014b) characterized two thin-film composite NF mambranes (DK and DL membrane) in a cross-flow system. Authors used CP film theory, HagenPoiseuille equation, SK-SHM, and DSPM to calculate the membrane parameters and to analyze the rejection of arsenate. The coupling motion of components was neglected because feed solutions are diluted in the DSPM model mentioned by the authors. Dielectric interactions were also neglected due to the larger pore size of the membrane. The concentration of arsenate at the membrane surface was evaluated by CP film theory. The SK-SHM was used to calculate the effective pore radius by fitting the rejection of glucose into the model. Membrane thickness to porosity ratio was evaluted using the HagenPoiseuille equation. Membrane charge density was obtained by fitting the chloride rejection data into DSPM. Fourth-order RungeKutta method was used to calculate the ENP equation. Diffusive transport; 2 Ki;d Di;N Electromigration transport;

dci 100 3 Ji dx

ci;avg 3 zi Ki;d Di;P F dΨ 100 3 3 RT dx Ji


100 Convective transport; 2Ki;c Jv ci;avg 3 Ji

(17.6) (17.7) (17.8)

The percentage contributions of each transport mechanism (diffusion, electromigration, and convection transport mechanisms) are calculated using a one-step central difference estimate as mentioned in Eqs. (17.6) and (17.8) (Fang & Deng, 2014b). The authors concluded that diffusion transport is directly proportional to concentration gradient. Arsenate diffusive transport was dominated compared to convective transport and electromigration transport for both DK and DL membranes. Jadhav et al. (2016) investigated the performance of two NF membranes (NF90 and NF270) for the simultaneous removal of arsenic, fluoride, nitrate, and sulfate. The central composite design (CCD) approach was used in this work. The design included six factors: cross-flow velocity, transmembrane pressure, feed concentration of arsenic, fluoride, nitrate, and

Table 17.4: Recent literature overview of transport models for arsenic separation using NF. Model

Membrane

Donnan steric NF1 PA, Sepro pore model membranes Inc. (United States) NF2 PA, Sepro membranes Inc. (United States) NF20 PA, Sepro membranes Inc. (United States) Genetic UTC-70 UB Flat sheet programming aromatic PA, Toray Japanese Company TFC SR 2, Koch DSPM combined with CP and SK-SHM NF270, Dow Filmtec

DSPM combined with CP and SK-SHM RSM  CCD; multiple solute model

DK TFC membranes, GE Osmonics DL TFC membranes, GE Osmonics PA NF90, Dow Filmtec PA NF270, Dow Filmtec

Feed solution (mg/L)

Pressure (bar)

pH

TA 5 0.1704 (arseniccontaminated groundwater)

516

10

Oxidant used

%R

KMnO4 98.5 (8 mg/L) (NF1);

Parameters

Mode and area (cm2)

References

rp ; Δx; Xd

Cross flow (100 cm2)

Pal, Chakraborty, and Roy (2012)

%R

Cross flow (63.6 cm2)

Okhovat and Mousavi (2012) Fang and Deng (2014a)

96 (NF2)

95 (NF20) As(V) 5 0.1 0.4 514 (simulated water) As(V) 5 0.0480.100; As(III) 5 0.0550.108; (simulated water)

15

7.1



410 

As(V) 5 8.3 and 0.0520.886 13.79 (simulated water)

410 

TA 5 0.10.2 515 (simulated water)

7

84.993

% R; 9093 Cm ; k; Xd ; [As(V)] , 40 [As (III)] 8891 [As(V)] , 36 [As (III)] 43.196.3 rp ; Δx Ak ; Cm ; k; Xd 55.895.5



9398

Dead-end system (28.2 cm2)

Cross flow Fang and (138.7 cm2) Deng (2014b)

Jv, % R, Cg, Pm Cross flow (160 cm2)

Jadhav, Marathe, and Rathod (2016)

CP, concentration polarization; RSM, response surface methodology; CCD, Central composite design; SK-SHM, SpieglerKedem steric-hindrance model; PA, Polyamide; PS, polysulfone; TA, total arsenic; DSPM, Donnan steric pore model; NF, nanofiltration.

Study of transport models for arsenic removal using nanofiltration process 401 sulfate. The multiple solute model was used to estimate membrane resistance, mass transfer coefficient from gel polarization term, and true rejection. The simulation was done using the GaussNewton algorithm in MATLab. The results showed less than 5% error. The effect of ionic strength has shown no effect on arsenic removal (Jadhav et al., 2016). However, solutesolute interactions and electronic migrations are not considered in the multiple solute model, which is the main drawback of this approach. The recent study models for arsenic removal are compared in Table 17.4. It has been found that there is enough scope for the modeling of As(III) rejection using the NF membrane. In most of the work, a binary system was assumed for the modeling approach. A multiple solute model which contained different heavy metals tried to model the arsenic rejection. However, solutesolute interactions were neglected in this model, which as mentioned previously is a major drawback.

17.5 Conclusion and future perspective In this chapter, the recent modeling approaches for arsenic removal using NF membranes have been discussed. It was found that the NF process can be a promising option for the rejection of arsenic from an aqueous stream. Better efficiency of As(V) rejection can be achieved solely using NF. In the case of As(III), rejection efficiency can be enhanced by combining oxidation with NF. We have discussed various transport models applied for the NF process. Transport of As(V) covered diffusion, convection, and electromigration terms. Researchers preferred to study As(V) over As(III) in transport modeling. As(III) is electrically neutral and did not have an electromigration term in the transport models. A binary system has been considered for most of the modeling studies. Extensive research is needed for the modeling of arsenic rejection in the presence of foreign ions. Not much modeling work has been done for arsenic rejection with real groundwater, bigger surface areas of membrane, pilot-plant systems, higher concentrations of arsenic in feed, co- and counterions in the feed, spiral wound modules, and oxidant additions. These can be the future objectives for transport models for arsenic removal using NF membranes.

Nomenclature Am Δci CF ci Cp Cg Di,P DB F Jv

membrane area concentration difference of ion “i” across the pore thickness feed concentration of ion “i” concentration of ion in pore concentration of permeate ion “i” concentration of a compound in gel layer pore diffusivity of ion “i” bulk diffusivity Faraday constant (96,487 C/mol) volumetric flux

402 Chapter 17 Ki,c Lp Pe Pm Rg ri rp rstirrer T v Xd

convective diffusion, dimensionless hydraulic permeability Peclet number, dimensionless permeability coefficient gas constant (8.314 m3 kPa/mol/K) ionic radius ore radius radius of stirrer temperature (K) stirring speed membrane charge density (mol/m)

Greek symbols ϕD ΔϕD Ψm;x50 ψm;x5Δx Θ ω σ

potential Donnan potential (ψm;x5Δx 2 Ψm;x50 Þ=Δx membrane potential at feed side membrane potential at permeate side steric partition coefficient radius of the membrane cell reflection coefficient

Abbreviations CCD CP DSPM ENP GP IMIS MF MSP MWCO NF PA PS RSM SK SK-SHM TMP WHO

central composite design concentration polarization Donnan steric pore model extended NerstPlank model (ENP model) genetic programming integrated management information system microfiltration membrane separation processes molecular weight cutoff nanofiltration polyamide membrane polysulfone membrane response surface methodology SpieglerKedem model SpieglerKedem steric hindrance model transmembrane pressure World Health Organization

References Abedin, M. J., Cotter-Howells, J., & Meharg, A. A. (2002). Arsenic uptake and accumulation in rice (Oryza sativa L.) irrigated with contaminated water. Plant and Soil, 240(2), 311319. Available from https://doi. org/10.1023/A:1015792723288.

Study of transport models for arsenic removal using nanofiltration process 403 ATSDR. (2010). 2007 CERCLA priority list of hazardous substances. Department of Health and Human Sciences—Agency for Toxic Substances & Disease Registry. Available from https://doi.org/http://www. atsdr.cdc.gov/cercla/05list.html. Bajpai, S., Gupta, S. K., Dey, A., Jha, M. K., Bajpai, V., Joshi, S., & Gupta, A. (2012). Application of Central Composite Design approach for removal of chromium (VI) from aqueous solution using weakly anionic resin: Modeling, optimization, and study of interactive variables. Journal of Hazardous Materials, 227228, 436444. Available from https://doi.org/10.1016/j.jhazmat.2012.05.016. Choong, T. S. Y., Chuah, T. G., Robiah, Y., Gregory Koay, F. L., & Azni, I. (2007). Arsenic toxicity, health hazards and removal techniques from water: An overview. Desalination, 217(13), 139166. Available from https://doi.org/10.1016/j.desal.2007.01.015. Criscuoli, A., & Figoli, A. (2018). Pressure-driven and thermally-driven membrane operations for the treatment of arsenic-contaminated waters: a comparison. Journal of Hazardous Materials. Available from https://doi. org/10.1016/J.JHAZMAT.2018.07.047. De´on, S., Dutournie´, P., Fievet, P., Limousy, L., & Bourseau, P. (2013). Concentration polarization phenomenon during the nanofiltration of multi-ionic solutions: Influence of the filtrated solution and operating conditions. Water Research. Available from https://doi.org/10.1016/j.watres.2013.01.044. Dong, X., Ma, L. Q., Gress, J., Harris, W., & Li, Y. (2014). Enhanced Cr(VI) reduction and As(III) oxidation in ice phase: Important role of dissolved organic matter from biochar. Journal of Hazardous Materials, 267, 6270. Available from https://doi.org/10.1016/j.jhazmat.2013.12.027. Ergican, E., Gecol, H., & Fuchs, A. (2005). The effect of co-occurring inorganic solutes on the removal of arsenic (V) from water using cationic surfactant micelles and an ultrafiltration membrane. Desalination, 181(13), 926. Available from https://doi.org/10.1016/j.desal.2005.02.011. Fang, J., & Deng, B. (2014a). Arsenic rejection by nanofiltration membranes: Effect of operating parameters and model analysis. Environmental Engineering Science. Available from https://doi.org/10.1089/ees.2013.0460. Fang, J., & Deng, B. (2014b). Rejection and modeling of arsenate by nanofiltration: Contributions of convection, diffusion and electromigration to arsenic transport. Journal of Membrane Science, 453, 4251. Available from https://doi.org/10.1016/j.memsci.2013.10.056. Figoli, A., Cassano, A., Criscuoli, A., Mozumder, M. S. I., Uddin, M. T., Islam, M. A., & Drioli, E. (2010). Influence of operating parameters on the arsenic removal by nanofiltration. Water Research, 44(1), 97104. Available from https://doi.org/10.1016/j.watres.2009.09.007. Giles, D. E., Mohapatra, M., Issa, T. B., Anand, S., & Singh, P. (2011). Iron and aluminium based adsorption strategies for removing arsenic from water. Journal of Environmental Management. Available from https:// doi.org/10.1016/j.jenvman.2011.07.018. Gorchev, H. G., & Ozolins, G. (2011). WHO guidelines for drinking-water quality. WHO chronicle (Vol. 38). https://doi.org/10.1016/S1462-0758(00)00006-6. Han, B., Runnells, T., Zimbron, J., & Wickramasinghe, R. (2002). Arsenic removal from drinking water by flocculation and microfiltration. Desalination, 145(13), 293298. Available from https://doi.org/10.1016/ S0011-9164(02)00425-3. Harisha, R. S., Hosamani, K. M., Keri, R. S., Nataraj, S. K., & Aminabhavi, T. M. (2010). Arsenic removal from drinking water using thin film composite nanofiltration membrane. Desalination, 252(13), 7580. Available from https://doi.org/10.1016/j.desal.2009.10.022. Harper, T., & Kingham, N. (1992). Removal of arsenic from wastewater using chemical precipitation methods. Water Environment Research. Available from https://doi.org/10.2175/WER.64.3.2. Jadhav, S. V., Marathe, K. V., & Rathod, V. K. (2016). A pilot scale concurrent removal of fluoride, arsenic, sulfate and nitrate by using nanofiltration: Competing ion interaction and modelling approach. Journal of Water Process Engineering, 13, 153167. Available from https://doi.org/10.1016/j.jwpe.2016.04.008. Jha, S. K., Mishra, V. K., Damodaran, T., Sharma, D. K., & Kumar, P. (2017). Arsenic in the groundwater: Occurrence, toxicological activities, and remedies. Journal of Environmental Science and Health—Part C: Environmental Carcinogenesis and Ecotoxicology Reviews, 35(2), 84103. Available from https://doi.org/ 10.1080/10590501.2017.1298359. Kim, J., & Benjamin, M. M. M. M. (2004). Modeling a novel ion exchange process for arsenic and nitrate removal. Water Research, 38(8), 20532062. Available from https://doi.org/10.1016/j.watres.2004.01.012.

404 Chapter 17 Kumar, P. R., Chaudhari, S., Khilar, K. C., & Mahajan, S. P. (2004). Removal of arsenic from water by electrocoagulation. Chemosphere. Available from https://doi.org/10.1016/j.chemosphere.2003.12.025. Mindat.org - Mines, Minerals and More. (n.d.). Retrieved July 4, 2018, from https://www.mindat.org/. Mulder, M. (1996). Basic Principles of Membrane Technology. Kluwer Academic Publishers. Available from https://doi.org/10.1016/0376-7388(92)85058-Q. Nguyen, C. M., Bang, S., Cho, J., & Kim, K. W. (2009). Performance and mechanism of arsenic removal from water by a nanofiltration membrane. Desalination, 245(13), 8294. Available from https://doi.org/ 10.1016/j.desal.2008.04.047. Nimick, D. A., Moore, J. N., Dalby, C. E., & Savka, M. W. (1998). The fate of geothermal arsenic in the Madison and Missouri Rivers, Montana and Wyoming. Water Resources Research, 34(11), 30513067. Available from https://doi.org/10.1029/98WR01704. Okhovat, A., & Mousavi, S. M. (2012). Modeling of arsenic, chromium and cadmium removal by nanofiltration process using genetic programming. Applied Soft Computing Journal, 12(2), 793799. Available from https://doi.org/10.1016/j.asoc.2011.10.012. Oliveira, D. Q. L., Gonc¸alves, M., Oliveira, L. C. A., & Guilherme, L. R. G. (2008). Removal of As(V) and Cr(VI) from aqueous solutions using solid waste from leather industry. Journal of Hazardous Materials, 151(1), 280284. Available from https://doi.org/10.1016/j.jhazmat.2007.11.001. Ortega, A., Oliva, I., Contreras, K. E., Gonza´lez, I., Cruz-Dı´az, M. R., & Rivero, E. P. (2017). Arsenic removal from water by hybrid electro-regenerated anion exchange resin/electrodialysis process. Separation and Purification Technology. Available from https://doi.org/10.1016/j.seppur.2017.04.050. Pal, P., Chakraborty, S., & Roy, M. (2012). Arsenic separation by a membrane-integrated hybrid treatment system: Modeling, simulation, and techno-economic evaluation. Separation Science and Technology (Philadelphia). Available from https://doi.org/10.1080/01496395.2011.652754. Peryea, F. J., & Creger, T. L. (1994). Vertical distribution of lead and arsenic in soils contaminated with lead arsenate pesticide residues. Water, Air, and Soil Pollution. Available from https://doi.org/10.1007/BF00483038. Sabha, L. (n.d.). Government of India Ministry of Drinking Water & Sanitation. Retrieved from http://www. indiaenvironmentportal.org.in/files/file/ContaminatedDrinkingWater_2.pdf. Sarkar, A., & Paul, B. (2017). Corrigendum to “The global menace of arsenic and its conventional remediation—a critical review” [Chemosphere 158 (September) (2016) 3749](S004565351630683X) (10.1016/j.chemosphere.2016.05.043). Chemosphere. Available from https://doi.org/10.1016/j. chemosphere.2017.01.076. Sen, M., Manna, A., & Pal, P. (2010). Removal of arsenic from contaminated groundwater by membraneintegrated hybrid treatment system. Journal of Membrane Science. Available from https://doi.org/10.1016/j. memsci.2010.02.063. Shih, M. C. (2005). An overview of arsenic removal by pressure-driven membrane processes. Desalination, 172(1), 8597. Available from https://doi.org/10.1016/j.desal.2004.07.031. Smith, E., Naidu, R., Alston, A. M., & Donald, L. S. (1998). Arsenic in the soil environment: A review. Advances in Agronomy, 64, 149195. Available from https://doi.org/10.1016/S0065-2113(08)60504-0. Song, S., Lopez-Valdivieso, A., Hernandez-Campos, D. J., Peng, C., Monroy-Fernandez, M. G., & Razo-Soto, I. (2006). Arsenic removal from high-arsenic water by enhanced coagulation with ferric ions and coarse calcite. Water Research, 40(2), 364372. Available from https://doi.org/10.1016/j.watres.2005.09.046. Teychene, B., Collet, G., Gallard, H., & Croue, J. P. (2014). Corrigendum to: “A comparative study of boron and arsenic (III) rejection from brackish water by reverse osmosis membranes. Desalination, 310(2013), 109114. Available from https://doi.org/10.1016/j.desal.2012.05.034. Desalination. https://doi.org/10.1016/ j.desal.2014.10.019. Thirunavukkarasu, O. S., Viraraghavan, T., Subramanian, K. S., & Tanjore, S. (2002). Organic arsenic removal from drinking water. Urban Water, 4(4), 415421. Available from https://doi.org/10.1016/S1462-0758(02) 00029-8. Uddin, M. T., Mozumder, M. S. I., Figoli, A., Islam, M. A., & Drioli, E. (2007). Arsenic removal by conventional and membrane technology: An overview. Indian Journal of Chemical Technology.

Study of transport models for arsenic removal using nanofiltration process 405 Urase, T., Oh, J. I., & Yamamoto, K. (1998). Effect of pH on rejection of different species of arsenic by nanofiltration. Desalination, 117(13), 1118. Available from https://doi.org/10.1016/S0011-9164(98) 00062-9. Vrijenhoek, E. M., & Waypa, J. J. (2000). Arsenic removal from drinking water by a “loose” nanofiltration membrane. Desalination, 130(3), 265277. Available from https://doi.org/10.1016/S0011-9164(00)00091-6. Xia, S., Dong, B., Zhang, Q., Xu, B., Gao, N., & Causseranda, C. (2007). Study of arsenic removal by nanofiltration and its application in China. Desalination, 204(13 SPEC. ISS), 374379. Available from https://doi.org/10.1016/j.desal.2006.04.035. Yoon S.-H. (2016). Classification of Membranes According to Pore Size. Retrieved from http://onlinembr.info/ membrane-process/classification-of-membranes-according-to-pore-size/

Further Reading Tahura, S., Shahidullah, S. M., Rahman, T., & Milton, A. H. (1998). Evaluation of an arsenic removal household device: Bucket Treatment Unit (BTU) (pp. 158170). Retrieved from http://archive.unu.edu/ env/Arsenic/Tahura.pdf.

CHAPTER 18

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms ¨rstmann, J. Peens and H.G. Brink B. van Veenhuyzen, C. Ho Department of Chemical Engineering, Faculty of Engineering, Built Environment and Information Technology, University of Pretoria, Hatfield, Pretoria, South Africa

18.1 Introduction Five million tons of lead (Pb) ore are mined annually to be refined for use in a variety of industrial applications (International Lead Association, 2019). The most common application of Pb is in the manufacturing of Pb-acid batteries for energy storage. These batteries are mostly utilized in automotive applications, as well as emergency power supplies for various critical services such as hospitals, communication networks, public buildings, and emergency services. Additionally, Pb batteries arguably form the backbone of renewable energy industries such as solar photovoltaic implementations, wind turbines, and electric or hybrid vehicles. The extremely high density of Pb provides unrivaled radiation protection which is essential for medical, dental, research, and nuclear installations. Further, Pb is used as a stabilizer to increase the durability of polyvinyl chloride products and has been applied to thousands of kilometers of underwater power and communication cabling (International Lead Association, 2019). Two problems, however, arise from lead industries. Firstly, the present rate of lead extraction considered with the most recent estimate of current lead ore reserves (88 Mt) means that raw lead could potentially be depleted by 2035 (Statista, 2019). Secondly, the mining, processing, and disposal of lead products introduce lead pollutants into the environment. Pb is highly toxic and has been found to accumulate through different trophic levels of ecosystems (Naik, Khanolkar, & Dubey, 2013a, 2013b). Lead serves no biological purpose, but rather harms organisms either directly by damaging cell structure, or indirectly by impairing enzymes and substituting cationic nutrients. In plants, this results in necrosis, the inhibition of growth, and a reduction in biomass (Shakoor et al., 2013). For animals, this harm is seen in the form of damage done to the central nervous system, kidneys, reproductive system, and immune system (United Nations Environment Programme, 2010).

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00018-3 © 2020 Elsevier Inc. All rights reserved.

407

408 Chapter 18 Cases of lead poisoning have been reported in local communities close to lead-acid battery recycling plants, lead transportation routes, and lead ore mines (Shakoor et al., 2013). In humans, exposure to high concentrations of lead has been shown to cause neuronal encephalopathy and gastrointestinal colic, which includes symptoms such as constipation, abdominal pain, and intestinal paralysis (Mudipalli, 2007). The majority of these ill effects are caused by Pb in its ionic or aqueous form, Pb(II). In this oxidation state it is more likely to form organic complexes or attach to colloidal particles, making it exceptionally mobile and bioavailable (Naik et al., 2013a, 2013b). This chapter presents the current research and understanding in terms of Pb(II) immobilization through bioremediative techniques, with specific focus on work done using locally obtained Pb-resistant consortia. The consortia have been shown not only to possess the ability to remove Pb(II) from contaminated water, but also immobilize the Pb(II) in forms that could enhance the recovery for industrial reuse.

18.2 Remediation of Pb 18.2.1 Conventional methods for Pb remediation or recovery Various processes are currently used to remove Pb(II) from water. A summary of the most widely employed techniques is presented in Table 18.1. The general goal of these methods is either to immobilize the Pb(II) and prevent its proliferation in the environment or to transform it into a less harmful state. The common failure of adsorption processes is the lack of selectivity for the target metals. The presence of other minerals such as Na, Ca, K, and Mg can inhibit the adsorption processes due to competition for adsorption sites. In contrast, many of these techniques are favorable due to their high selectivity yet carry the disadvantages of high operating energy requirements and the requirement for additional processes to handle Pb waste after extraction. A large proportion of these methods also prove uneconomical for treating water with low Pb concentrations (Fu & Wang, 2011).

18.2.2 Bioremediation of Pb A promising alternative to the conventional techniques discussed above is bioremediation, where organisms are used to remove or detoxify aqueous Pb (Philp, Barnforth, Singleton, & Atlas, 2005). Bioremediation is attractive due to the variety of biomaterials applicable (such as algae, fungi, plants, and bacteria) and its potential for low cost and high efficiency operation at low Pb concentrations (Kang et al., 2015). The bioremediation of Pb has mostly been limited to sorption with biomass (Chatterjee, Mukherjee, Sarkar, & Roy, 2012). This includes the use of plants for phytoextraction (Shakoor et al., 2013) and fungi for mycelial biosorption (Chakraborty, Mukherjee, & Das, 2013). Some microorganisms have been

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 409 Table 18.1: Current techniques employed for removal of Pb. Treatment method Adsorption

Pressure driven membrane filtration Electrodialysis

Electrochemical treatment

Chemical precipitation Ion exchange

Description

Advantages

Disadvantages

Aqueous Pb is transferred from the liquid phase to the surface of a solid Polluted water is forced through pores with sizes ranging from ultra- to nanofiltration and reverse osmosis An electric field draws Pb ions out of solution across a charged membrane Aqueous Pb is reduced to metallic Pb by means of an applied current and deposited onto an electrode Pb ions are reacted with various reagents to form an insoluble precipitate Resin exchanges embedded cations for aqueous Pb

Low selectivity Low capital cost, adsorbent regeneration is viable High efficiency and High operating and selectivity maintenance costs

High selectivity

High operating cost

Recovery of elemental Pb

Large capital investment and high operating cost

Low capital cost, simple operation

Sludge generation

High treatment capacity and efficiency. Process kinetics are rapid

Sensitivity to operating conditions such as pH, temperature, concentration, and contact time

Source: From Fu, F., & Wang, Q. (2011). Removal of heavy metal ions from wastewaters: A review. Journal of Environmental Management, 92(3), 407 418. Available from https://doi.org/10.1016/j.jenvman.2010.11.011.

discovered that reduce the bioavailability and toxicity of Pb by precipitating it out as an insoluble complex. Staphylococcus aureus, for example, has been found to take up Pb(II) and precipitate it out as Pb3n(PO4)2n crystals (Levinson, Mahler, Blackwelder, & Hood, 1996; Naik et al., 2013a, 2013b). Additionally, several species of Pb-resistant marine bacteria tend to form insoluble PbS (De, Ramaiah, & Vardanyan, 2008) when exposed to Pb(II). An overview of known microorganisms or biomass studied for the removal of aqueous Pb from solution is given in Table 18.2.

18.2.3 Mechanisms of bioremediation Due to the harmful effects of Pb(II), microorganisms have adapted various defense mechanisms to detoxify or immobilize this pollutant (Kessi, Ramuz, Wehrli, Spycher, & Bachofen, 1999). A variety of mechanisms have been observed in different microorganisms and are presented in Table 18.3.

410 Chapter 18 Table 18.2: Summary of research on species used for Pb(II) bioremediation and their performance. Microbe or material used in bioremediation process Bacillus sp. Pseudomonas sp. Corynebacterium sp. Staphylococcus sp. Escherichia coli Rhodopseudomonas palustris Adsorption onto microparticles of dry plants Enterobacter cloacae Adsorption on activated carbon developed from Tamarind wood Adsorption onto activated Cassia grandis seed gum Sulfate-reducing consortium Sulfate-reducing bacteria (continuous operation) Enterobacter sp.

Pb Time (ppm) period

Pb removal (%) References

450

11 h

100 100

4days 30 min

90 88 87 65 60 96 97

7 50

48 h 50 min

68 94

200

16 h

60

80

7 days

98

50

30 min

90

1000

24 h

90

Kafilzadeh, Afrough, Johari, and Tahery (2012)

Sinha and Biswas (2014) Benhima, Chiban, Sinan, Seta, and Persin (2008) Kang et al. (2015) Acharya, Sahu, Mohanty, and Meikap (2009) Singh, Tiwari, Sharma, and Sanghi (2007) Verma, Bishnoi, and Gupta (2017) Hien Hoa and Liamleam (2007) Jiang et al. (2019)

While the biological reduction of metal species to lower oxidation states is thermodynamically feasible, the requirement for large activation energy can potentially render these processes unfeasible for environmental applications. Biological systems utilize enzymes as biocatalysts which reduce the activation energy requirements, thereby facilitating ambient temperature reactions (Karp, 2009). Several bacterial cultures have demonstrated the ability to biologically reduce metal pollutants into less toxic oxidation states, namely Cr(VI) to Cr(III) (Molokwane & Chirwa, 2009), U(VI) to U(IV) (Chabalala & Chirwa, 2010), and Se(VI) to Se(0) (Li et al., 2014). Despite the energetic barrier, in each case microorganisms were discovered that facilitated the biotransformation of the metals to the lower oxidation state. The bioreduction of aqueous Pb(II) to its more stable elemental state (Pb0) is still relatively unexplored. The first reported attempt at biological reduction of Pb(II) to Pb0 using an anaerobic culture of Moraxella bovis reported a dark precipitate of unknown identity (Saiz & Barton, 1992). In another study, the abiotic reduction of Pb(II) to the metallic state was achieved using a purified cytochrome 3 from Desulfomicrobium baculatum (Abdelouas et al., 1999). In a recent report by Moreno et al. (2019), the successful reduction of Pb(II) to elemental Pb was reported for three Candida species in an oxidizing environment (100 mM H2O2).

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 411 Table 18.3: Biochemical and molecular mechanisms of Pb-resistance in bacteria. Mechanism

Description

Efflux

Pb concentration is regulated by transporter proteins that prevent overaccumulation of highly toxic and reactive heavy metal ions Intracellular Metal-binding proteins facilitate the accumulation of bioaccumulation heavy metals inside the cell and protect metabolic processes catalyzed by enzymes Extracellular sequestration

High molecular weight biopolymers are secreted by bacterial cells to bind and immobilize the metal to counteract the toxic effects

Surface biosorption

The presence of a variety of negatively charged chemical groups on the bacterial cell surface allows adsorption of metal ions Bioprecipitation The microorganism facilitates precipitation of the ionic metal into a less harmful or bioavailable insoluble state Alteration in cell Adjustments in cell size, shape, aggregation, and morphology outer membrane surface area are used to reduce the pollutant impact Metal ligand complex formation

Example species Ralstonia metallidurans CH34 Staphylococcus aureus Oscillatoria brevis (BmtA) Synechococcus PCC 7942 (SmtA) Pseudomonas aeruginosa strain WI-1 (Naik, Pandey, & Dubey, 2012b) Pseudomonas marginalis Paenibacillus jamilae Enterobacter cloacae strain P2B (Naik, Pandey, & Dubey, 2012a) Bacillus subtilis Pseudomonas aeruginosa PU21

Providencia alcalifaciens strain 2EA (Naik et al., 2013a, 2013b) Bacillus iodinium Escherichia cloacae Pseudomonas putida Pseudomonas aeruginosa strain 4EA (Naik & Dubey, 2011) Pseudomonas vesicularis Compounds are secreted by microorganisms for Streptomyces sp. transporting iron which ends up forming stable metal ligand complexes with the Pb, reducing Pseudomonas aeruginosa strain 4EA (Naik & Dubey, 2011) the bioavailability and mobility of the pollutant

Source: Based on Naik, M. M., Khanolkar, D., & Dubey, S. (2013a). Lead resistant bacteria: Lead resistance mechanisms, their applications in lead bioremediation and biomonitoring. Ecotoxicology and Environmental Safety, 98, 1 7. Available from https://doi.org/10.1016/j. ecoenv.2013.09.039; Naik, M. M., Khanolkar, D., & Dubey, S. K. (2013b). Lead-resistant Providencia alcalifaciens strain 2EA bioprecipitates Pb 1 2 as lead phosphate. Letters in Applied Microbiology, 56(2), 99 104. Available from https://doi.org/10.1111/lam.12026.

18.3 Case study of aqueous Pb biorecovery by local Pb-resistant organisms This section presents results from laboratory-scale batch reactors as studied by Peens (2018). These reactors were inoculated with local, Pb-resistant organisms known to precipitate Pb(II) from aqueous solution (Brink, Lategan, & Chirwa, 2017). Their study allowed for a comprehensive understanding of how the Pb remediation process works and where future research is required.

18.3.1 Characterization of bacterial consortia The bacteria for this study were sourced from soil at a Pb battery recycling plant (referred to as the B-culture) and a Pb ore mine (referred to as the S-culture) in South Africa. It

412 Chapter 18 was hypothesized that these cultures would be well adapted to survive in high Pb concentration environments and possess a mechanism capable of reducing Pb toxicity or mobility. Such characteristics can be taken advantage of to reduce the Pb(II) concentration of a solution. The identification of the bacterial consortia is motivated by several benefits. Importantly, it provides access to prior research done on specific species and contributes to the general understanding of bioremediation mechanisms. Two analyses were done to classify the consortia. Firstly, streak plates were prepared for both S- and B-cultures at 80 and 500 ppm Pb(II). The entire culture of the streak plates was sequenced and then analyzed using the Basic Local Alignment Search Tool to compare genetic information with known microbial species. Secondly, spread plates of the consortia were prepared for both cultures at 80 and 500 ppm Pb(II). Colonies exhibiting precipitate formation were selected from the spread plates and isolated before being sequenced and analyzed using 16s rRNA genetic fingerprinting analysis. The streak plate analyses were able to confirm the presence and relative abundance of a consortia of bacteria while the isolates from spread plates were able to identify dominant bacteria. For the B-culture, the streak plates showed that at 80 ppm Pb(II) Enterococcus sp. and Clostridium botulinum were present in the largest quantity, but at 500 ppm Pb(II) Ralstonia solanacearum and Klebsiella pneumoniae were the most abundant. Furthermore, K. pneumoniae was identified as the dominant bacteria in spread plate isolates from both 80 and 500 ppm Pb(II). The S-culture results showed Clostridium bifermentans, K. pneumoniae, and Listeria monocytogenes as having major populations present at 80 ppm for the streak plate analysis. For the same analysis at 500 ppm, Enterococcus faecium was identified as the dominant bacteria with an insignificant K. pneumoniae population. This suggests that E. faecium outcompetes K. pneumoniae in terms of growth at higher concentrations of Pb(II). The spread plates analyzed for the S-culture yielded E. faecium as the dominant bacterium at both 80 ppm and 500 ppm Pb(II).

18.3.2 Precipitate identification To better understand the processes and verify the possibility of recovering elemental Pb through bioremediation, the precipitates formed by the S- and B-cultures were analyzed. Four different sets of batch reactors were operated for the Pb-precipitate analyses. Batch reactors were set up for 80 and 500 ppm Pb. The 80 ppm batch reactors were made up of 25 g/L standard Miller lysogeny broth (LB) and spiked with Pb(NO3)2. For the 500 ppm batch reactors, a simulated LB broth was used consisting of 20 g/L tryptone, 10 g/L yeast extract, and only 1 g/L of NaCl. This mixture possesses twice the concentration of nutrients when compared to the 80 ppm reactors to accommodate the high Pb(II) concentration, but only contains one tenth of the NaCl to prevent the formation of PbCl2 precipitate. The

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 413 reactors were operated anaerobically by purging with N2 before being placed on a shaker at 150 rpm and 35 C. The batch reactors were incubated for at least two weeks to allow for maximum precipitate formation. An analysis of the precipitate was undertaken using X-ray photoelectron spectroscopy and a scanning electron microscope (SEM) coupled with energy dispersive X-ray spectroscopy. The results identified the precipitation of PbS and PbO from solution. It was also found that the two consortia produced precipitates of equivalent composition under the same reactor operating conditions. The most probable source of sulfur is from two amino acids found in yeast extract and tryptone, namely cysteine and methionine (GRiSP Research Solutions, 2016a, 2016b). Several possible mechanisms are deemed possible for the liberation of sulfur from these two amino acids in the reactor (Peens, 2018). PbS was found to be the most abundant precipitate. It was discovered that PbS made up approximately 80% of the precipitated Pb when 80 ppm Pb(II) was used and around 40% with 500 ppm Pb(II) (Peens, Wu, & Brink, 2018). The presence of PbO was attributed to the oxidation of Pb(0) during the drying of the samples for analysis. This has been confirmed by similar experiments (Brink, Ho¨rstmann, & Feucht, 2019). More elemental Pb is formed at elevated Pb(II) concentrations in the media, suggesting that Pb(II) is more likely to be used as an electron acceptor than nitrates or sulfates, especially if fewer sulfur compounds are present under anaerobic respiration. A precipitate of PbCl2 has also been shown to form at high Pb concentrations where the solubility product is exceeded. The salt, however, has been found to dissolve again once enough Pb(II) has precipitated out of solution (Peens et al., 2018). In addition, the SEM afforded insight into the mechanism. As seen in Fig. 18.1, a precipitate appears to blanket the microbe. This is indicative of an extracellular precipitation to detoxify Pb(II).

18.3.3 Microbiological and kinetic study This section of the case study focuses on investigating and measuring how the behavior of the microbes changed as a function of time. 18.3.3.1 Laboratory-scale system design and experimental approach While only the B-culture was employed for this study, the two consortia do not show significant variations in results (Brink et al., 2017). This is most likely because these organisms have convergently adapted to similar environments. Two experiments were performed in parallel for this investigation using the same microbial culture. The first experiment, named B80, was completed in triplicate batch reactors at a

414 Chapter 18

Figure 18.1 Scanning electron microscope (SEM) micrograph of the precipitate and biomass of S80. Source: From Peens, J. (2018). Pb(II)-removal from water using microorganisms naturally evolved to tolerate Pb(II)-toxicity. Pretoria: University of Pretoria.

concentration of 80 ppm Pb(II). Each batch reactor was set up by using a 100 mL serum bottle. Firstly, stock solutions of all reagents were prepared and the growth media and Pb (II) stock solutions were autoclaved separately. For B80, the standard Miller LB broth was used with a final concentration of 25 g/L. The Pb(II) stock solution was prepared with Pb (NO3)2. After autoclaving the relevant solutions and glassware, the growth media and Pb(II) stock solutions were added to each batch reactor, working in a biological safety cabinet and near an open flame to ensure sterile conditions. The serum bottles were inoculated, sealed with parafilm, purged with N2-gas for 3 min, sealed with a rubber stopper, and clamped with a metal cap to ensure anaerobic conditions. Finally, the batch reactors were placed on a shaker at 150 rpm and incubated at 35 C for the 15-day duration of the experiment. 18.3.3.2 Laboratory-scale system results and discussion The microbiological activity and kinetics will be discussed in terms of four prominent variables measured from samples over the duration of the case study. These are the colony forming unit (CFU) count in Fig. 18.2, Pb(II) concentration in Fig. 18.3, metabolic activity (MA) in Fig. 18.4, nitrate concentration in Fig. 18.5, and Pb concentration in Fig. 18.6. As shown in the figures, three distinct stages in reactor operation were identified and are highlighted in different colors. A fourth stage is identified in MA. Abiotic controls were able to confirm that the precipitation of Pb was not caused by physical or chemical means, but rather by biological processes (Fig. 18.3).

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 415

Figure 18.2 Colony forming unit (CFU) count rate distribution. Source: From Peens, J. (2018). Pb(II)-Removal from water using microorganisms naturally evolved to tolerate Pb(II)-toxicity. Pretoria: University of Pretoria.

Figure 18.3 Rate distribution Pb concentrations over time. Source: From Peens, J. (2018). Pb(II)-Removal from water using microorganisms naturally evolved to tolerate Pb(II)-toxicity. Pretoria: University of Pretoria.

The first stage spans the initial two days. From Figs. 18.2 and 18.4 it can be seen that this first phase was characterized by a sudden increase in MA and CFU count. The behavior of these two variables is indicative of rapid substrate metabolism as biomass was produced. During anaerobic respiration, K. pneumoniae uses NO2 3 as the terminal electron acceptor instead of oxygen and

416 Chapter 18

Figure 18.4 Metabolic activity (MA) rate distribution over time. Source: From Peens, J. (2018). Pb(II)-Removal from water using microorganisms naturally evolved to tolerate Pb(II)-toxicity. Pretoria: University of Pretoria.

Figure 18.5 Nitrate rate distribution over time. Source: From Peens, J. (2018). Pb(II)-Removal from water using microorganisms naturally evolved to tolerate Pb(II)-toxicity. Pretoria: University of Pretoria. 2 reduces NO2 3 to NH3. The result of this can be seen in Fig. 18.5 as NO3 concentration fell steeply during the first phase. Sulfide was released in this phase by the catabolism of methionine and cysteine. This is supported by the visible change in the appearance of the reactor to dark gray/black as PbS began to precipitate. In Fig. 18.3, the sharp drop observed in Pb(II) concentration during the first phase also suggests that this may be the case.

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 417 The second stage spans days 2 3. Here, the drastic decreases in MA and CFU counts show a transition phase as the cells switch from growth to maintenance. Between days 3 and 15, the third stage is observed where activity stabilizes across all variables. The maintenance phase of the microbes is evidenced by steady values of biomass and MA with gradual cell death. This phase also exhibits the use of less energy than required for growth in order to regenerate damaged biomass, maintain gradients, and maintain electric potential (Villadsen, Nielsen, & Lide´n, 2011). After the depletion of nitrate, it is likely that the microbes began using Pb(II) as an electron acceptor in stage 3. The reason Pb(II) reduction becomes prominent in this stage is most probably due to the fact that nitrate has more energy available as an electron acceptor (Wielinga, 2009) and was therefore preferred in the first phase. The analysis of the precipitate supports this idea by showing that the fraction of Pb reduced to Pb(0) corresponds with the fraction of Pb(II) removed during the third phase. Although not shown in the figures, the study did also find an increase in aqueous nitrogen concentration toward the end of the run. This may be attributed to nitrogen fixation by K. pneumoniae (Klopprogge & Schmitz, 1999), which is repressed by the presence of O2 and NO2 3 (Hom et al., 1980). The breakdown of biomass may also be the cause of nitrogen released into the reactor medium; this is supported by the drop in CFU count and MA as time progresses in the third stage. Stage 4 presents as a noticeable dip in MA, most likely brought about by cell death.

18.3.4 Case studies of varied operating conditions In this section, a variety of case studies are considered in terms of the effects of their respective operating conditions on the reactor performance. 18.3.4.1 Effect of elevated heavy metal concentrations In the study presented earlier by Peens (2018), parallel experimental runs were performed at Pb(II) concentrations of 80 and 500 ppm to study the effect of varying Pb(II) concentrations on the reactor performance. Since these reactors contained a higher concentration of Pb(II), a simulated LB broth was made up that contained double the amount of nutrients and less NaCl to stimulate growth while preventing the formation of PbCl2. The final concentration of each constituent was as follows: 20 g/L tryptone, 10 g/L yeast extract, and 1 g/L of NaCl. Experimental results showed very similar behavior in 80 and 500 ppm batches. The extra nutrients allowed for higher CFU counts and higher MA per CFU than with the 80 ppm reactor. This growth may suggest that the higher Pb concentration has a limited inhibitory effect, but it is worth noting that the 80 ppm reactor took a day less to form the dark precipitate and reach the end of the first phase. After running for 15 days, analysis from atomic adsorption spectroscopy showed 90% of bioavailable Pb removed from the 80 ppm

418 Chapter 18

Figure 18.6 Measured Pb concentrations in the aqueous and precipitate phases for reactors with initial loadings of (A) 80 ppm Pb(II) and (B) 500 ppm Pb(II). Source: From Peens, J. (2018). Pb(II)Removal from water using microorganisms naturally evolved to tolerate Pb(II)-toxicity. Pretoria: University of Pretoria.

reactor and 72% removed from 500 ppm reactor (Fig. 18.6). Even though Pb removal occurred more rapidly in the 80 ppm reactor, more Pb was removed in total from 500 ppm reactor. This displays the capacity for Pb removal at higher concentrations. In another study on the B-consortium obtained from the Pb battery recycling plant, the effect of elevated Pb(II) concentrations was evaluated (Peens et al., 2018). Experiments were performed with concentrations of Pb(II) ranging from 80 to 1000 ppm. The consortium was shown to remove Pb(II) at concentrations as high as 1000 ppm (Fig. 18.7). The experimental results also showed a nonlinear relationship between initial Pb concentration and percentage removal, as indicated by Table 18.4. 18.3.4.2 Effect of Zn(II) or Cu(II) ions on Pb(II) bioprecipitation Due to Pb being commonly found with zinc and copper, a study was done to determine the effect of these other metals on the bioprecipitation mechanism of the B-culture consortium (Ho¨rstmann & Brink, 2019). It was found that both Cu(II) and Zn(II) tend to inhibit the removal of Pb(II). In the results of this study, the percentage of Cu(II) removed from solution in conjunction with Pb was found to be greater than the percentage of Pb removed. This suggests a competitive remediation mechanism for the two metals. The precipitation of Pb was not observed in reactors with a concentration of 80 ppm Pb(II) and 80 ppm Zn(II),

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 419

Figure 18.7 Change in Pb(II) concentration measurements of the (A) 25 g/L and (B) 100 g/L concentration simulated lysogeny broth (LB). Source: From Peens, J., Wu, W. Y., & Brink, H. G. (2018). Microbial Pb(II) precipitation: The influence of elevated Pb(II) concentrations. Chemical Engineering Transactions, 64, 439 444. Available from https://doi.org/10.3303/CET1757069. Table 18.4: Relationship between Pb concentration and removal efficiency for B-culture bacteria with a substrate of 24.5 g/L tryptone, 24.5 g/L yeast extract and 1 g/L NaCl. Initial Pb(II) concentration (ppm) 80 250 500 750 1000

Pb(II) removal after 11 12 days (%) 82.8 90.2 99.2 61.0 55.3

Source: From Peens, J., Wu, W. Y., & Brink, H. G. (2018). Microbial Pb(II) precipitation: The influence of elevated Pb(II) concentrations. Chemical Engineering Transactions, 64, 439 444. Available from https://doi.org/10.3303/CET1757069.

even though a small amount of Pb was removed from solution. This removal is most likely due to biosorption. Precipitate did, however, form for a concentration of 80 ppm Pb(II) in conjunction with 40 ppm Zn(II). These results demonstrated the inhibitory effect of Zn(II) and Cu(II) on the bioprecipitation of Pb by the industrial consortium, implying the requirement for Cu(II) and Zn(II) removal prior to Pb(II) biorecovery using the specific consortium. 18.3.4.3 The minimum inhibitory concentration of Pb(II) for the B-consortium In a study by Brink et al. (2019) the minimum inhibitory concentrations (MICs) at which B-culture consortium would cease to grow was determined. A simulated nutrient agar with reduced NaCl concentration was prepared and inoculated with the B-culture. A heated glass rod was used to drill 10 mm wells in the agar in which Pb(II) solutions of different

420 Chapter 18 concentrations were poured. The MIC was deemed to be that concentration of Pb(II) which would prevent observable microbial growth in the agar. The experimental results showed the MIC to correspond to approximately 30,000 ppm Pb(II). Interestingly, runs without added NaCl showed inhibitory concentrations of approximately 8000 ppm lower than runs with NaCl added. This suggests an osmotic shielding mechanism provided by NaCl that protects the microbe during growth. Previous results by the same research group support this observation as an improved removal of Pb in the presence of NaCl was observed (Brink & Mahlangu, 2017). 18.3.4.4 Effect of substrate composition Experimental results from Peens et al. (2018) have shown that growth substrates play a significant role in the rate and capacity of the B-consortium to precipitate Pb(II). Fig. 18.7 shows a study where bacterial consortia were employed in reactors of varying Pb(II) concentrations and either 25 or 100 g/L simulated LB broth. The reactors of higher substrate concentrations show rapid removal of Pb(II) from solution when compared to the more stagnant concentrations seen in the reactors with less substrate. On day 16 of runs with high substrate concentration, the 750 and 1000 ppm reactors were spiked with more LB broth to determine whether nutrients were the limiting factor in Pb precipitation. Fig. 18.6 illustrated how this spike reinvigorated the removal of Pb(II) from solution. Substrate limitations affecting the bioremediation by the B-consortium were also observed by Brink et al. (2017). In these experimental runs, reactors were spiked with additional LB broth after 96 h to determine whether more Pb removal was possible and whether low removal rates were the result of substrate limitation. The spike managed to allow for significantly more Pb removal and showed substrate to be the limiting factor. This same study also demonstrated how the addition of glucose (60 g/L) significantly inhibits Pb removal and microbial growth. In another study, the effects of different types of substrates were investigated (Brink & Mahlangu, 2017). The investigation indicated that any glucose- and xylose-supplementation of the fermentation medium allows for the removal of Pb from solution (up to 75% 90% in 48 72 h) but tends to inhibit the bioprecipitation of Pb(II). This is most likely a result of the sugar-promoting fermentative organisms in the consortium that compete for nutritional resources but do not aid in the removal of Pb(II) from solution. This was supported by the observed reduction of the pH from above 6 to approximately 4.5 in both sugar supplemented runs. Additionally, the study tested the effectiveness of various combinations of LB broth constituents (yeast extract, tryptone, and NaCl) along with corn steep liquor. The results showed that while LB broth performed the best (89% Pb removal), the main constituent in LB broth required by the Pb-reducing population is yeast extract with NaCl as a

Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 421 supplementary nutrient (80% removal). Yeast extract alone only removed 56% Pb. It was also shown that clarified corn steep liquor used to replace the tryptone in LB broth does not improve the final removal of Pb when compared to the yeast extract NaCl medium (80% removal). No microbial growth occurred in a tryptone NaCl medium.

18.4 Conclusion and outlook Considering the current estimated global lead reserves as well as the historical and current nonlocalized outputs of lead to the environment, it has become imperative that an environmentally sustainable method for the remediation and recovery of lead be developed. As the state of knowledge on locally isolated industrial consortia develops, the potential for the use of these local lead-resistant organisms for the bioremediation and biorecovery of lead becomes more apparent. The current research has shown that the microorganisms used are extremely resilient to industrially significant concentrations of Pb (II) in solution. This was supplemented by significant Pb(II) precipitation potential demonstrated under batch conditions, usually only limited by the availability of sufficient substrate and nutrients to the organisms. Currently, the most prominent limitations observed are the requirements for a complex feed (mostly containing yeast extract), and substantial inhibition by Zn(II) and Cu(II), that is, those metals most commonly found in Pb(II)-containing ores. Future anticipated endeavors in this project include: • • •

•

• •

Exploring the biochemical lead precipitation and reduction kinetics of the consortia in greater detail to provide more insight into the lead removal mechanisms. Studying of the biochemical electron transport mechanism responsible for the reduction of Pb(II) to elemental lead to elucidate the enzymatic processes responsible. Comparing the bioremediation effectiveness of the consortia to the isolated microbial strains constituting the consortia to determine the respective contributions of the strains to the system. Researching and developing a more economically suitable substrate. For example, unpublished work on yeast extract from spent brewer’s yeast has found it to be a promising alternative to LB broth. Substrates can also be developed to minimize their sulfur content to ensure that PbS precipitate is limited and the production of elemental lead is maximized. Investigating methods of recovering Pb(0) from PbS to improve the economic viability of this form of remediation, since the majority product during bioprecipitation is PbS. Considering designs for the operation of this method in a continuous steady-state reactor such as the upflow anaerobic sludge blanket reactor or anaerobic hybrid reactor to further the investigation of this method for industrial application.

422 Chapter 18

Acknowledgment This work is based on the research supported in part by the National Research Foundation of South Africa for the grant, Unique Grant No. 106938.

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Bioremediation and biorecovery of aqueous lead by local lead-resistant organisms 423 International Lead Association. (2019). Lead production & statistics. ,https://www.ila-lead.org/lead-facts/leadproduction--statistics. Accessed 19.06.23. Jiang, Z., Jiang, L., Zhang, L., Su, M., Tian, D., Wang, T., . . . Li, Z. (2019). Contrasting the Pb(II) and Cd(II) tolerance of Enterobacter sp. via its cellular stress responses. Environmental Microbiology. Available from https://doi.org/10.1111/1462-2920.14719. Kafilzadeh, F., Afrough, R., Johari, H., & Tahery, Y. (2012). Range determination for resistance/tolerance and growth kinetic of indigenous bacteria isolated from lead contaminated soils near gas stations (Iran). European Journal of Experimental Biology, 2(1), 62 69. Kang, C. H., Oh, S. J., Shin, Y., Han, S. H., Nam, I. H., & So, J. S. (2015). Bioremediation of lead by ureolytic bacteria isolated from soil at abandoned metal mines in South Korea. Ecological Engineering, 402 407. Available from https://doi.org/10.1016/j.ecoleng.2014.10.009. Karp, G. (2009). Cell and molecular biology: Concepts and experiments. Hoboken, NJ: John Wiley & Sons. Kessi, J., Ramuz, M., Wehrli, E., Spycher, M., & Bachofen, R. (1999). Reduction of selenite and detoxification of elemental selenium by the phototrophic bacterium Rhodospirillum rubrum. Applied and Environmental Microbiology, 65(11), 4734 4740. Klopprogge, K., & Schmitz, R. A. (1999). NifL of Klebsiella pneumoniae: redox characterization in relation to the nitrogen source (2)Biochimica et Biophysica Acta (1431, pp. 462 470). Elsevier. Available from https://doi.org/10.1016/S0167-4838(99)00075-8. Levinson, H., Mahler, I., Blackwelder, P., & Hood, T. (1996). Lead resistance and sensitivity in Staphylococcus aureus. FEMS Microbiology Letters, 145, 421 425. Li, B., Liu, N., Li, Y., Jing, W., Fan, J., Li, D., . . . Wang, L. (2014). Reduction of selenite to red elemental selenium by Rhodopseudomonas palustris Strain N. PLoS One, 9(4). Available from https://doi.org/ 10.1371/journal.pone.0095955. Molokwane, P., & Chirwa, E. (2009). Cr(VI) reduction in packed-column microcosm reactors using chromium. Chemical Engineering Transactions, 18, 863 868. Available from https://doi.org/10.3303/CET0918141. Moreno, A., Demitri, N., Ruiz-Baca, E., Vega-Gonza´lez, A., Polentarutti, M., & Cue´llar-Cruz, M. (2019). Bioreduction of precious and heavy metals by Candida species under oxidative stress conditions. Microbial Biotechnology. Available from https://doi.org/10.1111/1751-7915.13364. Mudipalli, A. (2007). Lead hepatotoxicity & potential health effects. Indian Journal of Medical Research, 126 (6), 518 527. Naik, M. M., & Dubey, S. K. (2011). Lead-enhanced siderophore production and alteration in cell morphology in a Pb-resistant Pseudomonas aeruginosa strain 4EA. Current Microbiology, 62(2), 409 414. Available from https://doi.org/10.1007/s00284-010-9722-2. Naik, M. M., Khanolkar, D., & Dubey, S. (2013a). Lead resistant bacteria: Lead resistance mechanisms, their applications in lead bioremediation and biomonitoring. Ecotoxicology and Environmental Safety, 98, 1 7. Available from https://doi.org/10.1016/j.ecoenv.2013.09.039. Naik, M. M., Khanolkar, D., & Dubey, S. K. (2013b). Lead-resistant Providencia alcalifaciens strain 2EA bioprecipitates Pb 1 2 as lead phosphate. Letters in Applied Microbiology, 56(2), 99 104. Available from https://doi.org/10.1111/lam.12026. Naik, M. M., Pandey, A., & Dubey, S. K. (2012a). Biological characterization of lead-enhanced exopolysaccharide produced by a lead resistant Enterobacter cloacae strain P2B. Biodegradation, 23(5), 775 783. Available from https://doi.org/10.1007/s10532-012-9552-y. Naik, M. M., Pandey, A., & Dubey, S. K. (2012b). Pseudomonas aeruginosa strain WI-1 from Mandovi estuary possesses metallothionein to alleviate lead toxicity and promotes plant growth. Ecotoxicology and Environmental Safety, 79, 129 133. Available from https://doi.org/10.1016/j.ecoenv.2011.12.015. Peens, J. (2018). Pb(II)-removal from water using microorganisms naturally evolved to tolerate Pb(II)-toxicity. Pretoria: University of Pretoria. Peens, J., Wu, W. Y., & Brink, H. G. (2018). Microbial Pb(II) precipitation: The influence of elevated Pb(II) concentrations. Chemical Engineering Transactions, 64(2018), 439 444. Available from https://doi.org/ 10.3303/CET1864074.

424 Chapter 18 Philp, J. C., Barnforth, S. M., Singleton, I., & Atlas, R. M. (2005). Environmental pollution and restoration: A role for bioremediation. Bioremediation applied microbial solutions for real-world environmental cleanup (pp. 1 24). Washington, D.C.: ASM Press. Saiz B.L., Barton L.L. (1992). Transformation of Pb(II) to lead colloid using Moraxella bovis. American Society of Microbiology Meeting, Abstract No. 347. Shakoor, M., Ali, S., Farid, M., Farooq, M., Tauqeer, H., Iftikhar, U., . . . Bharwana, S. (2013). Heavy metal pollution, a global problem and its remediation by chemically enhanced phytoremediation: A review. Journal of Biodiversity and Environmental Sciences, 3(3), 12 20. Singh, V., Tiwari, S., Sharma, A. K., & Sanghi, R. (2007). Removal of lead from aqueous solutions using Cassia grandis seed gum-graft-poly(methylmethacrylate). Journal of Colloid and Interface Science, 316(2), 224 232. Available from https://doi.org/10.1016/j.jcis.2007.07.061. Sinha, S. N., & Biswas, K. (2014). Bioremediation of lead from river water through lead-resistant purplenonsulfur bacteria. International Research Publication House, 2(1), 11 14. Statista. (2019). Lead reserves worldwide by country 2018 (in million metric tons). ,https://www.statista.com/ statistics/273652/global-lead-reserves-by-selected-countries/. Accessed 19.06.23. United Nations Environment Programme. (2010). Final review of scientific information on lead. UNEP-DTIE Chemicals Branch. Verma, A., Bishnoi, N. R., & Gupta, A. (2017). Optimization study for Pb(II) and COD sequestration by consortium of sulphate-reducing bacteria. Applied Water Science, 7(5), 2309 2320. Available from https:// doi.org/10.1007/s13201-016-0402-7. Villadsen, J., Nielsen, J., & Lide´n, G. (2011). Bioreaction engineering principles (3rd ed.). New York: Springer. Wielinga, B. (2009). In situ bioremediation of pit lakes. In D. N. Castendyk, & L. E. Eary (Eds.), Mine Pit Lakes: Characteristics, Predictive Modeling and Sustainability (vol. 3). Littleton (pp. 215 223). Colorado: Society for Mining, Metallurgy & Exploration.

CHAPTER 19

Microbial bioremediation of azo dye through microbiological approach ˜ a2 Celia Vargas-de la Cruz1 and Daniela Landa-Acun 1

Latin American Center for Teaching and Research in food bacteriology (CLEIBA), Faculty of Pharmacy and Biochemistry, Major National University of San Marcos (UNMSM), Lima, Peru, 2 Laboratory of Microbial Ecology and Biotechnology, Department of Biology, Faculty of Sciences, National Agrarian University La Molina (UNALM), Lima, Peru

19.1 Introduction Microorganisms are widely distributed in the biosphere because their metabolic capacity is very impressive and they can easily grow in a wide range of environmental conditions. The nutritional versatility of microorganisms can also be exploited for the biodegradation of contaminants. This type of process is called bioremediation. It is continued through the ability of certain microorganisms to convert, modify, and use toxic pollutants to obtain the production of energy and biomass in the process (Das, Charumathi, & Das, 2010; Gonen & Aksu, 2009; Ponraj, Jamunarani, & Zambare, 2011). The quality of life on Earth is inextricably linked to the global quality of the environment. Unfortunately, advances in science, technology, and industry have led to large amounts of pollutants, ranging from dirty wastewater to nuclear waste, being extracted or introduced into the ecosystem, which is a serious problem for the survival of humanity itself on Earth. In the past, waste was traditionally disposed of by excavating a hole and filling it with waste. This way of eliminating waste was difficult to sustain because of the need for a new site every time. The new waste disposal technologies that use high-temperature incineration and chemical breakdown (e.g., catalytic base deformation, ultraviolet oxidation) have evolved. Although they can be very effective in reducing a wide range of pollutants, at the same time they have several disadvantages. These methods are complex, noneconomic, and lack public acceptance. The associated deficiencies in these methods have focused efforts to take advantage of the current bioremediation process as an appropriate alternative. Bioremediation is a transformation or degradation mediated by microorganisms of contaminants in nondangerous or less dangerous substances. Studies have been undertaken

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00019-5 © 2020 Elsevier Inc. All rights reserved.

425

426 Chapter 19 on the employability of various organisms such as bacteria, fungi, algae, and plants for the efficient bioremediation of contaminants (Charumathi & Nilanjana, 2010; Churchley, 1994; Khouni, Marrot, & Amar, 2012). The implication of plants in the bioremediation of pollutants is called phytoremediation. The phytoremediation process is an emerging green technology that facilitates the elimination or degradation of toxic chemicals in soils, sediments, groundwater, surface water, and air. Genetically engineered plants are also used. For example, arsenic is phytogenic for genetically modified plants such as Arabidopsis thaliana, which expresses two bacterial genes. One of these genes allows the plant to modify the arsenide to arsenite and the second modifies arsenite and stores it in vacuoles (Ogugbue, Sawidis, & Oranusi, 2012). Instead of simply collecting the pollutant and storing it, bioremediation is a well-organized microbiological procedure activity that is applied to break or transform pollutants into elemental and compound forms that are less toxic or nontoxic. Bioremediants are biological agents used for bioremediation to clean contaminated sites. Bacteria, archaea, and fungi are typical biological agents (Machado, Compart, Morais, Rosa, & Santos, 2006). The application of bioremediation is a biotechnological process that involves microorganisms to solve and eliminate the environmental dangers of many contaminants through biodegradation. The terms bioremediation and biodegradation are interchangeable. Microorganisms act as important tools for eliminating contaminants in the soil, water, and sediments; mainly due to their advantage over other protocols of remediation procedures. Microorganisms are recovering the original natural environment and avoiding contamination (Saratale, Saratale, Chang, & Govindwar, 2009). The objective of this chapter is to express the current trends in the application/role of microorganisms in bioremediation and to introduce the relevant findings to identify the gaps in this thematic area. Currently, it is an intense research area because microorganisms show promise of containing a variety of valuable genetic material that could solve environmental threats.

19.2 Classification of dyes All aromatic compounds absorb electromagnetic energy, but only those that absorb light with wavelengths in the visible range (350
700 nm) are colored. Toners contain chromophores, delocalized electron systems with conjugated double bonds and auxochromes, electron-withdrawing or electron-donating substituents that cause or enhance chromophore color by changing the total energy of the electron system. Conventional chromophores include
CQC
,
CQN
,
CQO,
NQN
,
NO2, and quinoid rings; conventional auxochromes are
NH3,
COOH,
SO3H, and
OH. Based on chemical structure or chromophore, 20
30 different groups of dyes can be detected. Azo, anthraquinone, phthalocyanine, and triarylmethane dyes are the quantitative major groups. Other groups include diarylmethane, indigo, azin, oxazine, thiazine, xanthan, nitro, nitroso, methine, thiazole, indamine, indophenol, lactone, aminoketone, and hydroxy ketone dyes

Microbial bioremediation of azo dye through microbiological approach 427

Figure 19.1 Chemical structure of C.I. Reactive Red.

and dyes that have an undefined structure. Reactive Red is one of the reactive azo dyes that are used commercially for dyeing textiles. In the Color Index, reactive dyes form the second largest class of dyes with regard to the number of active inputs, about 600 of 1050 different reactive dyes listed are in current production. The reactive dyes are dyes with reactive groups that form covalent bonds with the OH, NH, or SH groups in the fibers. The reactive group is often a substituted heterocyclic aromatic ring with chloride or fluoride, for example, dichlorotriazine. Upon reaction with reactive dyes (Fig. 19.1), hydrolysis of reactive groups is an undesirable side reaction that reduces the degree of fixation. Despite the addition of large amounts of salt and urea (up to 60 and 200 g/L respectively) to increase the degree of fixation, it is estimated that 10% to 50% will not react with the tissue and remain hydrolyzed in the aqueous phase. The problem of colored effluents is therefore mainly identified with the use of reactive dyes. The characteristics and types of dyes have been shown in Table 19.1.

19.3 Role of environmental parameters on microbial biodegradation and bioremediation of azo dye Bioremediation is involved in the degradation, elimination, alteration, immobilization, or detoxification of various chemical products and physical waste from the environment through the action of bacteria, fungi, and plants. Microorganisms are involved through their enzymatic pathways that act as biocatalysts and facilitate the progress of biochemical reactions that degrade the desired pollutant. Microorganisms only act against pollutants when they have access to various materials’ compounds to help them generate energy and nutrients to build more cells. The efficiency of bioremediation depends on many factors; including the chemical nature and the concentration of contaminants, the physicochemical characteristics of the environment, and their availability for microorganisms (Anjaneya, Souche, Santoshkumar, & Karegoudar, 2011; Meng, Liu, Zhou, Fu, & Wanga, 2012; Priya, Uma, Ahamed, Subramanian, & Prabaharan, 2011; Saratale, Saratale, Kalyani, Chang, & Govindwar, 2009). In addition,

428 Chapter 19 Table 19.1: Types of dyes and their characteristics. Dye type

Characteristic

Example

Basic dyes

Cationic compounds used for dyeing fibers containing acid groups, generally synthetic fibers such as modified polyacryl, bind to the fibers of acidic groups The reactive groups form covalent bonds with the OH
, NH
, or SH groups in fibers (cotton, wool, silk, and nylon)
Often a heterocyclic aromatic ring substituted by chloride or fluoride, for example, dichlorotriazine vinyl sulfone (like Reactive Orange 7).
During death with reactive dyes, hydrolysis (i.e., inactivation), an unwanted side reaction that decreases degree of fixation Anionic compounds, used for dyeing nitrogencontaining fabrics such as wool, polyamide, and silk; and modified acryl, bind to cationic NH4 1 /
Ions of these fibers
Acid means the pH in acidic dye dyes rather than the presence of acid groups (sulfonate, carboxyl) in the molecular structure of these dyes Strong complexes of a metal atom (usually chromium, copper, cobalt, or nickel) and one or two dye molecules, respectively, 1:1 and 1:2 metallic complex dyes Relatively large molecules with a high affinity for cellulose fibers in particular. Van der Waals forces bind them to the fiber Attached to the tissue by the addition of a mordant, a chemical that combines with the dye and fiber
used with wool, leather, silk, paper, and modified cellulose fibers Poorly soluble colorants that penetrate into synthetic fibers (cellulose acetate, polyester, polyamide, acrylic, etc.)
This diffusion requires swelling of the fiber, due to high temperatures ( . 120 C) or the help of chemical agents The dyeing takes place in dyebaths with fine dispersed solutions of these dyes Water insoluble dyes particularly and widely used for dyeing cellulose fibers. This method is based on the solubility of the vat dyes in their reduced form (leuco). Reduced with sodium dithionite, soluble leuco vat dyes permeate the tissue. Then the oxidation is applied to bring the dye in its insoluble form

Diarylmethane, triarylmethane, anthraquinone, or azo compounds

Reactive dyes

Acid dyes

Metal dyes

Direct dyes

Mordant dyes

Disperse dyes

Vat dyes

Most (80%) reactive dyes are azo or metal
azo complex compounds but also anthraquinone and reactive phthalocyanine dyes are applied, especially for green and blue

Azo (yellow to red, or a broader range of colors in the case of metal complex azo dyes), anthraquinone, or triarylmethane (blue and green) compounds

Metal complex dyes are usually azo compounds, phthalocyanine metal complex dye Mostly azo dyes with more than one azo bond or phthalocyanine, stilbene, or oxazine compounds Most mordant dyes are azo, oxazine/ triarylmethane compounds, usually dichromate or chromium complexes

Usually small azo or nitro compounds, anthraquinones, or complex metal
azo compounds

Anthraquinones or indigoids

(Continued)

Microbial bioremediation of azo dye through microbiological approach 429 Table 19.1: (Continued) Dye type

Characteristic

Example

Pigment dyes

These insoluble nonionic or insoluble salts retain their crystalline or particulate structure throughout their application. The pigment dyeing is obtained from a dispersed aqueous solution and therefore requires the use of dispersing agents. Pigments are generally used with thickeners in printing pastes for printing various fabrics Dyes (lysochromes) are nonionic dyes that are used for dyeing substrates in which they can dissolve Complex polymer aromatic rings containing heterocyclic S
Dye with sulfur involves reduction and oxidation, comparable to tincture of the tub. They are mainly used for the dyeing of cellulose fibers Dyes (naphthol dyes) are the insoluble products of a reaction between a coupling component (usually naphthols, phenols, or acetoacetylamides); diazotized aromatic amine Food dyes are not used as textile dyes and the use of natural dyes in textile-processing operations is very limited

Azo compounds (yellow, orange, and red) or metal complex phthalocyanines

Solvent dyes

Sulfur dyes

Anionic dyes

Other dyes

Diazo compounds, triarylmethane, anthraquinone, and phthalocyanine solvent dyes

All naphthol dyes are azo compounds

Anthraquinone, indigoid, flavenol, flavone, or chroman

microbes and pollutants are not spread evenly through the environment. Control and optimization of bioremediation processes is a complex system due to many factors. These factors are include the existence of a microbial population capable of degrading pollutants, the availability of contaminants in the microbial population, and environmental factors. Bioremediation is a technique used to eliminate environmental pollutants from the ecosystem. It uses the biological mechanisms inherent in microbes and plants to eradicate dangerous pollutants and restores the ecosystem to its original state (Dawkar, Jadhav, & Govindwar, 2009). The basic principles of bioremediation are to reduce the solubility of these environmental pollutants by modifying the pH, redox reactions, and the adsorption of pollutants from contaminated environments (Chen, Hsueh, Chen, & Li, 2011; Olukanni, Osuntoki, & Gbenle, 2009; Pilatin & Kunduhŏglu, 2011). Several reports have decribed how to improve the biosorption of pentachlorophenol (PCP) by modifying the pH levels in aqueous solutions. For example, the biosorption capabilities of Aspergillus niger (Ali, Hameed, Siddiqui, Ghumro, & Ahmed, 2009) and Mycobacterium chlorophenolicum (Ferraz et al., 2011; Solis-Oba, Eloy-Juarez, Teutli, Nava, & Gonzalez, 2009; Umbuzeiro et al., 2005) in

430 Chapter 19 the removal of PCP from aqueous solutions depended on the pH. Wong and Yuen (1998) also evaluated the influence of pH on the behavior of adsorption and desorption of PCP by M. chlorophenolicum and indicated that the pH values were an essential parameter affecting adsorption of PCP, with the adsorption capacity increasing with the decreasing pH. At pH 5.4, the adsorption by the bacteria was completely irreversible, while complete desorption was achieved at pH 7.0. At pH 6
8, Mansour et al. (2010) obtained better results on the behavior of adsorption of PCP by microbial biomass in aqueous solution. The results obtained by several authors underline the importance of using the appropriate pH to obtain an optimal performance of the microorganisms used during bioremediation. Bioremediation technologies are based on oxidation
reduction processes aimed at modifying the chemistry and microbiology of water by injecting selected reagents into contaminated water to improve the degradation and extraction of various contaminants (Kurade, Waghmode, Kagalkar, & Govindwar, 2012; Parshetti, Telke, Kalyani, & Govindwar, 2010). Redox reactions involve the chemical conversion of harmful pollutants into harmless or less toxic compounds that are more stable, less mobile, or inert (Olukanni et al., 2009). It plays a vital role in the conversion of heavy metals, especially As, Cr, Hg, and Se, in soils and sediments to less toxic or harmless forms (Ali et al., 2009; Solis-Oba et al., 2009). The physicochemical properties of the environment often affect redox reactions in sediments and ground-contaminated groundwater, but can be manipulated by the addition of organic and inorganic modifications such as composts and biochar (Ferraz et al., 2011; Umbuzeiro et al., 2005). Application of organic additives such as compost in metal-contaminated soil may result in differences in the soil microbial population by altering the pH, reducing the solubility of heavy metals and increasing the microbial biomass of algae and available nutrients (Mansour et al., 2010; Wong & Yuen, 1998). The efficacy of bioremediation depends on several factors, such as the nature of the organisms used, the predominant environmental factors in the contaminated site and the degree of contaminants in this environment (Parshetti et al., 2010). Bioremediation can be carried out using microorganisms (microbial bioremediation) that depend on the metabolic potential of microorganisms to degrade environmental pollutants and transform them into harmless forms through oxidation
reduction processes (Kumar, Sathyaselvabala, Premkumar, Vidyadevi, & Sivanesan, 2012; Kurade et al., 2012; Phugare, Kalyani, Patil, & Jadhav, 2011). It can also be achieved by plants that unite, extract, and eliminate contaminants from the environment. The level of contaminated soil, the bioavailability of the metal contaminant, and the accumulation of metals such as biomass by the plant are fundamental to the success of phytosanitary as a means of eradicating heavy metals from the contaminated sites of plants (Zhuo et al., 2011). Bioremediation may be in situ or ex situ. On-site bioremediation is a process of cleaning in situ contaminated environments that involves adding nutrients to contaminated soils to stimulate the ability of microorganisms to degrade contaminants and add new microorganisms to the environment or to improve local microorganisms to degrade specific contaminants through genetic engineering (Anastasi

Microbial bioremediation of azo dye through microbiological approach 431 et al., 2011; Bohmer, Kirsten, Bley, & Noack, 2010; Casieri et al., 2008; Jonstrup, Punzi, & Mattiasson, 2011; Phugare, Kalyani, Surwase, & Jadhav, 2011). The use of natural microorganisms in the environment for in situ bioremediation is affected by the lack of availability of adequate nutrient levels and/or by the environmental configuration of the contaminated site (Arroyo-Figueroa et al., 2011; Carita & Marin-Morales, 2008; GarciaMonta˜no, Domenech, Garcia, Torrades, & Peral, 2008; Kolekar et al., 2012; Porri et al., 2011; Rizzo, 2011). Ex situ bioremediation involves transporting the contaminated environment from the place of origin to another treatment site based on the cost of the treatment, the depth of contamination, the type of contaminant, and the extent of the contamination. Biotic factors affect the degradation of organic compounds through competition between microorganisms for limited carbon sources, antagonistic interactions between microorganisms or the depredation of microorganisms for protozoa and bacteriophages. The rate of degradation of pollutants often depends on the concentration of the pollutant and the amount of “catalyst” present. In this context, the amount of “catalyst” represents the number of organisms capable of metabolizing the pollutant, as well as the amount of enzymes produced by each cell. The expression of specific enzymes by the cells can increase or decrease the degradation rate of the pollutants. In addition, it is necessary to participate in the scope of specific enzymes for the metabolism of pollutants and in their “affinity” for the pollutant and, in addition, in the availability of the contaminant. The main biological factors included mutation, horizontal transfer of genes, enzymatic activity, interaction, growth to the achievement of critical biomass, population size, and composition. Some metabolic characteristics have been investigated related to the efficiency of biodegradation for microorganisms (Acuner & Dilek, 2004). Any factor that may alter growth or metabolism would also affect biodegradation. Therefore the physicochemical characteristics of the environmental matrix, such as temperature, pH, water potential, availability of oxygen and substrate, would influence the efficiency of biodegradation (Fig. 19.1). We should mention two more factors: cometabolism and consortium condition. Some biodegraders need other substrates to degrade pollutants (Dubey, Dubey, Viswas, & Tiwari, 2011). This phenomenon is called cometabolism and is especially necessary for organochlorine compounds. On the contrary, it has been shown that the presence of other carbon sources reduces the biodegradation of organophosphorus. Pollutants can suffer from biodegradation reactions such as dechlorination, division, oxidation, and reduction by different enzymes. Since biodegradation capacity is based on enzymes that are promiscuous and have evolved into detoxifying enzymes, the body’s duplication time is shorter, the body most appropriate is the organism, and it is easier to obtain biodegraders. Thus bacteria, with a duplication time of minutes, are sympathetic to responding to the evolutionary pressure induced by natural or artificial contaminants; this response is to select biotransformation enzymes capable of degrading them. These promiscuous

432 Chapter 19 enzymes are present in organisms even before they carry out the evolutionary pressure, and could induce genetic recombination or mutation leading to enzymes with a better biodegradation capacity. Saha, Swaminathan, Raghavan, Uma, and Subramanian (2010) has consistently reviewed the evolution of metabolic pathways and the factors that affect the efficiency of the biodegradation of pollutants. Although bacteria have been proven to be good biodegraders and biomeasurers, some fungi, plants, and algae could also biodegrade pesticides. Knowledge of the metabolism of these biodegradable species or their stents improves the bioremediation strategy for each site, either through the biostimulation of indigenous biodegraders or by adding exogenously to the site. In addition, thanks to molecular biology, the ability of metabolic biodegradation could be transferred from one biodegrader to another, thus improving its degrading capabilities. Given all this, it is clear that biodegradation enzymes play a key role in bioremediation processes and their knowledge can help into design or choose the most appropriate strategy. The metabolic characteristics of the microorganisms and the physicochemical properties of the target contaminants determine the possible interactions during the process. The real interaction between the two, owever, depends on the environmental conditions of the place of interaction. The growth and activity of microorganisms are affected by pH, temperature, humidity, soil structure, water solubility, nutrients, site characteristics, redox potential and oxygen content, and the physicochemical bioavailability of contaminants (concentration of contaminants, type, solubility, chemical structure, and toxicity). These aforementioned factors determine the kinetics of degradation (Ergene, Ada, Tan, & Katırcıoˇglu, 2009; Mohan, Ramanaiah, & Sarma, 2008; Sivarajasekar, Baskar, & Balakrishnan, 2009). Overall, the bioremediation strategy is shown in Fig. 19.2. Biodegradation can occur under a wide range of pH; however, a pH of 6.5
8.5 is generally optimal for biodegradation in most aquatic and terrestrial systems. Humidity influences the rhythm of the metabolism of contaminants, as it influences the type and quantity of available soluble materials, as well as the osmotic pressure and the pH of terrestrial and aquatic systems (Aksu & Tezer, 2005).

19.4 Effects of environmental parameters on azo dye degradation Research has shown that the yield of biological treatment systems is very significantly influenced by operating parameters. It is necessary to optimize the level of aeration, temperature, and pH in order to optimize the redox potential of the system to produce the maximum color rate reduction The concentrations of the electron donor and the redox mediator must be balanced with the amount of biomass of the system and the amount of dye present in the wastewater. The composition of textile wastewater is varied and may include organic matter, nutrients, salts, sulfur compounds, and toxins, as well as color. Any of these compounds may have an inhibitory effect on the dye reduction process.

Microbial bioremediation of azo dye through microbiological approach 433

Figure 19.2 Overall explanatory mechanism of bioremediation.

19.4.1 Temperature In many systems, the rate of decolorization increases with the increase of temperature, within a defined range that depends on the system. Required temperature in order to achieve the maximum color removal rate, there is a tendency to respond optimally growth of the cell culture at 35
45 C at higher temperatures can be attributed to loss of cell viability or azoreductase enzyme activity (Chu, See, & Phang, 2009). However, it has been shown in certain whole bacterial cell preparations that the azoreductase enzyme is relatively heat stable and can remain active at 60 C for short periods of time. Immobilization of the cell culture in the support material leads to the displacement of the optimum color removal temperature toward higher values due to the microenvironment of the cells.

19.4.2 Oxygen The most important factor to consider is the effect of oxygen on cell growth and reduction of dyes. As mentioned above, during the cell growth stage, oxygen will have one significant effect on the physiological characteristics of the cells. During the dye reduction phase, if the extracellular environment is aerobic, the high redox potential acceptor of electrons, oxygen, can inhibit the dye reduction mechanism. This is due to the electrons released from the oxidation of electron donors by the cells which are preferably used to reduce the oxygen instead of the azo dye (Ertugrul, Bakır, & Donmez, 2008). In addition, the postulated intermediates of the dye reduction reaction, which include the hydrazine form of the dye

434 Chapter 19 and the free radical form of the azo anion of the colorant, tend to be reoxidated by the molecular oxygen. Under anaerobic conditions, no further degradation of the dye molecule is observed after reduction of the azo bond. Aerobic conditions are necessary for complete mineralization of the reactive azo dye molecule, as simple aromatic compounds produced by the initial reduction are degraded by hydroxylation and cycles occurring in the presence of oxygen. Thus for the most efficient wastewater treatment, a two-step process is necessary in which oxygen is introduced after the first anaerobic reduction of the azo bond has occurred. The balance between the anaerobic and aerobic stages of this treatment system must be carefully controlled as it is possible for the reaeration of a reduced dye solution to cause the color of the solution to darken. This is to be expected, as aromatic amines, produced when azo dyes are reduced under anaerobic conditions, are spontaneously unstable in the presence of oxygen. This leads to the oxidation of hydroxyl groups and amino groups to quinines and imines. Such compounds may undergo dimerization or polymerization, leading to the development of new dark color chromophores, which are clearly undesirable by-products.

19.4.3 Dye concentration The concentration of the colored substrate may affect the dye removal efficiency through a combination of factors including dye toxicity at higher concentrations and the ability of the enzyme to recognize the substrate at very low concentrations that may be present in some wastewater. Indeed kinetic constants that control process efficiency, along with other enzyme-catalyzed processes, can be described by Michaelis
Menten kinetics: V 5 Vmax ½S=KM 1 ½S The application of the kinetics of Michaelis
Menten, for example, allow predictions to be made on the efficiency of the process, including the degree of biomass load or the operating temperature necessary to maintain the elimination of dyes at a given efficiency with the constraints defined by the available reactor volume, the composition of the background solution, and flow rates. Fairhead and Thony-Meyer (2012) observed that after a rapid initial reduction of color, the rate of color removal decreased faster than expected by a firstorder reaction. This effect was attributed to the toxicity of the metabolites that were formed during dye reduction. The higher the dye concentration, the longer the time is necessary to remove the color. Yang, Angly, Wang, and Zhang (2011) found that dyes containing 1
10 µM are easily discolored, but when the dye concentration increases to 30 µM the elimination of the color is reduced. Yu and Wen (2005), however, reported the absence of any effect of the dye concentration on the rate of reduction. This observation is compatible with a nonenzymatic reduction mechanism that is controlled by processes that are independent of dye concentration.

Microbial bioremediation of azo dye through microbiological approach 435

19.4.4 Electron donor Oxidation of organic electron donors and/or hydrogen is associated with color removal. Yu and Wen (2005) have shown that the addition of electron donors such as glucose or acetate ions apparently stimulate reductive cleavage of azo bonds. The thermodynamics of different half-electron reactions is different. Therefore the reaction rate is probably influenced by the type of electron donor. It is important to determine a physiological electron donor for each biological color removal process, as this will not only increase the rate of dye reduction, but will also indicate the enzyme pathway responsible for the reduction reaction. For example, if it is a format the most effective electron donor for the anaerobically induced electron transfer pathway to the dye molecule, then it can be concluded that the pathway must include format enzyme dehydrogenase.

19.4.5 pH The optimal pH for color removal is often neutral pH or slightly alkaline pH and color removal rates tend to drop rapidly under strongly acidic conditions or strongly alkaline pH values. As a result, colored wastewater is often increasing cell culture removal efficiency in biological reduction. The azo bond may result in an increase in pH due to the formation of aromatic amine metabolites that are more basic than the original azo compounds. A change in pH between 7.0 and 9.5 has very little effect on the dye reduction process. Aksu and Donmez (2005) found that the dye reduction rate increased almost 2.5 times as pH increased from 5.0 to 7.0, while the speed became insensitive to pH in the range of 7.0
9.5.

19.4.6 Dye structure Some azo dyes are more resistant to bacterial cell removal (Aksu & Donmez, 2005). Simple structure and low-molecular-weight dyes show a higher rate of removal, while removal of paint is more difficult in highly substituted high-molecular-weight dyes. In the case of a terminal nonenzymatic reduction mechanism, the reduction rate is affected by changes in electron density in the azo region group. The substitution of electron-withdrawing groups in the para position of the phenyl ring, relative to the azo group, causes an increase in the reduction rate. Aksu and Donmez (2005) found that the azo compounds with a hydroxyl group or some of the amino groups are more likely to be degraded than the methyl, methoxy, sulfo, or nitro groups. Color removal is also related to the number of azo bonds in the dye molecule. The color of the monoazo dye is removed faster than the color of the ˇ ˇ rik (2005) have shown that diazo or triazo dyes. Safarikova, Ptaˇckova, Kibrikova, and Safa the conversion rate of monoazo dyes increases with increasing dye concentration, while the rate of turnover of diazo dyes and the triazo dye remained constant when the dye concentration increased. A number of researchers correlated the level of color removal with

436 Chapter 19 the dye class, not molecular properties. Safarikova et al. (2005) concluded that acid dyes have low color removal due to the number of sulfonate groups in the dye, direct dyes show high color removal levels that are independent of the number of sulfonate groups in the dye, and reactive dyes show low levels of color removal. The effect of sulfonate groups in color removal is related to the mechanism by which the color is removed. If dye reduction occurs within the cell, the presence of sulfonate groups will prevent the transfer of the dye molecule by the cell membrane. The oxidation of organic electron donors and/or hydrogen is coupled to the color removal process. Safarikova et al. (2005) showed that the addition of electron donors as glucose or acetate apparently stimulates the reduction of the division of the azo links. The thermodynamics of the different electron donation reactions are different. So the reaction rate is likely to be influenced by the type of electron donor. It is important to determine the physiological donor of electrons for each process of biological color elimination, as this will not only increase the rate of reduction of dyes, but also give an indication of the enzymatic route responsible for the reduction reaction. For example, if it gives more effective electrons from the anaerobically induced electron transfer pathway to the dye molecule, then it can be concluded that the path must involve a form of enzyme dehydrogenase. The concentration of the auxiliary substrate controls the speed of formation of reduction equivalents. Safarik et al. (2007) found that cellulose lysis products can function as an electron donor for anaerobic azo dye. Some chemicals, for example, thimerosal and p-chlorameric benzoate, inhibit the alcohol dehydrogenase from nicotinamide adenine dinucleotide (NADH)-generating systems that is necessary to generate reduction equivalents for the reduction of dyes. Therefore the rate of NADH formation would be a limitation of the rate, causing an inhibition of the reduction of azo dyes (Safarik et al., 2007).

19.4.7 Redox potential The elimination of color depends on the redox potential of the electron donors and acceptors (Fig. 19.3), because the speed control step implies a redox balance between the colorant and the extracellular reducer. The redox potential is a measure of the ease with which a molecule will accept electrons and can be reduced. Therefore the more positive the redox potential, the more easily reduced is the molecule (Safarik et al., 2007). As a result, the color elimination rate will increase with the increased medium wave potential of the azo substrate. Safarikova et al. (2005) showed there is a linear relationship between the logarithm of the color elimination rate and the half-wave substrate potential. The electrochemical property of the dye substrate suggests the determining step of the velocity. The reduction of bacterial dyes does not imply a specific structural phenomenon like selective permeation of the membrane or enzymatic junction (Yu & Wen, 2005). The establishment of low reduction and oxidation potentials (, 2 400 mV) for the system,

Microbial bioremediation of azo dye through microbiological approach 437 Colored solution containing dye

Colorless solution containing amines X

X

N Chromophore N

Redox mediatorred

Redox mediatorox

NH2 NH2

Azoreductase NADH

X

NAD+

Carbon complexes

Dehydrogenase (enzyme liberating e–)

X Oxidation products

Cell

Figure 19.3 Schematic representation of azo dye reduction.

below anaerobic conditions, is necessary for high color elimination rates and has an effect on the profile of metabolites that are generated during the reduction process (Safarikova et al., 2005). The color elimination rate is higher when the redox potential of the system is the most negative and the rate falls as the redox potential of the system increases (Safarik et al., 2007). Azo dye reduction is shown systematically in Fig. 19.3.

19.4.8 Redox mediator Highly sulfonated azo dyes are unlikely to pass through the cell membrane, and therefore the dye reduction reaction should be an extracellular reduction activity (Fairhead & ThonyMeyer, 2012; Yang et al., 2011; Yu & Wen, 2005). This reduced activity is achieved using redox mediator compounds, such as Flavin’s, to facilitate equivalent transfers of cell launchers the nonenzymatic reduction of the extracellular azo dye (Chu et al., 2009). A small concentration of the redox mediator is sufficient for this electron transfer to take place. The redox mediators are characterized by a redox potential that oscillates between 2200 and 2350 mV (Yang et al., 2011). The addition of synthetic electron carriers improves the pace of reduction of azo dyes for bacterial cells. Quinone
hydroquinone round pairs are known to act as redox mediators (Yu & Wen, 2005). The quinones are formed via a reduction of hydroquinone radicals by one electron or the reduction of hydroquinone by two electrons. For the reduction of the azoic dyes, the reduction to the anion radical is produced by a rapid transfer reaction of a single electron, followed by a

438 Chapter 19 second slower electron transfer event to produce a stable diazo molecule (Safarik et al., 2007). Hydroquinone is then avoided by the dye in a direct chemical reaction (Aksu & Tezer, 2005; Chu et al., 2009; Ergene et al., 2009). The application of natural and biodegradable quinones, such as laws, has technical potential for treatment systems for color removal because the reduction rate is increased without adding any substances with environmental problems (Chen et al., 2011; Olukanni et al., 2009; Pilatin & Kunduhŏglu 2011). Safarikova et al. (2005) observed a certain elimination of color in the presence of autoclave cells, suggesting the existence of an active reducing factor capable of reducing dyes in the absence of microbial activity. The reduction reaction, which includes the redox mediator, includes the redox potential of the mediator in relation to coloring azo and with the specificity of reducing enzymes with regarding the mediator

19.4.9 Decolorization by genetically modified organisms Bioremediation is a useful methodology for the treatment of textile wastewater, but the variation in physicochemical characteristics of the effluents, including the pH, NaCl and other salt contents, temperature, and the presence of organic compounds, can result in the deactivation of enzymes and microbial cells. Thus more active and versatile enzymes and microorganisms are required with high stability, high production, and low suitable cost in order to meet the needs of the textile industry wastewater treatment. There are molecular biology methodologies, such as cloning, heterologous expression, random mutagenesis, directed mutagenesis, recombination techniques for targeted genes evolution, rational and metagenomics design, to accelerate evolution processes. In addition, advances in genetic and molecular genetics engineering has allowed the cloning and expression of virtually any gene in a suitable microbial host.

19.5 Conclusion Residual dyes, together with the other auxiliary chemical reagents used for processing, impurities of the raw materials, and other hazardous materials applied to the finishing process, impose massive difficulties for wastewater treatment systems. The release of colored wastewater into the ecosystem is a remarkable source of esthetic contamination, eutrophication, and disturbances in aquatic life. These concerns have given rise to new and/ or stricter regulations on wastewater discharges of color.

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72. Kurade, M. B., Waghmode, T. R., Kagalkar, A. N., & Govindwar, S. P. (2012). Decolorization of textile industry effluent containing disperse dye Scarlet RR by a newly developed bacterial-yeast consortium BL-GG. Chemical Engineering Journal, 184, 33
41. Machado, K. M. G., Compart, L. C. A., Morais, R. O., Rosa, L. H., & Santos, M. H. (2006). Biodegradation of reactive textile dyes by basidiomycetous fungi from Brazilian ecosystems. Brazilian Journal of Microbiology, 37, 481
487. Mansour, H. B., Ayed-Ajmi, Y., Mosrati, R., Corroler, D., Ghedira, K., Barillier, D., et al. (2010). Acid violet 7 and its biodegradation products induce chromosome aberrations, lipid peroxidation and cholinesterase inhibition in mouse bone marrow. Environmental Science and Pollution Research International, 17, 1371
1378. Meng, X., Liu, G., Zhou, J., Fu, Q. S., & Wanga, G. (2012). Azo dye decolorization by Shewanella aquimarina under saline conditions. Bioresource Technology, 114, 95
101. Mohan, S. V., Ramanaiah, S. V., & Sarma, P. N. (2008). Biosorption of direct azo dye from aqueous phase onto Spirogyra sp. I02: Evaluation of kinetics and mechanistic aspects. Biochemical Engineering Journal, 38, 61
69. Ogugbue, C. J., Sawidis, T., & Oranusi, N. A. (2012). Bioremoval of chemically different synthetic dyes by Aeromonas hydrophila in simulated wastewater containing dyeing auxiliaries. Annals of Microbiology. Available from https://doi.org/10.1007/s13213-011-0354-y. Olukanni, O. D., Osuntoki, A. A., & Gbenle, G. O. (2009). Decolourizatioin of azo dyes by a strain of Micrococcus isolated from a refuse dump soil. Biotechnology, 8, 442
448. Parshetti, G. K., Telke, A. A., Kalyani, D. C., & Govindwar, S. P. (2010). Decolorization and detoxification of sulfonated azo dye Methyl Orange by Kocuria rosea MTCC 1532. Journal of Hazardous Materials, 176, 503
509. Phugare, S. S., Kalyani, D. C., Patil, A. V., & Jadhav, J. P. (2011). Textile dye degradation by bacterial consortium and subsequent toxicological analysis of dye and dye metabolites using cytotoxicity, genotoxicity and oxidative stress studies. Journal of Hazardous Materials, 186, 713
723. Phugare, S. S., Kalyani, D. C., Surwase, S. N., & Jadhav, J. P. (2011). Ecofriendly degradation, decolorization and detoxification of textile effluent by a developed bacterial consortium. Ecotoxicology and Environmental Safety, 74, 1288
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135. Ponraj, M., Jamunarani, P., & Zambare, V. (2011). Isolation and optimization of culture conditions for decolorization of True Blue using dye decolorizing fungi. Asian Journal of Experimental Biological Sciences, 2, 270
277. Porri, A., Baroncelli, R., Guglielminetti, L., Sarrocco, S., Guazzelli, L., Forti, M., et al. (2011). Fusarium oxysporum degradation and detoxification of a new textileglycoconjugate azo dye (GAD). Fungal Biology, 115, 30
37.

Microbial bioremediation of azo dye through microbiological approach 441 Priya, B., Uma, L., Ahamed, A. K., Subramanian, G., & Prabaharan, D. (2011). Ability to use the diazo dye C. I. Acid Black 1 as a nitrogen source by the marine cyanobacterium Oscillatoria curviceps BDU92191. Bioresource Technology, 102, 7218
7223. Rizzo, L. (2011). Bioassays as a tool for evaluating advanced oxidation processes in water and wastewater treatment. Water Research, 45, 4311
4340. Safarik, I., Rego, L. F. T., Borovska, M., Mosiniewicz-Szablewska, E., Weyda, F., & Safarikova, M. (2007). New magnetically responsive yeast-based biosorbent for the efficient removal of water-soluble dyes. Enzyme and Microbial Technology, 40, 1551
1556. ˇ ˇ rik, I. (2005). Biosorption of water-soluble dyes on Safarikova, M., Ptaˇckova, L., Kibrikova, I., & Safa magnetically modified Saccharomyces cerevisiae subsp. uvarum cells. Chemosphere, 59, 831
835. Saha, S. K., Swaminathan, P., Raghavan, C., Uma, L., & Subramanian, G. (2010). Ligninolytic and antioxidative enzymes of a marine cyanobacteium Oscillatoria willei BDU 130511 during Poly R-478 decolourization. Bioresource Technology, 101, 3076
3084. Saratale, R. G., Saratale, G. D., Chang, J. S., & Govindwar, S. P. (2009). Ecofriendly degradation of sulfonated diazo dye C.I Reactive Green 19A using Micrococcus glutamicus NCIM-2168. Bioresource Technology, 100, 3897
3905. Saratale, R. G., Saratale, G. D., Kalyani, D. C., Chang, J. S., & Govindwar, S. P. (2009). Enhanced decolorization and biodegradation of textile azo dye Scarlet R by using developed microbial consortiumGR. Bioresource Technology, 100, 2493
2500. Sivarajasekar, N., Baskar, R., & Balakrishnan, V. (2009). Biosorption of an azo dye from aqueous solutions onto Spirogyra. Journal of the University of Chemical Technology and Metallurgy, 44, 157
164. Solis-Oba, M., Eloy-Juarez, M., Teutli, M., Nava, J. L., & Gonzalez, I. (2009). Comparison of advanced techniques for the treatment of an indigo model solution: Electro incineration, chemical coagulation and enzymatic. Revista Mexicana de Ingenieria Quimica, 8, 275
282. Umbuzeiro, G. A., Freeman, H. S., Warren, S. H., Palma de Oliveira, D., Terao, Y., Watanabe, T., et al. (2005). The contribution of azo dyes to the mutagenic activity of the Cristais River. Chemosphere, 60, 55
64. Wong, P. K., & Yuen, P. Y. (1998). Decolourization and biodegradation of N,N0 -dimethylpphenylenediamine by Klebsiella pneumoniae RS-13 and Acetobacter liquefaciens S-1. Journal of Applied Microbiology, 85, 79
87. Yang, Q., Angly, F. E., Wang, Z., & Zhang, H. (2011). Wastewater treatment systems harbour specific and diverse yeast communities. Biochemical Engineering Journal, 58, 168
176. Yu, Z., & Wen, X. (2005). Screening and identification of yeasts for decolorizing synthetic dyes in industrial wastewater. International Biodeterioration and Biodegradation, 56, 109
114. Zhuo, R., Ma, L., Fan, F., Gong, Y., Wan, X., Jiang, M., et al. (2011). Decolorization of different dyes by a newly isolated white-rot fungi strain Ganoderma sp. En3 and cloning and functional analysis of its laccase gene. Journal of Hazardous Materials, 192, 855
873.

Further reading Das, D., Charumathi, D., & Das, N. (2011). Bioaccumulation of the synthetic dye Basic Violet 3 and heavy metals in single and binary systems by Candida tropicalis grown in a sugarcane bagasse extract medium: Modelling optimal conditions using response surface methodology (RSM) and inhibition kinetics. Journal of Hazardous Materials, 186, 1541
1552.

CHAPTER 20

Novel process of ellagic acid synthesis from waste generated from mango pulp processing industries Murugan Athiappan1, Shantkriti Srinivasan2, Rubavathi Anandan1 and Janani Rajaram1 1

Department of Microbiology, Periyar University, Salem, India, 2Department of Biotechnology, Kalasalingam Academy of Research and Education, Krishnankoil, India

20.1 Introduction 20.1.1 Waste from mango pulp processing industries Mango (Mangifera indica L.) is listed among the most important tropical fruits in the world, ranking among the top 10 fruits in the overall production of major fruit crops all over the world. Salem, Tamil Nadu is popularly known as “mango city” and produces 12 million tons of mangos per year. Being seasonal fruit, mangos are harvested and processed throughout the year for pulp, peels, and kernels (Beerh, Raghuramaiah, Krishnamurthy, & Gridhar, 1976). About 54 mango processing units are housed in Tamil Nadu, 46 units are located in Salem and Krishnagiri district (Jagadeesan & Shankar, 2014). Most of the waste disposal in the fruit processing sector is due the poor quality of the fruits. About 30% 50% of the mango fruit goes into the waste, resulting in around 200,000 lakh tons of waste generated per season by the pulp industries in these area which are being discharged without proper treatment methods (El-Kholy, Solta, Abd El-Rahman, El-Saidy, & Foda, 2008; Jedele, Hau, & Von Oppen, 2003). Being small-scale industries, the disposal of solid waste is a burden (Carlasson & Amand, 2012). Industries discharge of waste on the outskirts of towns and cities has created overflowing landfills, which are not only impossible to reclaim but also have serious environmental implications in terms of groundwater pollution and contributions to global warming (Manigat, Wallet, & Claude Andre, 2010; Shah & Masoodi, 1994). It has been estimated that Rs. 50,000 per month is being spent on waste disposal (Shah & Masoodi, 1994).

Emerging Technologies in Environmental Bioremediation. DOI: https://doi.org/10.1016/B978-0-12-819860-5.00020-1 © 2020 Elsevier Inc. All rights reserved.

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444 Chapter 20

20.2 Composition of mango wastes By-products from mango industrial processing represent between 35% and 60% of the fresh fruit weight (Larrauri et al., 1996). The by-products are mainly peel, seed, and pulp residues. A regular 3-month production season constitutes over 7,500 tons of polluting byproducts. The residues represent a waste of nutrients and a source of environmental contaminants. Such by-products could be valuable sources of dietary fiber, antioxidant compounds, and single carbohydrates. Tannins are a group of naturally occurring complex phenolic compounds influencing the nutritive value of various foods by protein precipitation and tend to have an acrid taste. Tannins are secondary metabolites of plants (Bhat, Singh, & Sharma, 1998). They are located in the cytoplasm and vacuoles of plants from where they play a defense role (Khadem & Marles, 2010). Based on their sugar content and their degrees of polymerization and esterification the tannins can be broken down into four groups: condensed tannins (CT), complex tannins, gallotannins, and ellagitannins (ETs) (Khanbabee & Van Ree, 2001).

20.3 Types of tannins CT are the polymers of flavan-3-ol or flavan-3,4-diol, which do not have sugar residues (Lekha & Lonsane, 1997). They are known as polymeric proanthocianidins and leucoanthocyanidins based on flavan units, mainly catechin or epicatechin. Among the different kinds of tannins, gallotannin hydrolysis for the production of gallic acid is well studied (Aguilar et al., 2007). Ellagitannins have been less explored due to their diversity and chemical complexity (Aguilera-Carbo et al., 2007; Huang et al., 2007; Huang, Ni, & Borthwick, 2005; Huang, Niu, Gong, & Lu, 2007; Robledo-Olivo, Martinez, AguileraCarbo, Garza-Garcia, & Aguilar, 2006). Complex tannins are the compounds which occur due to the linkage of catechins or epicatechins with gallic or ellagic acids (EAs) and this reaction is catalyzed by light, heat, and oxygen. Catechin gallate, which contains hydrolysable tannin (HT) differ from the condensed moieties in their chemical structure (Khanbabee & Van Ree, 2001). Gallotannins, which are polyphenols, are susceptible to hydrolysis by enzymes (tannase activity), acids, or alkalis (Huang et al., 2005). Gallotannins have a central glucose moiety with five (pantagalloyl glucose) to nine (nanogalloyl glucose) galloyl residues and is a tannic acid (Belmares-Cerda, Contreras-Esquivel, Rodrı´guezHerrera, Ramı´rez Coronel, & Aguilar, 2004). Under present-day conditions, not only are tannins becoming of increasing importance as raw materials for industrial chemical processing in their own right, but their production also represents an agricultural operation of appreciable magnitude. Also one must consider the energetic costs and the reasons for such practices, especially when plants devote so much carbon to the production of tannins. Tannins are divided into two main groups: HT and CT

Novel process of ellagic acid synthesis from waste generated from mango pulp processing

445

Figure 20.1 Classification of tannins.

(Maheri-Sis et al., 2011). Hexahydroxydyphenic acid is present in ETs, and it is converted into EA by hydrolysis where it is dehydrated, followed by spontaneous lactonization. Gallic, ellagic, quinic, caffeic, and ferulic acids are among the monophenols derived from tannins and are an important group of molecules with interesting biological activities (Fig. 20.1).

20.4 Bioconversion of tannin to ellagic acid Mango wastes are processed for the production of ethanol by yeast or bacteria, and methane by microbial consortia. Anaerobic microbial fermentation may also lead to the conversion of wastes to fuels, including ethanol. Bioconversion of waste into different valuable products has been achieved (Table 20.1). Waste containing ETs, low-biodegradable phytochemicals, can be converted into value products such as EA. According to Prabha and Patwardhan (1986) EA is the predominant substrate in mango peel. EA was described in unripe mango fruits for the first time by Saleh and El-Ansari (1975). EEA produced from fruit and vegetable processing industries, including the pomace of apple, currant, citrus fruit, carrot, tomato, melon, mango, grapes or spinach, is convenient, cost-effective, and enables the rational management of troublesome waste (Duda-Chodak & Tarko, 2007).

Table 20.1: Sources of substrates that are used for the production of ellagic acid (EA). Crops

Waste type

Bioactive compounds

Activity

Mango

Peels, pits/seeds, and pulp Pomace (peels, core, seeds, calyces, and stem) Rinds and flesh

Tannins, vanillin, and mangiferin

Antioxidant

Pectin, catechin, hydroxycinnamates, phloretin glycosides, quercetin glycosides, and procyanidins

Antioxidant

Apple

Watermelon

Lycopene, citrulline, and phenolic compounds Rambutan Peels Tannins Mangosteen Pericarps Proanthoccyanidins Guajuva Bagasse Epicatechin, quercetin, syringic acid, and mirycetin Sugars, α-amylase, laccase, and citric Banana Dried leaves, pseudostems, and acid peels Lemon Peels Flavonoids, saponins, tannins, alkaloids, steroids, triterpenes, essential oil, and limonene Avocado Peel and seeds Epicatechin, gallic acid, and chloronic acid Grapes Seeds and skin Coumaric acid, EA, ellagitannins (ETs), gallotannins, and reversatrol Lichee Pericarp and Epicatechin and gallic acid seeds Pomegranate Peel and pericarp EA, gallic acid, and coumaric acid

Carrot Cucumber

Peel Peel

β-Carotene Gallic acid and coumaric acid

Potato

Peel

Tomato Barley Rice

Skin and pomace Bran Bran

Vanillic acid, caffeic acid, and gallic acid Carotenoids β-Glucan γ-O-Oryzanol oil

Wheat Strawberry

Bran and gram Seeds and pulp

Phenolic acids EA, gallic acid, and coumaric acid

Raspberry

Seeds and pulp

EA, gallic acid, and coumaric acid

Blackberry bush

Seeds and pulp

EA, gallic acid, and coumaric acid

Antioxidant and food additives Antioxidant Antioxidant Antimicrobial Fermentation and enzyme production Antimicrobial, nematostatic activity, and insecticidal (larvicidal) Antiobesity, antioxidant, and antidiabetic Antiobesity, antioxidant, antidiabetic, and antimicrobial Antimicrobial and antioxidant Antiobesity, antiviral, antioxidant, antidiabetic, and antimicrobial Antioxidant and food additives Antioxidant, food additives, and antiobesity Skin whitening and anitoxidant Antioxidant and anticancer Antiobesity Antioxidant and antiinflammatory Antioxidants Antiobesity, antiviral, antioxidant, antidiabetic, antimicrobial, and antiinflammatory Antiobesity, antiviral, antioxidant, antidiabetic, antimicrobial, and antiinflammatory Antiobesity, antiviral, antioxidant, antidiabetic, antimicrobial, and antiinflammatory (Continued)

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Table 20.1: (Continued) Crops

Waste type

Bioactive compounds

Peach

Seeds and pulp

EA, gallic acid, and vanillic acid

Clove Chestnut

Clove Nut

Cashew

Pericarp

Hazel nut

Pericarp and nut

Activity

Antiobesity, antiviral, antioxidant, antidiabetic, antimicrobial, and antiinflammatory Epicatechin and gallic acid Antioxidant Ellagitannins, gallotannins, and Antiviral, antioxidant, and reversatol antidiabetic Coumaric acid, EA, ETs, gallotannins, Antiobesity, antiviral, and reversatol antioxidant, antidiabetic, antimicrobial, and antiinflammatory Gallic acid, EA, vanillic acid, ETs, Antiobesity, antiviral, gallotannins, and reversatol antioxidant, antidiabetic, antimicrobial, and antiinflammatory

20.5 Microbes involved in the production of ellagic acid It is well-known that the advantages of microorganisms in the preparations of natural products are not only technical (high specificity, mild reaction conditions) but also economic (costs of industrial processes). Although ETs have the typical C C bond, which is more difficult to degrade than in gallotannins, concerted efforts are still in progress to improve ET degradation and utilization (Li, Kai, Qiang, & Dongying, 2006; Li et al., 2006). In recent years, more attention has been mainly focused on intestinal microflora biodegradation of tannins, especially ETs that can contribute to the definition of their bioavailability for both human beings and ruminants (Goel, Puniya, Aguilar, & Singh, 2005). Furthermore, there have been endeavors to utilize the tannin-degrading activity of different fungi for ET-rich biomass, which will facilitate the application of tannin-degrading enzymes in strategies for improving industrial and livestock production. Microbes used for the bioconversion of agroindustrial waste containing ET are organisms that are considered Generally Recognized As Safe (Robledo-Olivo, Martı´nez, Aguilera-Carbo´, Garza-Garcı´a, & Aguilar, 2008). There are a few studies about ET biodegradation by fungal enzymes. The enzymatic hydrolysis of ETs (novotannin O and novotannin P) from Tibouchina multiflora by an Aspergillus niger tannase for EA production was demonstrated in 1999; this kind of substrate requires a specific biocatalyst (such as ellagitannase) that is different from the classical tannase used for gallic acid production (Yoshida, Amakura, Yokura, & Ito, 1999). Different types of microorganisms have been exploited for the production of EA from ETs. A. niger, Mucor sp., Penicillium sp., Penicillium verrucosum, and Fusarium sp. have been reported to produce ellagitannase (De la Cruz Medina et al., 2002; Shi, Qiang, Kai, Huang, & Quin, 2005).

448 Chapter 20 Table 20.2: Microbes involved in the bioconversion of ellagitannin (ET). Microorganisms

Sources of ETs

Lentinus eddes Rhizopus oligosporous Aspergillus niger/Candida utilis Aspergillus SHL6 Aspergillus niger GH1 Aspergillus oryzae Aspergillus oryzae/Trichoderma ressei Aspergillus niger PSH A. niger GH1

Cranberry pomace Cranberry pomace Fruit shell of Quercus aegilops (Valonea) Fruit shell of Q. aegilops (Valonea) Pomegranate husk Acorn fringe Acorn cups Leaves of creosote bush and tar bush Leaves of creosote bush

EA was produced using cranberry pomace and Rhizopus oligosporous by solid-state fermentation (Vattem & Shetty, 2002). Lentinus edodes also was found to produce EA from the same cranberry pomace (Vattem and Shetty, 2003). A. niger and Candida utilis demonstrated a bioconversion process of valonia tannins from oak trees through solid-state fermentation with a correspondent accumulation of EA (Shi et al., 2005). Aspergillus niger SHL6 strain produces EA by submerged fermentation (Huang et al., 2005). Aspergillus oryzae, Endomyces fibulige, and Trichoderma reesei bioconverted oak acorns into EA by solid-state fermentation (Huang et al., 2008). A. niger GH1 strain produces EA from the ETs of pomegranate peel by solid-state fermentation (Herna´ndez, Aguilera-Carbo´, Rodrı´guez-Herrera, Martı´nez, & Aguilar, 2008). EA production was also obtained by A. niger PSH strain from the ETs of tar bush by solid-state fermentation (Ventura et al., 2007) (Table 20.2). In addition, bacterial tannase can also degrade and hydrolyze natural tannins very efficiently (Deschamps, Otuk, & Lebeault, 1983; Lewis & Starkey, 1969) (Fig. 20.2).

20.6 Applications of ellagic acid Over the past decades, there has been a significantly increased use of dietary supplements, such as vitamins, herbal supplements, and medicinal foods, in the general population. From 1990 to 1997, more than 15 million adults in the United States reported using herbal supplements in conjunction with prescribed medications (Eisenberg, Sundgren, & Wells, 1998). In addition, from 1997 to 2002, there was a further increase of more than 50% in the consumption of dietary supplements (Tindle, Davis, Phillips, & Eisenberg, 2005). Ellagic acid is a bioactive molecule of added value with biological properties and can be applied in food technology as a nutraceutical agent in order to improve food quality and in the healthcare industries for medical, pharmaceutical, and cosmetics uses. EA has been reported to have antimutagenic, anticarcinogenic, antioxidant, antiviral, and antiinflammatory activities (Ogawa et al., 2002; Priyadarsini et al., 2002). The key activities

Novel process of ellagic acid synthesis from waste generated from mango pulp processing

449

Ellagitannins E? HHDP +

E?

Galloyl glucose

Ellagiloyl glucose E?

TAH Glucose + gallic acid

+ HHDP

HHDP + glucose

Q? Glycolysis PO/PFO

GAD Pyrogallol

Gallic acid

Ellagic acid

Pyruvate

Scheme of bioegradation of ellagitannin E? unknown enzyme, TAH tannin acyl hydrolase, PO peroxidase PFO Polyphenolcidase

Figure 20.2 Pathway leading to the formation of ellagic acid (EA).

of ETs include antioxidant property that has been used for the prevention of cardiovascular diseases and cancer (Manach & Scalber 2004; Madrigal-Carballo et al., 2009). EA has antimicrobial, antihepatotoxic, antisteatosic, anticholestic, antifibrogenic, and antihepatocarcinogenic properties (Garcı´a-Nino & Zazueta, 2015; Rasool et al., 2015; Rubial et al., 2003; Williner, Pirovani, & Gu¨emes, 2003). It is also a potent gas-protective drug, reduces birth defects, and has antiaging properties (Ahn et al., 1996; Presser, Attias, Liker, & Hayek, 2004; Vattem et al., 2005). EA acts as a remarkable depigmentation agent that inhibits melanin production (Bae, 2011). Current research seems to indicate that the most therapeutically beneficial plant constituents are ETs (including EA), punicic acid, flavonoids, anthocyanidins, anthocyanins, and estrogenic flavonols and flavones (Espı´n, Garcı´a-Conesa, & Toma´s-Barbera´n, 2007). In the human body EA may promote good health and be effective in the prevention of cancer, heart disease, and other chronic diseases. EA and ETs are active agents that induce vasorelaxation, oxygen free radical scavenging, as well as hypolipidemic, antiinflammatory, and anticarcinogenic activities. Antioxidants are compounds that can delay, inhibit, or prevent the oxidation of compounds, trapping free radicals and reducing oxidative stress.

450 Chapter 20

Anticancer

Antiinflammatory

Antioxidant

Congulant

Ellagic acid

Skin whitening

Cardioprote ction

Antidepressant Antidiabetic

Figure 20.3 Prospective applications of ellagic acid (EA).

EA prevents the formation of various tumors, this mechanism of action can be possible because compounds such as ETs and EA explicitly interact with the cells walls or sites with the facility to complex proteins, preventing the proliferation of metastatic cells (Williner et al., 2003). EA was characterized to have the antineurogenerative properties (MertensTalcott, Bomser, Romero, Talcott, & Percival, 2005), and also was shown to have antioxidant, antihepatotoxic, antisteatosic, anticholestatic, antifibrogenic, antihepatocarcinogenic, and antiviral properties (Garcı´a-Nino & Zazueta, 2015; Rasool et al., 2015; Williner et al., 2003) (Fig. 20.3). A nutraceutical is a food which has medical and health benefits, including the prevention and/or treatment of disease. Nutraceuticals have health-promoting effects and are mediated through biochemical and cellular interactions. The various classes of nutraceutical are polyphenolics, dietary fiber, polysaccharides, proteins, biopigments, edible oils, and emulsifiers. In the food industry, synthetic antioxidants, such as butylated hydroxyanizole and butylated hydroxytoluene, have long been widely used as antioxidant additives to preserve and stabilize the nutritive value of foods. Mango fruits and their by-products, that is, mango pulp, peel, and seed, provide a whole range of nutritional and medicinal properties due to their high contents of polyphenols and specific enzymes. A ripe mango

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has high antioxidant and anticarcinogenic properties. In the pharmaceutical industries, mango seed kernel extract and mango peel have various uses since they contain ethyl gallate and penta-O-galloyl-glucoside (PGG) which possess potent hydroxyl radicals, superoxide anions, and singlet oxygen scavenging activity. PGG also possesses bioactivity including antitumor, antioxidant, anticardiovascular, and hepatoprotective effects. Certain compounds in passion fruit peel have a bronchodilator effect and can help relieve bronchospasm in asthma patients. Oral administration of the purple passion fruit peel extract is considered to reduce wheeze and cough and improve shortness of breath in adults with asthma (Wadhwa1, Bakshi1, & Makkar, 2015). Iokused is a dietary fiber supplement as fruit and vegetable by-products such as apple, pear, orange, peach, black currant, cherry, artichoke, asparagus, mango peels, onion and carrot pomace. Dietary fibers and phytochemicals are gaining increased attention because of their antioxidant, and anticarcinogenic properties. The presence of bound phenolics in dietary fiber adds an additional antioxidant property and its antioxidant capacity is greater than that of DL-alpha-tocophenol which is an important source of high-quality antioxidant dietary fiber, pectin, polyphenols, and carotenoids. Mango peel can be used as a functional ingredient to develop healthy food products such as noodles, bread, biscuits, sponge cakes, and other bakery products. The dried carrot pomace can be used in wheat bread at a level of 5% to produce high-fiber biscuits, cake, dressing, and pickles as a supplement of carotenoides, fiber, and minerals and in functional drinks (Chaisawadi, Suwanyeun, & Boonnumma, 2014).

20.7 Conclusion In this chapter the issues related to the discharge of waste from the mango pulp processing industries were critically analyzed. The resources like tannins available in the mango waste can be harnessed by microbial conversion into EA. EA produced through microbial synthesis would be compatible for human use. Solid-state fermentation using A. niger or submerged fermentation with Micrococcus luteus would be economically feasible and safe. Instead of extracting EA from plants, synthesizing it from agroindustrial waste will be a wiser method of waste disposal.

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454 Chapter 20 Saleh, N. A. M., & El-Ansari, M. A. I. (1975). Polyphenolics of twenty local verities of Mangifera indica. Planta Medica, 28, 124 130. Shah, G. H., & Masoodi, F. A. (1994). Studies on the utilization of wastes from apple processing plants. Indian Food Packer, 48(5), 47 52. Shi, B., Qiang, H., Kai, Y., Huang, W., & Quin, L. (2005). Production Of ellagic acid from degradation of valonea tannins By Aspergillus niger and Candida utilis. Journal of Chemical Technology and Biotechnology, 80, 1154 1159. Tindle, H. A., Davis, R. B., Phillips, R. S., & Eisenberg, D. M. (2005). Trends in use of complementary and alternative medicine by US adults: 1997 2002. Alternative Therapies in Health and Medicine, 11, 42 49. Vattem, D. A., & Shetty, K. (2002). Solid-state production of phenolic antioxidants from cranberry pomace by Rhizopus oligosporum. Food Biotechnology, 16, 189 210. Ventura, J., Belmares-Cerda, R., Aguilera-Carbo´, A., Contreras-Esquivel, J. C., Rodrı´guez-Herrera, R., & Aguilar, C. N. (2007). Fungal biodegradation of tannins from creosote bush (Larreatri dentata) and tar bush (Fluorensia cernua) for gallic and ellagic acid production. Food Technology and Biotechnology, 46(2), 213 217. Wadhwal, M., Bakshil, M. P. S., & Makkar, H. P. S. (2015). Wastes to worth: Value added products from fruit and vegetable wastes. CAB Reviews, No. 043, 10. Williner, M. R., Pirovani, M. E., & Gu¨emes, D. R. (2003). Ellagic acid content in strawberries of different cultivars and ripening stages. Journal of the Science of Food and Agriculture, 83, 842 845. Yoshida, T., Amakura, Y., Yokura, N., & Ito, H. (1999). Oligomeric hydrolysable tannins from Tibouchina multiflora. Phytochemistry, 52, 1661 1666.

Further reading Ehrenberg, L., Osterman-Golkar, S., Segernick, D., SvenssonAnd, K., & Calleman, C. J. (1977). Evaluation of genetic risks of alkylating agents. III. Alkylation of haemoglobin after metabolic conversion of ethene to ethene oxide in vivo. Mutation Research, 45, 175 184. Li, Y., Zhao, X., Zu, Y., Zhang, Y., Ge, Y., Zhong, C., & Wu, W. (2015). Preparation and characterization of micronized ellagic acid using antisolvent precipitation for oral delivery. International Journal of Pharmaceutics, 486(1 2), 207 216. Li, Z. J., Guo, X., Dawuti, G., & Aibai, S. (2015). Antifungal activity of ellagic acid in vitro and in vivo. Phytotherapy Research: PTR, 29, 1019 1025. Prashanth, D., Asha, M. K., & Amit, A. (2001). Antibacterial activity of Punica granatum. Fitoterapia, 72, 171 173. Wang, Y., Fang, X., An, F., Wang, G., & Zhang, X. (2011). Improvement of antibiotic activity of Xenorhabdus bovienii by medium optimization using response surface methodology. Microbial Cell Factories, 10, Article No. 98. Wang, Y., et al. (2015). In vitro antiproliferative and antioxidant effects of urolithin A, the colonic metabolite of ellagic acid, on hepatocellular carcinomas HepG2 cells. Toxicology In Vitro: An International Journal Published in Association With BIBRA, 29, 1107 1115. Zhang, L., Ji, X.-J., Cong, Z.-X., Zhu, J.-H., & Zhou, Y. (2013). FTY720 for cancer therapy. Oncology Reports, 30(6), 2571 2578.

Index Note: Page numbers followed by “f” and “t” refer to figures and tables, respectively.

A Absorption of light, by PAHs, 169 ABTS. See 3-Ethylbenzthiazoline6-sulfonate (ABTS) Accelerated bioremediation, 44 49 bioaugmentation, 47 49 biostimulation, 46 47 Acetoclastic methanogens, 23 25 Acetyl-coenzyme A (Acetyl-CoA), 41 43 Acidification volatilization reneutralization process, 279 280 Acidithiobacillus ferrooxidans, 329 330 Acid mine drainage wastewater, 356 357 Acidobacteria, 31 35 Acidobacterium spp., 210 211 Acid Red 18 degradation, 237 Acinetobacter sp., 47 49 Acorus calamus, 249 Actinobacteria, 31 35, 205 206, 211 212 Actinomycetes, 174 175 Activated sludge system, 313 314 cyclic multistage reactors with, 314 multistage biological reactors with, 313 314 Active/inactive microbial biomass, 382 Acyl homoserine lactones (AHLs), 153 154, 155f, 344 Adapted versus nonadapted strains, 296 298

ADMI. See American Dye Manufacturer’s Institute (ADMI) Adsorption processes, 74, 177, 395, 408, 429 431 and electrostatic binding, 181 heavy metal pollutants, 72 Advanced attached biological treatment systems, 316t Advanced oxidation process (AOP), 74 Advanced treatment. See Tertiary treatment AEOs. See Alkylphenol ethoxylates (AEOs) Aeration/oxygenation, 46 47 Aerobic and anaerobic systems, 73, 75 Aerobic degradation, 263 264 of pollutants, 176 Aerobic granular sludge (AGS), 158 159 Aerobic heterotrophic bacteria, 349 Aerobic metabolism, 200 202 Aerobic treatment system, of wastewater, 312 314, 313t and anaerobic treatment systems, 313t attached growth system, 314 316 secondary treatment, 315 317 attached growth system, 315 316 cyclic activated sludge system, 316 integrated fixed film activated sludge system, 317

455

with membrane bioreactor, 317f membrane bioreactor system, 317 suspended growth system, 316 317 suspended growth system, 312 314 Aeromonas hydrophila, 289, 297 300 Aeromonas sp., 290 294 AFFR. See Anaerobic fixed film reactor (AFFR) AF systems. See Anaerobic Filter (AF) systems Agarose gel electrophoresis, 207 210 AGS. See Aerobic granular sludge (AGS) AHLs. See Acyl homoserine lactones (AHLs) α 2 Hydroxy carboxylate siderophores, 346 347 AIs. See Aromatic autoinducer signals (AIs) Algae bacterial consortium, 176 Algae-mediated azo dye degradation, 300 301 Algal bacterial consortium, 341 342 advantage of, 342 applications in wastewater treatment, 351 362 emerging contaminants removal, 359 361 metal removal, 356 357 nutrients removal, 353 356

456 Index Algal bacterial consortium (Continued) organic matter removal, 357 359 removal of refractory compounds, 361 362 types of reactors, 351 353 biofilm, 356 357 capability, 365 coculture, growth of, 347 348 COD and nutrient removal with, 354t emerging contaminants (ECs) investigated with, 360t energy generation, 362 365 algal biohydrogen production, 363 algal lipid production, 363 364 microbial fuel cell reactor using, 363 364 extracellular organic compounds, 343 interactions in, 348f maturation ponds with, 352 metal removal with, 357t performance, 343 refractory compounds investigated with, 361t in suspended system, 353 symbiotic process, 342 351 biomass growth in, 351 exchange of information in form of bioactive compounds for, 343 345 factors affecting, 349 351 inhibit microbes in vicinity, 345 346 quorum sensing, 344 345 stimulate microbes in vicinity, 346 348 wastewater treatment with, 365 366 Algal biohydrogen production, 363 Algal lipid bioaccumulation, 363 364 Algal lipid production, 363 364 Alicyclobacillus, 210 211 alkBFGHJKL operon, 41 43 alkB gene, 41 43

Alkylbenzene sulfonate degradation pathways, 264 Alkylphenol ethoxylates (AEOs), 172 173 Alphaproteobacteria, 28 29, 31 35, 211 212 Alternanthera philoxeroides, 140 Alternative electron acceptors, 73 Alteromonos sp., 345 346 Amensalism, 342 American Dye Manufacturer’s Institute (ADMI), 139 Ammonia, 351 Ammonium distribution coefficient, 11 Ammonium oxidizing bacteria (AOB), 7 amoA gene, 205 206 Amplified ribosomal DNA restriction analysis (ARDRA), 28 29, 207 208 Anaerobic ammonium oxidation (anammox) bacteria, 2 autotrophic growth behavior of, 2 autotrophic sulfate reducing, 3 4 cell immobilization. See Cell immobilization entrapment of, 2 3, 9 10 gel immobilization advantageous, 8 immobilization, 8 gel materials used for, 9 11 technology, commercialization of, 13 17 and metabolic process, 3 5 metabolism, 4 5, 4f plodding growth of, 2 3 start-up time of, 8 suspended growth, 2 3 Anaerobic and facultative ponds, 282 Anaerobic enzyme-mediated machinery, 288 289 Anaerobic Filter (AF) systems, 320 Anaerobic fixed film reactor (AFFR), 318f, 320

Anaerobic metabolism, 200 202 Anaerobic microbial fermentation, 445 Anaerobic treatment systems, of wastewater, 314 aerobic treatment system and, 313t, 315 317 representation of, 319f secondary treatment, 317 320 anaerobic fixed film reactors, 320 expanded granular sludge bed, 319 up-flow anaerobic sludge bed, 318 319 types of, 318 320 Anaerolineae, 31 35 Analyte-induced (AI) reporter gene, 154 155 Analytical methods, in azo dye degradation, 294 295 Anammox. See Anaerobic ammonium oxidation (Anammox) bacteria Anammoxoglobus propionicus, 3 Anammoxoglobus sulfate, 3 4 Anammoxosome membrane, 4 5 Anode material, in MFC, 245 Anodic biofilm accelerates, 243 Anthracophyllum discolor, 332 333 Anthranilate biodegradation, 160 Anthropogenic sources, 392 Antioxidants, 449 450 AOB. See Ammonium oxidizing bacteria (AOB) Aquatic ecosystem, 285 286 Aquatic macrophytes, 138 Aquatic systems, treatment using, 277 278 Aqueous Pb biorecovery, 411 421 Aqueous solubility, 133 Arabidopsis cells, 214 215 Arabidopsis thaliana, 152, 425 426 Arbitrarily primed polymerase chain reaction, 209 210 Archaeal phyla, 31 35 Arcobacter, 31 35

Index ARDRA. See Amplified ribosomal DNA restriction analysis (ARDRA) ARISA. See Automated ribosomal intergenic spacer analysis (ARISA) Aromatic autoinducer signals (AIs), 344 Aromatics, 248 249 compounds, 91, 110 dye compounds, 138 Arsenate diffusive transport, 399 Arsenic, 391, 425 426 chemistry of, 395, 396t in groundwater, 391 health effects, 393 nanofiltration of, 397 401 oxidation of, 397 398 permissible limit, 393 394, 394f removal from water/wastewater, methods of, 395 396 membrane technology, 396, 397f sources, 391 393, 392f anthropogenic, 392 natural, 393 species, 395 transport models for, 400t Arsenic-bearing minerals, 393 Arsenicosis, 393 Arsenite (As(III)) rejection, 398 399 Artificial incorporation technique, 7 8 Ascomycetes, 87 88 Aspergillus niger, 429 431, 447 448 Atelocyanobacterium thalassa, 346 Atrazine concentration, 245 246 Attached growth system, 314 316 Autoinducers (AI), 151 152 Automated ribosomal intergenic spacer analysis (ARISA), 209 Autotrophic sulfate, 3 4 Azo dye degradation, 285 286 bacteria, 287 296 analytical methods, 294 295 mechanism, 288 289

phases of treatment, 289 290 recent studies in, 290 294 as source, 287 288 by toxicity tests, efficiency analysis, 295 296 biodegradation, computational inputs in, 296 300 adapted versus nonadapted strains, 296 298 in silico analysis as a valuable tool, 298 300 comprehensive representation of, 286f decolorization, 289 fungi, yeast, and algae-mediated, 300 301 Green HE4B, 138 physical and chemical methods, 286 287 Azo dye, microbial bioremediation, 425 bioremediation, 427 432 characteristics, 428t classification, 426 427 effects of environmental parameters on, 432 438 decolorization by genetically modified organisms, 438 dye concentration, 434 electron donor, 435 436 oxygen, 433 434 pH, 435 redox mediator, 437 438 redox potential, 436 437 structure, 435 436 temperature, 433 reactive group, 426 427 reduction, 437f role of environmental parameters on, 427 432 types of, 428t Azoreductases, 287 289, 294, 433 Azospirillum, 374, 378, 380 382 Azospirillum lipoferum, 87

B Bacillus, 174 175, 205 208, 210 211, 374 376 Bacillus cereus strain, 39 41 Bacillus laterosporus, 289

457

Bacillus lentus B1377, 289 Bacillus licheniformis, 87, 354 355 Bacillus pumilus, 139 140 Bacillus sp., 239 240, 289, 294, 299 Bacillus subtilis, 47 49, 87, 299 Bacteria, 136 137, 153, 168 algae and. See Algal bacterial consortium azo dye degradation, 287 296 analytical methods, 294 295 mechanism, 288 289 phases of treatment, 289 290 recent studies in, 290 294 source, 287 288 by toxicity tests, efficiency analysis, 295 296 B-culture, 411 413 concentration and removal efficiency for, 419t consortia characterization of, 411 412 identification of, 412 feed on organic carbon sources, 341 342 flocculent, 353 Gram-negative bacteria, 153 154 Gram-positive bacteria, 153 microbial bioremediation, 174 175 Pb-resistance in, 411t PBZ1 and B510 strain, 380 with photosynthetic organisms, interaction, 341 342 quorum sensing in. See Quorum sensing (QS) symbiotic growth of, 347 348 two-component system, 157 Bacterial and fungal strains, 73t Bacterial artificial chromosomes, 265 268 Bacterial biosensors, 157 Bacterial community abundance, 262 268 biodegradation pathway in degradation of organic compounds, 262 265

458 Index Bacterial community abundance (Continued) functional metagenomics, 265 268 Bacterial species of genera, 174 175, 359, 376 378 plant growth promoting rhizobacteria, 374 376 Rice (Oryza sativa L.), 374 Bacterial systems, 75, 357 358 Bacteria strains, 263 264 Banana, 328 Bardenpho technology, 313 314 Barren/freshwater rinse method, 279 280 Basidiomycete I-62 (CECT 20197), 90 91 Basidiomycetes, 87 88 Batch biofilter granular reactor (SBBGR) system, 280 B-consortium, 420 B-culture bacteria, 411 413 concentration and removal efficiency for, 419t Beneficial biotechnological products, 182 183 Benzene, toluene, ethylbenzene, and xylenes (BTEX), 23 27, 205 206 Benzidine, 110 Benzotriazole, 172 173 Benzylsuccinate synthase, 41 43 BESs. See Bioelectrochemical systems (BESs) β-proteobacteria, 205 206 Beta-ketoadipate pathways, 264 Betaproteobacteria, 31 35 Betaproteobacteria, 263 264 Bioaccumulation, 382 Bioactive compounds, for symbiosis, 343 345 quorum sensing, 344 345 Bioattenuation, 70 Bioaugmentation (BA), 23 25, 47 49, 70 method, 199 200 Bioavailability, 200 202 of iron, 346 347 of pollutants, 68 role of, 68 70

Biobed, 319 Biobulk, 314 Biochar anode, 247 Biochemical and molecular mechanisms of Pbresistance, 411t Biochemical oxygen demand (BOD), 139, 342 Biochips. See Microarrays Biodegradation, 69, 158 160, 431 computational inputs in, 296 300 in degradation of organic compounds, 262 265 of hydrocarbons, 41 43, 49 50 by indigenous microbes, 49 50 of organic compounds, 28 of organic pollutants, 200 202 of pollutants, 68 of polycyclic aromatic hydrocarbons, 68 69, 107 109 Biodegradation-based remediation, 23 25 Biodiesel production, 353 Bioelectrochemical systems (BESs), 135 137, 145 146, 154 155 Biofabrication of bifunctional macromolecules, 183 184 Biofilm formation, 7, 151 152, 158 160, 344 345 and granulation process, 8 Biofilters, 314 Biofuel production process, 175 176 Biogenic uraninite nanoparticles, 329 330 Biohydrogen production, 363 Bioinformatics, 183, 298 sequencing, 29 30 Bioleaching process, 383 Biological oxygen demand (BOD), 257 258, 280 281, 314 315 Biological wastewater treatment methods, 277 BIOLOG system, 41 Biomineralization process, 383 Bionanotechnology, 183 184

Bioreactor digestion, 159 Biorecovery, of aqueous lead by local Pb-resistant organisms, 411 421 Bioremediation, 68 70, 133, 197 199, 325, 425 426 application of, 234 235, 425 426 of azo dye. See Azo dye, microbial bioremediation bacterial growth, 49 50 by B-consortium, 420 best strategy for, 197 199 biodegradation, 69 classification of, 234 235 concept of, 70 of contaminated soil, 106 control and optimization of, 427 429 of dyes, 76t effectiveness of, 133, 144 efficacy, 429 431 of environmental pollutants, 204f application of laccases, 106 115 explanatory mechanism of, 433f ex situ. See Ex situ bioremediation factors affecting, 49 50 of heavy metals, 71 72 mechanisms of bioremediation of Pb, 409 410 of metal-contaminated sediment, 243 microbial bioremediation of waste sludge accelerated bioremediation, 44 49 microbial electrochemical technology, 235 molecular biology methodologies, 438 nanomaterials. See Nanomaterials nanoparticles mimicking and functioning as enzyme for, 331t nanotechnology and, 325 333 native microbial populations in, 25 26

Index notion of, 68 of oil hydrocarbons, 47 49 of organic waste, 258 of organopollutants, applications of fungal laccases for, 108t of Pb, 408 409 and performance, 410t on petroleum-contaminated samples with high TPH content, 37t of petroleum refinery sludge carcinogenic pollutants in, 26 27 composition and hazard, 26 27 harmful chemicals, 27 heterogeneous nature, 23 25 hydrocarbon-associated environments microbiology, 28 43 microbial bioremediation of waste sludge, 43 49 physicochemical properties, 26 27 physicochemical technologies for, 44 45 for pollution control, 234 235 principles of, 429 431 quorum sensing, 158 159 role of genetic engineering in, 76 78 through microbial systems biology, 77 78 in situ. See In situ bioremediation of soil, 332 333 microbial bioremediation, 332 333 nanomaterials based bioremediation, 333 phytoremediation, 332 surfactants, 68 69 systems, 39 41 technologies, 44 for organic and polyaromatic compounds, 265 268 types, 173 180, 174f enzymatic bioremediation, 179 microbial bioremediation, 173 175

mixed cell culture system, 176 phycoremediation, 175 176 phytoremediation, 176 179 vermiremediation, 179 180 zooremediation, 179 using native microbiomes, 49 50 waste, 26 27 of wastewater, 159 160 Biosensing activity in microbial biosensors, 154 155 Biosensors, 154 155, 157 Biosorption, 382 Biostimulation (BS), 23 25, 46 47, 70, 199 200 Biosurfactants, 68 69 application, 180 Biosynthesized nanoparticles, 325 326, 332 Biotic factors, 431 Biotin (vitamin B7), 347 Biotower, 317 Biotransformation, 382 383 Bisphenol A (BPA), 166 168 inorganic organic clays, 167 photocatalytic method of, 168 photodegradation, 168 “Blue baby syndrome”, 276 277 Blue laccases, 99 Blue multicopper oxidases, 86, 99 Blue oxidases, 97 98 B. malcolmii, 142 BOD. See Biochemical oxygen demand (BOD); Biological oxygen demand (BOD) Brasicca rapa, 140 Brevibacillus laterosporus MTCC 2298, 294 Brevibacillus sp., 174 175 Brilliant Red X-3B decolorization, 237 B510 strain, 380 BTEX. See Benzene, toluene, ethylbenzene, and xylenes (BTEX) Burkholderia sp., 378 Burkholderia sp. DW2-1, 158 159 Butylated hydroxyanizole, 450 451 Butylated hydroxytoluene, 450 451

459

By-products, mango pulp processing industries, 444

C Calcium alginate (CA), 327 Calcium chloride, 10 Callus and suspension culturing methods, 142 Calvin Benson cycle, 347 Candida tropicalis, 300 301 Carbonaceous polymers, 347 348 Carbon-based nanomaterials, 326 328 Carbon dioxide, 349 Carbon nanotubes (CNTs), 326 327 Carbon:nitrogen:phosphorus (C:N: P) ratio, 351 Carbon-to-nitrogen (C/N) ratio, 2 Carboxymethyl-β 2 cyclodextrin polymer modified Fe3O4 (CDpoly-MNPs) nanoparticles, 330 331 Carboxymethyl cellulose (CMC), 10 11 CARD-FISH. See Catalyzed reporter deposition fluorescence in situ hybridization (CARDFISH) Carpobrotus rossii, 209 Carrier size, 7 CASS. See Cyclic activated sludge system (CASS) Catalytic efficiency constant, 95 Catalyzed reporter deposition fluorescence in situ hybridization (CARDFISH), 212 213 Catechin gallate, 444 Cathode, in sediment-MFC, 242 CCD approach. See Central composite design (CCD) approach CDpoly-MNPs nanoparticles. See Carboxymethylβ 2 cyclodextrin polymer modified Fe3O4 (CDpolyMNPs) nanoparticles Cell immobilization, 2 3, 5 8, 6f application of, 12 13 full-scale application, 12

460 Index Cell immobilization (Continued) artificial technique for, 5 different approaches for, 5 8 biofilm formation, 7 gel entrapment, 7 8 granulation, 5 7 Central composite design (CCD) approach, 399 401 Central nervous system, 407 CFB bacteria. See Cytophaga/ Flavobacterium/Bacteroides (CFB) bacteria Chemical oxidation, 169 Chemical oxygen demand (COD), 136 137, 139, 257 258, 285 286, 314, 318 320, 349 350 Chemical surfactants, 180 Chemical wastewater treatment methods, 277 Chemisorption, 383 Chemistry of arsenic, 395, 396t Chimeric enzymes, 183 Chitosan-polypropylene imine dendrimer, 330 Chlorella pyrenoidosa, 301 Chlorella sp., 278 Chlorella vulgaris, 301 Chlorine, 315 Chondroitin lyase, 343 344 Chromatography, 294 295 Chromohalobacter salaxigenes sp., 220 223 Chromosomal and extrachromosomal DNA, 184 Chrysopogon zizanioides, 249 Circular plastic media, 315 316 Cis-dichloroethene (cDCE), 212 213 Citrobacter sp., 383 Clavariopsis aquatic, 113 Clostridium botulinum, 412 CLPP, 41 CMC. See Carboxymethyl cellulose (CMC) C:N:P ratio. See Carbon:nitrogen: phosphorus (C:N:P) ratio CNTs. See Carbon nanotubes (CNTs)

COD. See Chemical oxygen demand (COD) Cofactor auxotrophy, 347 348 Cognate receptor, 153 Collimonas spp., 210 211 Colony forming unit (CFU), 415f Color Index, 426 427 Cometabolic bioremediation, 186 Cometabolism, 200 202, 431 Commensalism, 342 Community proteogenomics, 218 219 Community proteomics, 218 219 Comprehensive Environmental Response, Compensation and Liability Act, 391 Concentration polarization (CP) film theory, 398 399 Condensed tannins (CT), 444 Conjugation, 178 Constructed wetlands (CWs), 138 139 Constructed wetlands-microbial fuel cells (CW-MFC), 235 241 for boron removal, 238 for degradation of hydrocarbons, 238 hybrid system of, 236, 240 241, 240t practicability of, 238 tetracycline and sulfamethoxazole in, 239 upflow configuration, 237 Contact beds, 314 Contaminants, 4, 167, 170 172, 180 organic, 183 184 removal of, 180f Contaminated environments, 199 202, 323 325 Conventional bioremediation, 145 146 Conventional chromophores, 426 427 Conventional wastewater treatment processes, 311 312 primary treatment, 312 grit chambers, 312

primary settlers or sedimentation, 312 screen, 312 secondary treatment, 312 314 aerobic treatment system, 312 314, 313t anaerobic treatment system, 313t, 314 tertiary treatment, 315 Copper-resistant system, 157 Coriolopsis gallica, 107 109, 113 114 Coriolus zonatus, 98 Corrinoid, 218 C. reinhardtii, 347 348 Crenarchaeota, 31 35 Crude oil biodegradation, 26 27, 31 35 Cryptococcus (Filobasidiella) neoformans, 87 88 Cultivation methods, 215 216 Culture-based techniques, 203 Culture-independent gene-targeted metagenomics approach, 258 259 Culture-independent methods, 203 215 nonpolymerase chain reaction based molecular techniques, 212 215 fluorescence in situ hybridization, 212 213 guanine plus cytosine content, 213 214 microarrays, 214 215 polymerase chain reaction based molecular techniques. See Polymerase chain reaction based molecular techniques Culture-independent molecular methods, 38 39 Cupriavidus, 210 211 C. vulgaris, 354 355, 364 365 CW-MFC. See Constructed wetlands-microbial fuel cells (CW-MFC) CWs. See Constructed wetlands (CWs) Cyanide, 279 280

Index Cyanobacteria, 301, 341 342, 346 Cyathus bulleri, 112 Cyclic activated sludge system (CASS), 316 Cyclic multistage reactors with activated sludge, 314 Cyclic tubular photobioreactors, 352 353 CYP153 genes, 41 43 Cytochrome P450, 183 Cytophaga/Flavobacterium/ Bacteroides (CFB) bacteria, 212 213, 343 344 Cytoplasmic response receptors, 153

D Dairy industry, 280 281 Debaryomyces polymorphus, 300 301 De-centralized system, 152 Decolorization, 297 298 of Acid Red 18, 237 by algae, 301 azo dye degradation, 289 of dyes, 75, 110 111 by genetically modified organisms, 438 of scarlet RR dye, 249 of synthetic dyes, 75 76, 110 113 Deeper roots system, 140 Deferribacteres, 31 35 Degradation. See also Biodegradation azo dye. See Azo dye degradation of Brilliant Red X-3B, 237 of organic pollution/compounds, 262 aerobic pathway, 263 264 anaerobic pathways, 264 265 of PAH, 248 249 of polycyclic aromatic hydrocarbons, 169 of xenobiotic compounds, 107 109 Degradation and/or detoxification, of waste

bioremediation and role of bioavailability, 68 70 biodegradation, 69 in situ and ex situ bioremediation, 70 surfactants, 68 69 of pollutants, 70 76 dyes, 73 76 heavy metal pollutants. See Heavy metals limitations and future prospect, 78 role of genetic engineering in bioremediation, 76 78 toxic pollutant, 76 Degradative genes, 184 Dehalococcoides mccartyi, 218 Dehalococcoides strains, 211 212 Delisea pulchra, 344 345 Deltaproteobacterial members, 31 35 Denaturing gradient gel electrophoresis (DGGE), 28 29, 41 43, 205 206 Dendritic polymeric nanoparticles, 330 Dendroremediation, 140 Dense root system, 177 Designer microbe and plant approach, 184 185 Desulfobacteraceae, 41 43 Desulfomicrobium baculatum, 410 Desulfosarcina cetonica, 264 265 Desulfosarcina Desulfococcus group, 212 213 Desulfosporosinus spp., 329 330 Detection system, 157 Detoxification. See Degradation and/or detoxification, of waste DGGE. See Denaturing gradient gel electrophoresis (DGGE) Dichloroethylene (DCE), 200 202 Diclofenac, 170 171 Dietary fiber supplements, 451 2,6-Dimethoxyphenol, 95 Dimethylsulfoniopropionate (DMSP), 344 345 Dioxygenases, 41 43, 200 202, 263

461

Direct enzymatic reduction, 382 383 Direct leaching, 383 Dissolved oxygen (DO), 238, 349 Distillery waste, 198f Diverse bacterial groups, 41 43 DL-alpha-tocophenol, 451 DMSP. See Dimethylsulfoniopropionate (DMSP) DNA chips. See Microarrays DNA microarray, 214 215 DNA-SIP. See DNA stable-isotope probing (DNA-SIP) DNA stable-isotope probing (DNA-SIP), 210 211 DO. See Dissolved oxygen (DO) Domestic and industrial wastewater, 357 358 Donnan steric pore model (DSPM), 397 399 Drinking water distribution systems (DWDS), 212 213 DSPM. See Donnan steric pore model (DSPM) dsrAB genes, 41 43 Dual-chamber MFCs, 136 137 Duckweed species, 277 Dunaliella sp., 278 Dutch beet sugar firm, 318 319 DWDS. See Drinking water distribution systems (DWDS) Dyes, 73 76 biological treatment methods, 75 76 compounds, 140 141 decolorization, 75, 110 111, 140 degradation, 136 137, 249 enzymes for bioremediation of, 76t physical and chemical removal of, 74, 74f in situ/ex situ treatment of, 138 139

E Earthworms, 279 EAs. See Ellagic acids (EAs)

462 Index EB irradiation. See Electron beam (EB) irradiation Ecofriendly approach, 326 327 Ecofriendly waste management technique, 258 Economic sector, heavy metals in, 70 71 Ecosystem, heavy metals’ effect on, 70 71 Ectomycorrhiza, 89 90 EDC. See Endocrine disrupting chemicals (EDC) Edo activity, 216 217 Efflux proteins, 157 EGSB. See Expanded granular sludge bed (EGSB) 18S rDNA, 205 Eisenia fetida, 278 Electrobioremediation, 134 135 Electrochemical cells, 136 Electrohydrogenesis systems, 135 136 Electrokinetic remediation, 182 Electron acceptor, nitrite as, 3 4 Electron beam (EB) irradiation, 142 144 advantage, 143 144 full/field-scale applications, 143 integration, 144 in wastewater treatment, 143 Electron donor, 435 436 Electron paramagnetic resonance (EPR) spectroscopy, 97 98 Electron transfer (ET), 104 Electrostatic binding, 181 Elevated heavy metal concentrations, effect of, 417 418 Ellagic acids (EAs), 444 applications of, 448 451 bioconversion of tannin to, 445 446 formation of, 449f mango pulp processing industries, 443 microbes involved in production of, 447 448 production, 446t, 447 prospective applications of, 450f Ellagitannins (ETs), 444, 447

bioconversion of, 448t degradation and utilization, 447 by fungal enzymes, 447 Emerging contaminants (EC), 233 removal, algal, 359 361 in situ removal of, 249 soil-MFC for degradation of, 244 245 Emerging organic contaminants (EOCs), 132 Emerging pollutants, 165, 172 173 bioremediation types, 173 180, 174f enzymatic bioremediation, 179 microbial bioremediation, 173 175 mixed cell culture system, 176 phycoremediation, 175 176 phytoremediation, 176 179 vermiremediation, 179 180 zooremediation, 179 Bisphenol A, 166 168 inorganic organic clays, 167 photodegradation, 168 hospital effluents as source of, 171 172 biological treatment, 172 pharmaceutical wastes, 170 171 photodegradation, 171 sludge treatment, 171 polychlorinated biphenyls, 169 170 biological degradation, 170 halogenated organic compounds technology for, 170 polycyclic aromatic hydrocarbons. See Polycyclic aromatic hydrocarbons (PAHs) sources and effects of, 167t types of, 166f Emerging techniques, 180 186, 180f adsorption and electrostatic binding, 181 bioinformatics, 183

biosurfactants, application of, 180 cometabolic bioremediation, 186 designer microbe and plant approach, 184 185 electrokinetic remediation, 182 genetic engineering, 184 immobilization, 181 metagenomics, 182 183 nanotechnology, 183 184 plant microbe symbiosis manipulation of, 185 porous matrix and encapsulation, entrapment in, 181 182 protein engineering, 183 rhizosphere engineering, 185 Encapsulation, 182 Endocrine disrupting chemicals (EDC), 166 167 Endophytes, 184 185 Energy generation, 362 365 algal biohydrogen production, 363 algal lipid production, 363 364 microbial fuel cell reactor using algal bacteria interaction, 363 364 Enterococcus, 207 208 Enterococcus faecium, 412 Enterococcus sp., 412 Entrapment of anammox, 2 3, 9 10 Environment, 165 166 emerging pollutants. See Emerging pollutants PAHs in. See Polycyclic aromatic hydrocarbons (PAHs) Environmental biotechnology, 154 160 biofilm formation, 159 160 bioremediation, 158 159 heavy metal detection, 154 157 hydrocarbon remediation, 160 pathogen detection, 158 Environmental contaminants. See Contaminated environments

Index Environmental parameters on azo dye degradation, 432 438 dye concentration, 434 electron donor, 435 oxygen, 433 434 temperature, 433 Environmental pollutants, 429 431 bioremediation of, 204f laccases application for, 106 115 Environmental pollution, 133, 138, 197 199, 275 Environmental product, 67 68 Environmental proteomics, 218 219 Enzymatic bioremediation, 179 Enzyme-based bleaching technologies, 113 114 Enzyme-based system, 75 Enzyme dye interaction, 299 300 Enzyme inhibitors, 97 Enzymes, catalytic power of, 183 Enzymes, defined, 85 Enzymes, plant-based, 140 141 EOCs. See Emerging organic contaminants (EOCs) Epoxy resins, 166 167 EPR spectroscopy. See Electron paramagnetic resonance (EPR)spectroscopy EPS. See Extracellular polymer substances (EPS) Epsilonproteobacteria, 31 35 EroS, 343 344 Escherichia coli, 280 Essential nutrients, scarcity of, 25 26, 49 50 ET. See Electron transfer (ET) 3-Ethylbenzthiazoline-6-sulfonate (ABTS), 95 97, 103 104, 105f Ethylenediaminetetraacetic acid, 97 Eukaryotic phytoplankton, 346 347 Euryarchaeota, 31 35 Eutrophication, 353 354 Exopolysaccharides, 158 159

Expanded granular sludge bed (EGSB), 319 Experimental approach, 413 414 Ex situ bioremediation, 44, 70, 76, 144, 182, 197 199, 234 235, 325, 429 431 microbial communities involved in, 202 215 Extended Nernst Plank (ENP) equation, 397 Extracellular environment, 153, 433 434 Extracellular fungal enzymes, 175 Extracellular plant growth promoting rhizobacteria, 376 377 Extracellular polymer substances (EPS), 352 353 Extracellular signaling molecules, 344 345 Extreme environments, 28 Exudates, algal inhibition of microbes in vicinity, 345 346 stimulate microbes in vicinity, 346 348 cofactor auxotrophy, 347 348

F Facultative interaction, 342 Fatty acid compounds, 345 346 Feed concentration, 398 Fe-hydrogenases, 363 Fermentative microbes, 179 Ferric oxide, 241 242 FGAs. See Functional gene arrays (FGAs) Field-scale applications, 250 251 Filamentous fungi, 300 Filtered algal exudates, 346 Filtering technique, 2 Filtration, 397 398 Fingerprinting techniques/methods, 205 206, 209 210 Firmicutes, 31 35 FISH. See Fluorescence in situ hybridization (FISH) Flavobacterium, 343 Flavodon flavus, 110, 113

463

Flocculent bacteria, 353 Fluorescence in situ hybridization (FISH), 28 29, 212 213 Fomes sclerodermeus, 112 113 Fourth-order Runge Kutta method, 399 Franconibacter pulveris DJ34, 39 41 Freshwater algae, 175 176 Full-scale application of anammox, 12, 17 Functional annotation databases, 30 31 Functional-based metagenomics, 216 217 Functional gene arrays (FGAs), 214 215 Functional metagenomics approach, 261, 265 268, 268f Function-directed remediation approach, 146 Fungal degradation, 175, 300 301 polycyclic aromatic hydrocarbons (PAHs), 169 Fungal laccases, 86, 88, 94 97, 106, 113 for bioremediation of organopollutants, 108t ligninolytic systems in, 90 molecular weights of, 94 95

G GACB-MFC. See Granular activated carbon-microbial fuel cell (GACB-MFC) Gallic acid, 88 89 Gallotannins, 444 Gammaproteobacteria, 31 35, 211 212, 263 264 Ganoderma lucidum, 89 90, 95 97, 110 111 Gas pocket, 8 Gastrointestinal colic, 408 Gauss Newton algorithm, 399 401 Gelatin, 328 Gel electrophoresis, 207 208 Gel immobilization advantageous, 8

464 Index Gel immobilization (Continued) materials used for, 9 11 polyethylene glycol gel (PEG), 11 polyvinyl alcohol and polyvinyl alcohol/sodium alginate, 10 waterborne polyurethane, 10 11 techniques, 5 Gene-targeted metagenomics, 258 262, 268 269 advantage of, 259 261 functional screening, 269 function-based approach, 261 qPCR approach in, 262 sequences-based approach, 259 261 whole genome metagenome sequencing, 261 262 Genetically modified (GM) organisms, 425 426, 438 plants, 181, 185 Genetic engineering, 91 94, 184 in bioremediation, 76 78 Genetic programming (GP) model, 398 Genome sequencing, 77, 220 Genomic DNA. See Metagenomics Genomic polymorphism, 209 210 Genomics, 77 78 Geobacillus thermoglucosidasius, 264 265 Geobacter metallireducens, 242 243, 329 330 Geobacter sulfurreducens, 243 Germ-free axenic cultures, 343 344 Glandularia pulchella, 139 Glycoproteins, 94 95, 346 GO. See Graphene oxide (GO) Gordonia sp. nov. Q8, 39 41 Gram-negative bacteria, 344, 382 acyl homoserine lactone in, 153 154, 155f strains, 263 264 Gram-negative rhizosphere bacteria, 374 376 Gram-positive bacteria, 153, 154f, 168, 344

sensor of peptides in, 153 species, 47 49 two-component system in, 153 Granular activated carbonmicrobial fuel cell (GACBMFC), 136 137 Granulation process, 5 7 biofilm formation and, 8 Granules settling ability, 5 7 Graphene-based nanomaterials, 327 328 Graphene oxide (GO), 327 328, 328f Gratuitous metabolism. See Cometabolism Green algae, 363 Greengenes, 29 30 Greenhouse gas emission, 247 248 Green light, 349 350 “Green Liver Model”, 178 Green product, 67 68 Green technology, 325 326, 332 Grit chambers, primary function of, 312 Groundwater, arsenic in, 391 G. thermoglucosidasius C56-Y593, 299 Guanine plus cytosine content (G 1 C content), 213 214

H HAEs. See Hydrocarbon-associated environments (HAEs) Hagen Poiseuille equation, 399 Hairy root culturing, 142 Halogenated organic compounds (HOCs), 170 Halophilic/halotolerant bacteria, 169 Halophyte, 138 Haptophytes, 346 Harmful chemicals, 27 HAT. See Hydrogen atom transfer (HAT) Hazardous waste cleanup systems, 174 175 Health effects for arsenic, 393 394

Heavy metals, 238, 240 241, 243, 247, 323 325 adsorption, 328f concentrations, 417 418 detection, 154 157 microbial quorum-sensing (QS) systems for, 156t effect on human health, 324t exchangeable state, 378 functional groups, 326 327 in iron and manganese oxidation state, 378 in organic state, 378 phytoextraction of, 138 pollutants, 70 73, 323 325 adsorption, 72 bioremediation of, 71 72 biosorbence for, 73t microorganisms for detoxification of, 73 natural and anthropogenic sources responsible for, 72f sources, 71 in urban sectors, 70 71 removal of, 235 236, 248t, 249 250, 329t in residual state, 378 in soil, plant growth promoting rhizobacteria role, 377 378 stress, 380 381 using hybrid microbial electrochemical technologies, 250 251 xenobiotics and, 248t Herbaceous plants, 140 141 Herbaceous species, 177 178 Herbicides concentration, 245 246 Heterologous production of laccases, 91 94 Heterotrophic biofilm, 239 240 Heterotrophic microorganisms, 200 202 Hexachlorobenzene, 244 245 Hexachlorocyclohexane (HCH), 220 223 Hexadecane mineralization, 160 Hexahydroxydyphenic acid, 444 445 Hexavalent chromium, 249 250 Hibiscus furcellatus, 141

Index High-molecular-weight organic compounds, 142 143 High-rate algal ponds (HRAP), 351 352 High-temperature natural processes, 393 High-throughput sequencing methods, 220 223 Hindrance factors, 398 Histidine kinase receptor, 153 HMETs. See Hybrid microbial electrochemical technologies (HMETs) HOCs. See Halogenated organic compounds (HOCs) Horizontal tubular photobioreactors, 352 Hospital effluents as source of emerging pollutants, 171 172 HRAP. See High-rate algal ponds (HRAP) HRT. See Hydraulic retention time (HRT) HT. See Hydrolysable tannin (HT) Hybrid microbial electrochemical technologies (HMETs), 250 251 applications of, 251 Hybrid systems of constructed wetlandsmicrobial fuel cells, 236, 240 241, 240t toxic wastewater control, 282 for xenobiotic compounds, 248t Hydraulic retention time (HRT), 236, 317 318, 342, 350 Hydrocarbon-associated environments (HAEs), 23 25 Betaproteobacteria, 31 35 distribution of microbial phyla in, 33f Epsilonproteobacteria, 31 35 metabolic processes in, 39f for microbial community distribution, 36t microbiology of, 28 43 Proteobacteria in, 31 35

Hydrocarbon-metabolizing microorganisms, 31 35 Hydrocarbonoclastic microbes, 23 25 Hydrocarbons, 23 25 biodegradation of, 49 50 contamination, 160 degradation, 41 43 initial concentrations of, 49 50 natural degradation of, 44 remediation, 160 Hydrodynamics of vertical photobioreactors, 352 Hydrogen production, 363 Hydrogen atom transfer (HAT), 104 Hydrogenotrophic methanogens, 23 25 Hydrolysable tannin (HT), 444 Hydrolyzing enzymes, 357 358 Hydrophilic groups, 68 69 Hydrophobic pollutants, 180 Hydroponic systems, 141 Hydroquinone, 437 438 Hydroxide anion, 95 97 Hydroxyapatite nanomaterials, 328 1-Hydroxybenzotriazole (HBT), 107 109 Hydroxymate siderophores, 346 347 Hypocrea lixii, 331

I IAA. See Indole-3-acetic acid (IAA) IFAS system. See Integrated fixed film activated sludge (IFAS) system Illumina sequencing technologies, 211 212, 220 223 Immobilization, 408 anammox application of, 12 13 for evaluation of different factors, 14t cell culture, 433 of cells. See Cell immobilization of chromium, 249 250 of genetically modified, 181 of laccases, 104 106

465

mechanism, 181 of microbial biomass, 3 techniques, emerging environmental pollutants, 181 technology, commercialization of, 13 17 Independent variables, 398 Indian Oil Cooperation Limited (IOCL), 26 27 Indigenous microbial community, 44 Indigenous microbial populations, 186 Indirect leaching, 383 Indole-3-acetic acid (IAA), 343 344 Industrial and municipal wastewaters, 280 281 Industrialization, 70 71 Industrial revolution, 144, 275 Industrial sector, heavy metals in, 70 71 Industrial wastewater, 257 258, 357 358 Initial algae:bacteria ratio, 350 Inorganic carbon (IC), concentration of, 13 Inorganic organic clays (IOCs), 167 Inorganic pollutants, 179 180, 197 202, 257 258 In silico techniques, 298 300 In situ bioremediation, 44 45, 70, 76, 144, 197 199, 202 215, 325, 429 431 basic methods of, 197 199 electron acceptors, 241 242 future scope of research, 250 251 hybrid microbial electrochemical technologies, 250 251 microbial communities involved in, 202 215 techniques, 234 using microbial electrochemical technologies, 235 250 constructed wetlandsmicrobial fuel cells (CWMFC), 235 241

466 Index In situ bioremediation (Continued) plant-microbial fuel cells (plant-MFC), 246 250 sediment-microbial fuel cells (sediment-MFC), 241 243 soil-microbial fuel cells (soilMFC), 243 246 Instrumental automatism, 209 Integrated fixed film activated sludge (IFAS) system, 317 Internal transcribed spacer (ITS), 205 International Atomic Energy Agency, 142 143 Intracellular and extracellular oxidoreductive enzymes, 288 Intracellular plant growth promoting rhizobacteria, 376 377 Intrinsic biodegradation, 46 47 Intrinsic bioremediation, 197 199 Invasive approach, 197 199 IOCL. See Indian Oil Cooperation Limited (IOCL) IOCs. See Inorganic organic clays (IOCs) Ion Torrent sequencing platform, 29 30, 217 218 Ipomoea palmate, 140 Irradiated wastewater, 143 Isoelectric points (pI), 94 95 Isoenzymes, 94 95 Isophthalate/3-carboxybenzoate, 264 Isozyme production, 89 90 “Itai-Itai” disease, 276 277 ITS. See Internal transcribed spacer (ITS)

K Klebsiella pneumoniae, 412 Klebsiella pneumoniae strain L17, 136 137

L Laboratory-scale system design, 413 417 Lab-scale batch reactors, 139 140, 411

LacA/ABTS system, 112 113 Laccase mediator system (LMS), 101 104, 106, 113 114 Laccase(s), 75 76, 86, 97f application for bioremediation of environmental pollutants, 106 115 in biotechnology, 107f decolorization of synthetic dyes, 110 113 degradation of xenobiotic compounds, 107 109 to develop ecofriendly processes, 114 115 pulp and paper industry, 113 114 treatment of industrial effluent, 113 of Bacillus licheniformis, 87 biochemical properties of, 94 97 effect of inhibitors on activity, 97 effect of pH and temperature on activity, 95 97 kinetic properties, 95 catalytic performance, 94, 100f classification of, 100 101 copper atoms and ligand distance, 99f in decolorizing azo dyes, 289 distribution and physiological functions, 87 88 genes, expression of, 90 94 immobilization of, 104 106 isoenzymes, 89 90, 94 95 kinetic parameters of, 96t limitations and future prospects, 115 116 mechanism, 98f mode of action, 97 100 Myceliophthora thermophila, 91 94 in organic syntheses, 114 115, 115f production, 87 106 cultural and nutritional conditions for, 89 91 heterologous, 91 94

screening of laccaseproducing fungi, 88 89 by white-rot fungi (WRF) and recombinants, 92t of redox mediators, 102f, 104f Rhus vernicifera, 86 87 temperature stability of, 95 97 of T. hirsuta, 110 Land-based wastewater treatment systems, 282 Laser-based fluorescence detection system, 209 LasR/LasI QS system, 157 Lead (Pb), 407 biochemical and molecular mechanisms of, 411t biorecovery by local Pb-resistant organisms, 411 421 characterization of bacterial consortia, 411 412 concentration and removal efficiency for B-culture bacteria, 419t concentration measurements, change in, 419f current techniques employed for, 409t metabolic activity (MA) rate distribution, 416f microbiological and kinetic study, 413 417 case studies of varied operating conditions, 417 421 colony forming unit (CFU) count rate distribution, 415f elevated heavy metal concentrations effect, 417 418 laboratory-scale system design and experimental approach, 413 414 substrate composition, effect of, 420 421 Zn(II) or Cu(II) ions on Pb(II) bioprecipitation effect, 418 419 minimum inhibitory concentration of, 419 420 nitrate rate distribution, 416f

Index precipitate identification, 412 413 rate distribution, 415f reduction, 410, 417 remediation, 408 410 bioremediation, 408 410 conventional methods for, 408 research on species used for, 410t Lead poisoning, 408 Legionella pneumophila, 212 213 Lemna, 277 Lentinus edodes, 448 Leptospirillum group II bacteria, 219 220 Leucoanthocyanidins, 444 Library-based metagenomic method, 215 216 Library-based targeted metagenomics, 215 216 Life-supporting ecosystems, 197 199 Light, symbiotic systems, 349 350 Ligninolytic systems in fungi, 86, 90, 169 Lignin peroxidase (LiP), 75 76 Lignin phenolic groups, 97f, 99 Lipid synthesis, 349 350 Listeria, 289 llumina-based platforms, 29 30 LMS. See Laccase mediator system (LMS) Lumbricus rubellus, 278 LuxI system, 153 154 LuxR system, 153 154, 344 LuxS/AI-2 system, 344 Lysinibacillus sphaericus SK13, 299 300 Lysogeny broth (LB), 412 413

M Magnaporthe oryzae, 379 380 Magnetic nanoparticles, 330 Malachite green, 112 113 Manganese peroxide (MnP), 332 333 Mangifera indica L. Anacardiaceae, 443 Mango city, 443

Mango pulp processing industries, 443 by-products, 444 composition of, 443 444 ellagic acids. See Ellagic acids (EAs) Marasmius quercophilus, 94 95 Marasmius querocophilus strain 17, 89 90 Mass spectrometry, 218 219 Mass transfer coefficient, 398 399 Maturation ponds, 282, 352 Maximum power density (MPD), 236 MBBR. See Moving bed bioreactor (MBBR) MBR system. See Membrane bioreactor (MBR) system mcrA gene, 41 43 MECs. See Microbial electrolysis cells (MECs) Membrane bioreactor (MBR) system, 280 281, 317 Membrane-bound receptor, 153 Membrane separation process, 396 Membrane technology, 396 comparison of, 397f MET. See Microbial electrochemical technology (MET) Metabolic activity (MA) rate distribution, 416f Metabolomics, 219 220 Metagenomics, 77, 182 183, 199 200, 258 259, 260f approaches in bioremediation, 224 225 bacterial community abundance, 262 268 of contaminated sites, 221t culture-based techniques, 203 culture-independent methods. See Culture-independent methods ex situ bioremediation, 197 199, 202 215 functional-based metagenomics, 216 217 gene-targeted. See Gene-targeted metagenomics

467

metabolomics, 219 220 metaproteomics, 218 219 metatranscriptomics, 218 and NGS technology, 268 269 sequence-based metagenomic, 217 218 sequencing strategies, 220 in situ bioremediation, 197 199, 202 215 studies, methods used for, 259 262 function-based metagenomics approach, 261 gene-targeted (sequencebased) approach, 259 261 whole genome metagenome sequencing, 261 262 Metal corrosion, heavy metals by, 323 325 Metal ions, 91 Metallic nanoparticles, 329 Metal-organic framework (MOF) materials, 323 325 Metal removal, algae for, 356 357 Metal toxicity, 380 381 Metaproteomics, 78, 218 219 Metatranscriptomics, 218, 264 metE genes, 347 348 Methanogens, 31 35 Methanosarcina gene, 218 219 Methemoglobinemia, 276 277 metH genes, 347 348 Methionine synthase, 347 348 2-methoxy-4-aminoazobenzene, 295 296 3-Methoxy-4-aminoazobenzene, 295 296 Methylobacterium oryzae, 378 Methylobacterium sp. strain BJ001, 140 141 Methyl-tert-butyl ether (MTBE), 178, 238 M10EXG, 264 265 MFCs. See Microbial fuel cells (MFCs) MG-RAST, 29 30 MIB. See Morphogenesis-inducing bacteria (MIB) Michaelis Menten constant, 95 Michaelis Menten kinetics, 434

468 Index Microalgae, 175 176 biomass, 362 363 treatment using, 278 Microalgae bacterial consortia, 176 Microarrays, 77 78, 214 215 Microbe plant symbiosis, 185 Microbes, 224 225, 325 326, 417 Microbial biodegradation, 210 211 Microbial bioremediation, 173 175, 200 202 bacterial bioremediation, 174 175 mycoremediation, 175 of waste sludge, 43 49 Microbial biosensors, 154 155 “Microbial cell factories”, 43 44 Microbial cell immobilization, 3 Microbial cells/enzymes, 181 184 Microbial communities, 199 200, 212 213 composition characterization of, 31 35 of hydrocarbon-associated environments, 28 43, 36t of oil-contaminated sediment, 29 30 sulfate-reducing, 41 43 culture-based techniques, 203 culture-independent methods. See Culture-independent methods FISH analysis, 212 213 fluctuations, 207 208 genetic diversity of, 258 259 metagenomics. See Metagenomics molecular approaches used for monitoring of, 204f native and nonnative, 258 and plant microbes interactions, 146 in situ and ex situ bioremediation, 202 215 structure and function, 220 223 Microbial diversity, 31 35, 39 41, 46 47 characterization of, 28 29

of refinery sludge, 38 39 Microbial electrochemical technology (MET), 235 in situ bioremediation using, 235 250 constructed wetlandsmicrobial fuel cells (CWMFC), 235 241 plant-microbial fuel cells (plant-MFC), 246 250 sediment-microbial fuel cells (sediment-MFC), 241 243 soil-microbial fuel cells (soilMFC), 243 246 Microbial electrolysis cells (MECs), 135 136, 235 Microbial enzymes, 85, 179 Microbial fuel cells (MFCs), 135 137, 235, 364 Microbial-mediated bioremediation, 197 199 next-generation sequencing technologies to, 220 223 Microbial phyla, distribution of, 33f Microbial processes, 197 199 Microbial quorum-sensing (QS) systems, 154, 156t, 161 Microbial strains, 170 Microbial surfactants, 68 69 Microbial systems biology, 77 78 genomics, 77 metagenomics, 77 proteomics and metaproteomics, 78 transcriptomics, 77 78 Microbial wastewater treatment, 5 8 Micrococcus glutamicus NCIM2168 (MG), 290 294 Micrococcus luteus, 451 Micromonas, 347 Microorganisms/microbes, 71 73, 173 174, 202, 344, 378, 425, 427 429 aerobic pathway degradation by, 263 anaerobic, 39 41 azo dye. See Azo dye, microbial bioremediation

in bioconversion of ellagitannin, 448t biodegradation of SOCs by, 133 degradation of organic pollution using, 262 degradative power of, 49 50 for detoxification of heavy metals, 73 exogenous and indigenous populations, 199 200 in extreme environments, 28 growth and activity, 432 hydrocarbon metabolism by, 28 immobilization, 5 biofilm formation, 7 in biological treatment systems, 5 7 gel entrapment, 7 8 granulation, 5 7 isolation and cultivation of, 39 41 metabolic characteristics of, 432 metabolic machinery of, 264 metabolic versatility and gene diversity, 184 metabolism, 135 nanomaterials, 331 nutritional versatility of, 425 in production of ellagic acid, 447 448 rhizospheric, 178 179 Microsystis aeruginosa, 354 355 MICs. See Minimum inhibitory concentrations (MICs) Miller lysogeny broth (LB), 412 414 Mineral aquifer interaction, 393 Minimum inhibitory concentrations (MICs), 419 420 Ministry of Drinking Water and Sanitation, 391 MiSeq sequencing, 220 223 Mixed cell culture system, 176 Mixed culture system, 343 Mixotrophic algae, 363 364 Mixotrophic strains, 175 176 MnP. See Manganese peroxide (MnP) Modeling arsenic transport, 400t

Index Donnan steric pore, 397 399 genetic programming, 398 multiple solute, 399 401 nanofiltration membranes, 397 401 Spiegler Kedem steric hindrance, 398 399 MOF materials. See Metal-organic framework (MOF) materials Molecular biology, 199 200, 205, 431 432 Monoazo dye, 435 436 Monooxygenases, 200 202, 263 Moraxella bovis, 410 Morphogenesis-inducing bacteria (MIB), 343 344 Moving bed bioreactor (MBBR), 315 317 MTBE. See Methyl-tert-butyl ether (MTBE) Multicopper oxidases, 106 Multiple metagenomic analysis methods, 218 Multiple solute model, 399 401 Multi-walled carbon nanotubes, 326 327 Municipal wastewater treatment, 282, 311 312 Mutant strain, 160 Mutualism, 342 Mycelial biosorption, 408 409 Myceliophthora thermophila laccase, 91 94 Mycobacterium chlorophenolicum, 429 431 Mycobacterium frederiksbergense LB501, 135 Mycoremediation, 175 Myriad of organisms, 341 342

N N-Acyl-homoserine-lactone (AHL), 152 NADH-DCIP, 290 294 Nanobioremediation, 183 184 Nanofiltration (NF), 395 of arsenic, 397 401 membranes modeling, 397 401

using genetic programming (GP), 398 Nanomaterials, 323 325 as adsorbents, 333 334 used for removing pollutants, 326 331 biogenic uraninite nanoparticles, 329 330 carbon-based nanomaterials, 326 328 dendrimers, 330 hydroxyapatite nanomaterials, 328 metallic nanoparticles, 329 microorganisms, 331 nanozymes, 331 polymeric nanocomposites, 330 331 Nanoobjects, 183 184 Nanoparticles, 325 326, 333 different methods for, 327f of Euphorbia macroclada, 333 mimicking and functioning as enzyme for bioremediation, 331t synthesized with biological agents for heavy metals, 329t Nanopore sequencing system, 261 262 Nanoremediation, 183 184, 325 326 Nanoscale sized materials, 323 325 Nanotechnology, 183 184 and bioremediation, 325 333 Nano zero-valent iron (nZVI), 239 240 Nanozymes, 331 Naphthalene, 68 69 Naphthalenic acids, 172 173 National Engineering and Environmental Laboratory (INEEL), 211 212 Native microbial community, 258 Native microflora, heavy metals by, 70 72 Natural and artificial materials, 7 8

469

Natural attenuation (NA), 23 25, 197 199 rates of, 199 200 Natural bioremediation process, 332 Natural degradation, of hydrocarbon, 44 Natural microorganisms, 429 431 Natural surfactants, 68 69 nbCOD. See Not biologically degradable (nbCOD) N-decanoyl-L-HSL (C10HSL), 158 159 Neuronal encephalopathy, 408 Neutral As(III) molecule, 398 399 Next generation artificial enzymes. See Nanozymes Next-generation sequencing (NGS), 23 25, 29 30, 38 39, 217 218, 258 261, 268 269 NGS. See Next-generation sequencing (NGS) Nicotinamide adenine dinucleotide (NADH)-generating systems, 436 Nitrate amendment, 25 26 Nitrate production, 4 5 Nitrate rate distribution, 416f Nitrate reductase (narG), 41 43 Nitration anammox biofilm process, 7 Nitrification denitrification (N DN), 2 Nitrifying bacteria, 172 Nitrite-oxidizing bacteria, 349 350 Nitrobenzene, 238 239 Nitrogen, 1 2 gas bubbles formation, 13 and organic acid, 3 pollution, 276 removal process, 2 Nitrogen-fixing bacteria, 364 365 Nitrogen removal rate (NRR), 12 Nonadapted strains, adapted versus, 296 298 Nonionic surfactants, 68 69, 172 173

470 Index Nonland-based systems, 282 Nonligninolytic fungi, 169 Nonnative microbial community, 258 Nonnative strains, 297 298 Nonpolymerase chain reaction based molecular techniques, 212 215 fluorescence in situ hybridization, 212 213 guanine plus cytosine content, 213 214 microarrays, 214 215 Nonsteroidal antiinflammatory drugs, 170 171 Nonylphenol polyethoxylates (NPEOs), 113 Nonylphenols, 113 Not biologically degradable (nbCOD), 357 358 Novel functional screening system, 216 217 N-(3-oxotetradecanoyl)-Lhomoserine lactone, 344 345 N-3-(oxyhexanoyl)-homoserine lactone, 153 154 2-n-pentyl 4-quinolinol, 345 346 NPEOs. See Nonylphenol polyethoxylates (NPEOs) NRR. See Nitrogen removal rate (NRR) N-terminal amino acid sequences, 99 Nutraceuticals, 450 451 Nutrients removal, 353 356 nZVI. See Nano zero-valent iron (nZVI)

O Obligate interaction, 342 Octanol water partition coefficient, 200 202 OECD. See Organization for Economic Cooperation and Development (OECD) Oenanthe javanica, 239 Oil-contaminated sediment, 29 30 Oil hydrocarbons, 47 49 Oily sludge, 26 27, 39 41

biodegradation, 43 44 microbial population of, 41 waste management, 44 45 Oligopeptides, 153 On-site bioremediation, 429 431 Operational taxonomic units (OTUs), 30 31, 209 Organic acids (propionate), 3 Organic compounds, 133, 141, 357 358 Organic matter removal, 357 359 Organic pollutants, 133, 179 180, 197 202, 257 258 by plants, 138 in situ microbial degradation of, 200 202 Organic syntheses, 114 115, 115f Organization for Economic Cooperation and Development (OECD), 27 Organosulfur compounds, 343 344 Oryza sativa L. See Rice (Oryza sativa L.) Oscillateria tenuis, 301 Oscillatoria sp., 342 343 Ostreococcus, 347 OTUs. See Operational taxonomic units (OTUs) Oxidases, 85 86. See also Blue multicopper oxidases Oxidation of arsenic, 397 398 of organic electron donors, 435 436 Oxidation reduction processes, 73, 429 431 Oxidized forms of metals, 382 383 Oxidoreductases, 85 86 Oxygen, 85 86, 343 azo dye degradation, 433 434

P PAB. See Precultured anammox bacteria (PAB) PacBio system, 261 262 P. aeruginosa, 157 159, 380 PAH-degrading genes, 41 43

PAHs. See Polyaromatic hydrocarbons (PAHs); Polycyclic aromatic hydrocarbons (PAHs) Paraconiothyrium variabile, 111 Parasitism, 342 PA4778 regulator, 157 Partial nitrification, application of, 12 13 Paspalum crinitum, 139 140 Passive bioremediation, 197 199 Pathogen detection, 158 Pathogenicity/virulence, 158 Pb. See Lead (Pb) Pb(II) stock solution, 413 414 PBZ1 strain, 380 PCA. See Principal component analysis (PCA) PCBs. See Polychlorinated biphenyls (PCBs) PEG. See Polyethylene glycol (PEG) gel PEM. See Proton exchange membrane (PEM) Penta-O-galloyl-glucoside (PGG), 450 451 Perfluorinated compounds (PFCs), 172 173 Permissible limit for arsenic, 393 394, 394f Peroxidases, 140 Persistent organic pollutants (POPs), 170 Pesticide degradation, 174 175 Pesticide remediation, 174 175 Petroleum hydrocarbon, 242 Petroleum hydrocarbon-impacted environments, 28 29 Petroleum microbiology, 23 25, 38 39 Petroleum refinery sludge, bioremediation, 26 27 carcinogenic pollutants in, 26 27 composition and hazard, 26 27 harmful chemicals, 27 heterogeneous nature, 23 25 hydrocarbon-associated environments microbiology, 28 43

Index microbial bioremediation of waste sludge, 43 49 Petroleum refinery wastewater, 137 PFC. See Photomicrobial fuel cell (PFC) PGG. See Penta-O-galloylglucoside (PGG) PGPR. See Plant growth promoting rhizobia (PGPR) pH, 429 431 algal bacterial symbiosis, 351 azo dye degradation, 435 effect on laccases, 95 97 Phanerochaete chrysosporium, 90, 300 Pharmaceutical wastes, 170 171 photodegradation, 171 sludge treatment, 171 Phase II metabolism, 178 Phenanthrene, 68 69 Phenolic compounds, 101 Phenolic substrates, 95 97 Phenol-oxidizing enzymes, 99 Phenol Red, 103 104 Phenols, 75 Phlebia brevispora, 113 114 Phlebia tremellosa, 113 114 Phosphorus, 353 354 Photobioreactor systems, 352, 365 366 Photochemical methods, 286 287 Photodegradation, 168 pharmaceutical waste, 171 polycyclic aromatic hydrocarbons, 169 Photolysis of water, 363 Photomicrobial fuel cell (PFC), 364 365 Photosynthesis, 341 342, 351, 363 Photosynthetic oxygenation, 343 by algae, 357 359 Phragmites australis, 139, 220 223 Phycoremediation, 175 176 Phycosphere, 342, 362 363 PhyloChips, 214 215 Phylogenetic investigation of communities by

reconstruction of unobserved states (PICRUSt), 30 31 Phylogenetic oligonucleotide arrays, 214 215. See also PhyloChips Physicochemical remediation methods, 43 44 Phytoaccumulation. See Phytoextraction Phytodegradation, 138, 178 Phytoextraction, 138, 176 177 Phytoreactors and constructed wetlands, 138 139 Phytoremediation process, 176 179, 236, 425 426. See also Phytotechnology phytodegradation, 178 phytoextraction, 176 177 phytostabilization, 177 178 phytovolatilization, 178 rhizofiltration, 177 rhizoremediation, 178 179 soil bioremediation, 332 types of, 177t Phytostabilization, 138, 177 178 Phytotechnology (phytoremediation), 138 142, 145 hydroponic systems, 141 phytoreactors and constructed wetlands, 138 139 plant enzymes and metabolites, 140 141 plant microbe phytoremediation, 139 140, 146 plant tissue culturing, 141 142 Phytotechnology-based approach, 138 Phytotoxic substances, 376 377 Phytotransformation, 178. See also Phytodegradation Phytovolatilization, 138, 178 Pichia pastoris, 107 109 Picochlorumatomus, 355 356 PICRUSt. See Phylogenetic investigation of communities by reconstruction of

471

unobserved states (PICRUSt) Planctomycetes, 3 Planomicrobium, 39 41 Plant-assisted remediation, 176 Plant-dependent phytoremediation, 140 Plant enzymes and metabolites, 140 141 Plant growth promoting rhizobia (PGPR), 178 179, 373, 375t different genera of, 374 377, 376t extracellular, 376 377 general characteristics, 374 376 in heavy metals dynamics in soil, 377 378 nutrient dynamics and pathogen biocontrol by, 379f Pantoea sp. strain EA106, 381 382 in remediation of environment, 380 383 for rice plant growth promotion, 375t role in controlling pathogens in rice, 379 380 Plant microbe phytoremediation, 139 140, 146 Plant microbe symbiosis, manipulation of, 185 Plant-microbial fuel cells (plantMFC), 246 250, 248t aromatics and hydrocarbons removal, 248 249 greenhouse gas emission reduction, 247 248 heavy metals removal, 249 250 hybrid technology, 247 for xenobiotics and heavy metal removal, 248t Plant tissue culturing, 141 142 Plasmid screening, 297 298 Plastic media, 315 316 Plastic media trickling filter, 317 Platinum, 171 172 Pleurotus ostreatus, 94 95, 100 Pleurotus pulmonarius, 89 90

472 Index Pleurotus sajor-caju, 90 91, 110 111 Po. grandiflora, 139 Po. grandiflora-based phytoreactor, 139 140 Pollutants, 68, 115, 135, 431 432 aerobic degradation of, 176 anodic degradation of, 239 240 bioavailability of, 68 biodegradation of, 68 bioremediation of, 425 426 cometabolism of, 200 202 degradation, 70 76, 183 184 detoxification, 70 76 heavy metals, 323 325 prominent types, 70 Polyacrylamide gel electrophoresis, 208 Polyamide membrane, 398 399 Polyamide NF membranes, 397 398 Polyaromatic hydrocarbons (PAHs), 23 25, 69, 132, 248 249 degradation, 362 removal efficiencies of, 361 362, 361t Poly(vinyl alcohol)-boric acid, 112 Polycarbonate plastics, 166 167 Polychlorinated biphenyls (PCBs), 140 141, 169 170, 242 243 biological degradation, 170 halogenated organic compounds technology for, 170 Polycyclic aromatic hydrocarbons (PAHs), 39 41, 86, 107, 168 169, 332 333 archaea for degradation of, 169 bacterial catabolism of, 168 chemical oxidation and photodegradation, 169 degradation of, 168 fungal degradation, 169 halophilic/halotolerant bacteria, 169 Polyethylene glycol (PEG) gel, 9, 11, 13, 17 Poly(MAalt-MVE)-g-PLA/ODAMMT, 112

Polymerase chain reaction (PCR)based approach, 259 261 Polymerase chain reaction based molecular techniques, 205 212 amplified ribosomal DNA restriction analysis, 207 208 automated ribosomal intergenic spacer analysis, 209 denaturing gradient gel electrophoresis, 205 206 quantitative polymerase chain reaction, 211 212 random amplified polymorphic DNA, 209 210 restriction fragment length polymorphisms, 207 208 single-strand conformation polymorphism, 206 207 stable-isotope probing, 210 211 temperature gradient gel electrophoresis, 205 206 terminal restriction fragment length polymorphisms, 208 Polymeric nanocomposites, 330 331 Polymeric proanthocianidins, 444 Polyphenol oxidases, 100 101 Poly R-478, 88 89, 141 Polysaccharides, 346 Polyvinyl alcohol (PVA), 10, 17 Polyvinyl alcohol/Sodium alginate (PVA/SA), 9 10, 12 Polyvinyl chloride products, 407 Pond systems, 351 352, 365 POPs. See Persistent organic pollutants (POPs) Porous matrix and encapsulation, entrapment in, 181 182 Potentially toxic elements, 275 Power generation performance, 238 P. putida PCL1444, 140 141 P. putida PML2, 140 141 Precipitate identification, 412 413 Precultured anammox bacteria (PAB), 12 Precursor oligopeptide, 153 Primary metabolism, 200 202

Primary settlers, primary function of, 312 Primary treatment conventional wastewater treatment processes, 312 grit chambers, 312 primary settlers or sedimentation, 312 screen, 312 wastewater treatment plant, 315 Principal component analysis (PCA), 205 206 Pristine environment, 131 132 Prokaryote eukaryote interactions, 152 Prokaryotic cell to cell signaling, 151 Prokaryotic microorganisms, 23 25 Protein engineering, 183 Proteobacteria, 31 35, 263 264 in petroleum hydrocarbon-rich environments, 31 35 Proteomics, 78 Proteus vulgaris NCIM-2027 (PV), 290 294 Proton exchange membrane (PEM), 235 Prototheca zopfii, 175 176 Pseudoalteromonas sp., 347 Pseudolagarobasidium acaciicola, 111 112 Pseudomonas, 211 212, 263 264, 343 Pseudomonas aeruginosa, 344 345 Pseudomonas delafieldii, 329, 333 Pseudomonas fluorescens strains, 75 Pseudomonas maltophilio, 381 382 Pseudomonas putida, 139 140, 295 296, 355 356, 358 359 Pseudomonas species, 158, 174 175, 211 212, 264 265, 290 294, 374 376, 379 380 Pseudomonas stutzeri, 217 218

Index Pulp and paper industry, potential applications in, 113 114 Pump and treat system, 234 235 PVA. See Polyvinyl alcohol (PVA) Pycnoporus sanguineus, 110 Pyrene degradation, 362 Pyrethroid hydrolyzing enzyme, 182 183 Pyrosequencing technology 454 (Roche), 29 30

Q QIIME. See Quantitative Insights Into Microbial Ecology (QIIME) qPCR. See Quantitative polymerase chain reaction (qPCR); Quantitative realtime PCR (qPCR) QS. See Quorum sensing (QS) Quantitative Insights Into Microbial Ecology (QIIME), 30 31 Quantitative polymerase chain reaction (qPCR), 211 212 Quantitative real-time PCR (qPCR), 262 “Quantum quenching”, 159 160 Quinone hydroquinone round pairs, 437 438 Quorum sensing (QS), 151, 344 345 acyl homoserine lactone in Gram-negative bacteria, 153 154, 155f in environmental biotechnology, 154 160 biofilm formation, 159 160 bioremediation, 158 159 heavy metal detection, 154 157 hydrocarbon remediation, 160 pathogen detection, 158 in Gram-positive bacteria, twocomponent system, 153, 154f limitations of, 161 mechanisms of, 152 154 peptides secretion for, 153 principles, 152

Quorum sensor-based detection, 152

R Radial diffusion, 178 Ralstonia pickettii, 210 211 Ralstonia solanacearum, 412 Random amplified polymorphic DNA (RAPD), 209 210 RAPD. See Random amplified polymorphic DNA (RAPD) RBBR. See Remazol Brilliant Blue R (RBBR) Reactive Black 5 (RB5), 110 111, 330 Reactive Red, 426 427, 427f Reactive Red 198 (RR198), 142, 330 Reactors, types of, 351 353 Real-time PCR (RT-PCR), 211 212, 262 Real-time quantitative PCR (RT-qPCR), 211 212 Recalcitrant pollutants, 158 159, 169 170, 186, 242 243, 265 268, 285 286 Recalcitrant water pollutant, 166 167 Recombinant DNA techniques, 184 185 Redox mediators (RMs), 91 94, 101, 104f, 437 438 Redox reactions, 429 431 Reduced graphene oxide (r-GO), 327 328 Reduction, of lead (Pb), 410, 417 Refinery sludge bioremediation. See Petroleum refinery sludge, bioremediation heterogeneous nature, 23 25 microbial diversity of, 38 39 microbiomes, 51 petroleum, 26 27 waste, 26 27 Refractory compounds, removal of, 361 362 Remazol Blue and Black B textile dyes, 300 301

473

Remazol Brilliant Blue R (RBBR), 88 89, 110 111 Restriction fragment length polymorphism (RFLP), 207 208 Reverse osmosis technique, 280 281 RFLP. See Restriction fragment length polymorphism (RFLP) r-GO. See Reduced graphene oxide (r-GO) rGO-Au NPs. See r-GO supported Au nanoparticles (rGO-Au NPs) r-GO supported Au nanoparticles (rGO-Au NPs), 327 328 Rheum palmatum, 142 Rhizobacteria, 376 377 Rhizoctonia oryzae-sativae, 380 Rhizofiltration, 177 Rhizopus oligosporous, 448 Rhizoremediation, 139 140, 178 179 Rhizosphere engineering, 185 Rhizospheric bacteria, 184 185, 373 Rhizospheric microorganisms, 178 179 Rhodococcus genes, 168 Rhodocyclaceae (Betaproteobacteria), 31 35 Rhus vernicifera, 86 87 Ribosomal Database Project-RDP, 29 30 Ribosomal RNA (rRNA) gene, 205, 209 Rice (Oryza sativa L.), 373 Azospirillum, 374 PGPR role in controlling pathogens in, 379 380 plant growth promoting rhizobia, 375t production management, 373 374 Rice blast disease, 379 380 RMs. See Redox mediators (RMs) RNASeq data analysis, 264 RNA sequencing, 77 78

474 Index Roche 454 GS-FLX Titanium, 29 30 Roche 454 sequencing, 259 261 Root culturing (hairy roots), 142 Rumex crispus, 141

S Saccharomyces cerevisiae, 300 301 Sandy clay soil, 138 139 Sanger sequencing technology, 220 223 SBBGR system. See Batch biofilter granular reactor (SBBGR) system SBR. See Sequencing batch reactor (SBR) Scanning electron microscope (SEM), 413, 414f Scenedesmus sp, 355 356 Screening of laccase-producing fungal species, 88 89 Screens, primary function of, 312 S-culture, 411 412 SDM. See Site-directed mutagenesis (SDM) SDS. See Sodium dodecyl sulfate (SDS) SDS-PAGE. See Sodium dodecyl sulfate-polyacrylamide gel electrophoresis (SDSPAGE) Secondary plant metabolites (SPMEs), 140 141 Secondary treatment conventional wastewater treatment processes, 312 314 aerobic treatment system, 312 314, 313t anaerobic treatment system, 313t, 314 wastewater treatment plant, 315 320 aerobic secondary treatment, 315 317 anaerobic secondary treatment, 317 320 attached growth system, 315 316

cyclic activated sludge system, 316 integrated fixed film activated sludge system, 317 membrane bioreactor system, 317 suspended growth system, 316 317 Secreted root exudates, 246 247 Sedimentation, primary function of, 312 Sediment-EC. See Sediment electrolysis cell (sedimentEC) Sediment electrolysis cell (sediment-EC), 243 Sediment-MFC. See Sedimentmicrobial fuel cells (sediment-MFC) Sediment-microbial fuel cells (sediment-MFC), 241 243, 241f SEM. See Scanning electron microscope (SEM) Semiquantitative method, 205 206 Separation of ions, 397 Sequence-based metagenomics, 217 218 Sequences-based approach, 259 261 Sequencing batch reactor (SBR), 280, 313 314, 316 Sequencing technologies, 220, 224 225 Sequential anaerobic aerobic phases, 289 290 Sequential batch systems, 365 Shear stress, 8 Shewanella oneidensis, 154 155, 333 Shewanella sp. strain KMK6, 75 Shotgun metagenomic sequencing methods, 215 216, 220 223, 261 262 Siderophores, 346 347, 378 Signaling molecules, 153 SILVA, 29 30 SILVAngs, 30 31 Single chamber MFC, 136 137

Single-strand conformation polymorphism (SSCP), 206 207 Single-walled carbon nanotubes, 326 327 SIP. See Stable-isotope probing (SIP) Site-directed mutagenesis (SDM), 183 16S rDNAs gene, 205 206 16S rRNA genes, 23 25, 28 31, 205 206, 210 211, 214 216, 220 225, 412 SK-SHM. See Spiegler Kedem steric hindrance model (SK-SHM) S-layer, 382 Sludge retention time (SRT), 172 Sludge treatment, pharmaceutical waste, 171 Small-subunit ribosomal RNA (SSU rRNA) gene, 205 Sobering reality, 285 286 Social insects, 152 SOCs. See Synthetic organic compounds (SOCs) Sodium dodecyl sulfate (SDS), 244 245 Sodium dodecyl sulfatepolyacrylamide gel electrophoresis (SDSPAGE), 94 95 Soil aging of pollutants in, 134 bioremediation, 332 333 microbial bioremediation, 332 333 nanomaterials based bioremediation, 333 phytoremediation, 332 heavy metals dynamics in, 377 378 microorganisms, 178 179, 203 and water contamination, 198f Soil-borne plant pathogens, 374 Soil-microbial fuel cells (soilMFC), 243 246 Solar photovoltaic implementations, 407 Sol-gel method, 325 326

Index Solid liquid separation, 353 Solid retention time, 317 318 Solute solute interactions, 399 401 Sorption, 359 Sources anthropogenic, 392 arsenic, 391 393, 392f of heavy metal pollutants, 71 Sphingobacterium sp., 294 Sphingomonadaceae, 28 29 Sphingomonas, 28 29 Sphingomonas sp. L138, 135 Spiegler Kedem steric hindrance model (SK-SHM), 398 399 Spirodela, 277 SPMEs. See Secondary plant metabolites (SPMEs) SSCP. See Single-strand conformation polymorphism (SSCP) Stabilization ponds, 352 Stable-isotope probing (SIP), 28 29, 210 211 Stacked multilayered anode configuration, 245 Stand-alone CW, 237 Standard culture techniques, 203 Staphylococcus aureus, 408 409 Start-up time of anammox, 8 and immobilization techniques, 9t State-wise arsenic-affected habitations, 392f Stenotrophomonas, 207 208 Stone/plastic media trickling filter, 317 Streptococcus pneumoniae, 151 Streptomyces cerevisiae, 91 94 Streptomyces cyaneus, 87 Streptomyces lavendulae, 87 Structure-based virtual screening, 298 299 Structure function relationships, 183 Substrate composition, effect of, 420 421

Substrate concentrations, algal bacterial symbiosis, 351 Subsurface pollutants, 177 178 Sulfamethoxazole, 239 Sulfide, 415 416 Sulfitobacter sp., 343 344 Sulfonated anthraquinones, 141 Sulfonate groups, 435 436 Sulfurospirillum, 31 35 Sulfur, probable source of, 413 Suprametabolism, 140 141 Supranol Green, 140 Surfactants, 68 69 Suspended growth system, 312 314, 316 317 cyclic activated sludge system, 316 integrated fixed film activated sludge system, 317 membrane bioreactor system, 317 wastewater treatment, 2 3 Sustainable environment, 70 71 degradation and/or detoxification. See Degradation and/or detoxification, of waste Symbiotic pond system, 356 357 Symbiotic process/systems, 342 351 algae in, 359 bioactive compounds for, 343 345 biomass growth in, 351 EC removal by, 359 exudates, algal, 345 348 factors affecting, 349 351 carbon dioxide, 349 dissolved oxygen, 349 hydraulic retention time, 350 initial algae:bacteria ratio, 350 light, 349 350 pH and temperature, 351 substrate concentrations, 351 Synthetic dyes, 75 76, 110 decolorization of, 110 Synthetic laccase mediators, 103f Synthetic organic compounds (SOCs), 131 132, 144

475

biodegradation, 133, 144 bioelectrochemical systems/ technology, 135 137 commercial synthesis of, 131 132 electrobioremediation, 134 135 electron beam irradiation. See Electron beam (EB) irradiation molecular characteristics of, 134 phytotechnology (phytoremediation), 138 142 sequestration, 134 transformation products, 132 Synthetic surfactants, 68 69 Synthetic textile wastewater, 355 356 “Syntrophic methanogenic hydrocarbon degradation”, 23 25 Syntrophic organisms, 31 35 Syntrophobacteraceae, 41 43 Syntrophorhabdus aromaticivorans, 264 Syringaldazine, 95, 101 Systems biology (SB), 76 77

T Tannic acid, 88 89 Tannins, 444 bioconversion of, to ellagic acid, 445 446 classification of, 445f production, 446t types of, 444 445 TEAs. See Terminal electron acceptors (TEAs) Temperature algal bacterial symbiosis, 351 azo dye degradation, 433 Temperature gradient gel electrophoresis (TGGE), 205 206 Tensile strength of pulp, 113 114 Terminal electron acceptors (TEAs), 25 26, 41 43 Terminal restriction fragment length polymorphisms (TRFLP), 208

476 Index Terminal restriction fragment length polymorphisms (TRFLP) (Continued) analysis, 13 Terrestrial pollution, 138 Tertiary treatment conventional wastewater treatment processes, 315 wastewater treatment plants, 320 Tetracycline, 239 concentration, 359 Textile effluent treatment, 138 139 Textile industry, 110 and dyeing industries, 285 286 Textiles dyes, 141 Textile wastewater, 137 140, 432 TGGE. See Temperature gradient gel electrophoresis (TGGE) Thalluscin, 343 344 Thermomucorindicae seudaticae, 75 Thiamine monophosphate, 347 Thiobacillus denitrificans, 329 330 Thlaspi caerulescens, 205 206 Thymol, 158 Time-tested bioremediation technique, 235 236 Tissue cultured plants, 141 142 Total metagenomics DNA, 261 262 Total petroleum hydrocarbons (TPHs), 26 27, 245 degradation, 47 49 Total suspended solids, 257 258 Totyltriazoles, 172 173 Toxicity tests, bacterial dye degradation by, 295 296 Toxic metals, 73, 377 378, 380 381 Toxic wastewater control, 275 biological treatment methods, 277 chemical treatment methods, 277 conventional treatment techniques, 280 281 current interventions in, 277 281

hybrid systems, 282 other interventions in, 279 281 pollution, causes and effects of, 276 277 reuse, 281 282 treatment using aquatic systems, 277 278 using microalgae, 278 using vermifiltration, 278 279, 279f wetland plant scheme for, 278f types of, 276t TPHs. See Total petroleum hydrocarbons (TPHs) Traditional cultivation-based methods, 41 Trametes gallica, 89 Trametes hirsuta, 107 110 Trametes pubescens, 115 Trametes versicolor, 110 111 ATCC 200801, 112 IBL-04, 112 Trametes villosa, 91, 113 Transcriptomics, 77 78 Transform pollutants, 425 426 Transmembrane pressure, 398 Transport of ions, 398 T-RFLP. See Terminal restriction fragment length polymorphisms (T-RFLP) Trichloroethylene (TCE), 200 202 Trichoderma koningiopsis, 331 Trichoderma viride, 331 Trickling filters, 314, 317 2,4,6-Trinitrotoluene, 140 141 Triplicate batch reactors, 413 414 Trivalent chromium, 249 250 True laccase, 87 88 T. trogii, 110 Tubular photobioreactors, 352 Two-component system in Grampositive bacteria, 153 Type I copper (T1), 98 Typha angustifolia, 249 Typha latifolia, 238 Typha orientalis, 239 240 Tyrosinase, 289

U UASB. See Up-flow anaerobic sludge bed (UASB) Ultraviolet (UV) light radiation, 320 Uncultured bacteria, 203 Uncultured Microbiota, 199 200 Up-flow anaerobic sludge bed (UASB), 318 319, 318f Uraninite nanoparticles, 329 330 Urban sector, heavy metals in, 70 71 UV light radiation. See Ultraviolet (UV) light radiation UV Vis. See UV Visible spectroscopy (UV Vis) UV Visible spectroscopy (UV Vis), 97 98, 294 295, 299 300

V Varied operating conditions, case studies of, 417 421 effect of elevated heavy metal concentrations, 417 418 VBNC. See Viable but nonculturable (VBNC) Vegetation-based remediation process, 176 Vermiculture or worm farming, 278 Vermifiltration, treatment using, 278 279, 279f Vermiremediation, 179 180 Vertical tubular photobioreactors, 352 Veterinary drugs, 172 173 Viable but nonculturable (VBNC), 203 Vibrio fischeri, 343 344 Vibrio harveyi, 157 Virulence gene expression, 158 Vitamin B12, 347 348 Vitamins, 347 VOCs. See Volatile organics carbons (VOCs) Vogesella, 210 211 Volatile organics carbons (VOCs), 132 Volatilization process, 178

Index W Waste discharge, 443 disposal technologies, 425 from mango pulp processing industries, 443 stabilization ponds, 282 Wastewater treatment plants (WWTPs) design, 138 139, 311 312. See also Wastewater treatment processes evolution in, 311 312 primary treatment, 315 recent advances achieved in, 315 320 secondary treatment, 315 320 aerobic secondary treatment, 315 317 anaerobic fixed film reactors, 320 anaerobic treatment technology, 318 320 attached growth system, 315 316 expanded granular sludge bed, 319 suspended growth system, 316 317 up-flow anaerobic sludge bed, 318 319 tertiary treatment, 320

Wastewater treatment processes, 135 136, 311 312 anammox. See Anaerobic ammonium oxidation (anammox) bacteria applications of algal bacterial symbiosis in emerging contaminants removal, 359 361 metal removal, 356 357 nutrients removal, 353 356 organic matter removal, 357 359 removal of refractory compounds, 361 362 types of reactors, 351 353 conventional. See Conventional wastewater treatment processes domestic and industrial, 357 358 electron beam (EB) irradiation, 143 suspended growth, 2 3 using phytoremediation, 141 142 Waterborne polyurethane (WPU), 9 11 Water hyacinth plant systems, 277 Water-quality regulations, 165 White-rot fungi (WRF), 87 88, 94 95, 107, 110, 169 extracellular laccases of, 89 90

477

and recombinants, 92t Whole genome metagenome sequencing, 261 262 Windblown dust, 177 178 Wood rot fungi, 94 World Health Organization (WHO), 393 394 WPU. See Waterborne polyurethane (WPU) WRF. See White-rot fungi (WRF)

X Xanthomonas oryzae, 380 Xenobiotic compounds, 90, 200 202, 233, 264, 288 degradation of, 107 109 industrial pollution due to, 264 phytotransformation of, 178 plant-MFC hybrid systems for, 248t X-ray photoelectron spectroscopy, 249 250, 413

Y Yeast, 300 301 Yellow laccases, 99

Z Zn(II) or Cu(II) ions on Pb(II) bioprecipitation, 418 419 Zooremediation, 179 Zymogram, 89 90