Bioremediation of Industrial Waste for Environmental Safety: Volume II: Biological Agents and Methods for Industrial Waste Management [1st ed.] 978-981-13-3425-2;978-981-13-3426-9

Table of contents :
Front Matter ....Pages i-xxviii
Genetically Modified Organisms (GMOs) and Their Potential in Environmental Management: Constraints, Prospects, and Challenges (Gaurav Saxena, Roop Kishor, Ganesh Dattatraya Saratale, Ram Naresh Bharagava)....Pages 1-19
Advances in Bioremediation of Toxic Heavy Metals and Radionuclides in Contaminated Soil and Aquatic Systems (Evans M. Nkhalambayausi-Chirwa, Pulane Elsie Molokwane, Tshilidzi Bridget Lutsinge, Tony Ebuka Igboamalu, Zainab S. Birungi)....Pages 21-52
Phycoremediation: Algae as Eco-friendly Tools for the Removal of Heavy Metals from Wastewaters (Shamshad Ahmad, Arya Pandey, Vinayak Vandan Pathak, Vineet Veer Tyagi, Richa Kothari)....Pages 53-76
Recent Advances in Phytoremediation of Toxic Metals from Contaminated Sites: A Road Map to a Safer Environment (Mukesh Kumar Awasthi, Di Guo, Sanjeev Kumar Awasthi, Quan Wang, Hongyu Chen, Tao Liu et al.)....Pages 77-112
Emerging and Ecofriendly Technologies for the Removal of Organic and Inorganic Pollutants from Industrial Wastewaters (Gaurav Saxena, Surya Pratap Goutam, Akash Mishra, Sikandar I. Mulla, Ram Naresh Bharagava)....Pages 113-126
Constructed Wetlands: A Clean-Green Technology for Degradation and Detoxification of Industrial Wastewaters (Sardar Khan, Javed Nawab, Muhammad Waqas)....Pages 127-163
Nano-bioremediation: An Innovative Remediation Technology for Treatment and Management of Contaminated Sites (Ritu Singh, Monalisha Behera, Sanjeev Kumar)....Pages 165-182
Electro-bioremediation: An Advanced Remediation Technology for the Treatment and Management of Contaminated Soil (Sivasankar Annamalai, Maruthamuthu Sundaram)....Pages 183-214
Microbial Fuel Cell (MFC): An Innovative Technology for Wastewater Treatment and Power Generation (Mostafa Rahimnejad, Maryam Asghary, Marjan Fallah)....Pages 215-235
Functional Diversity of Plant Endophytes and Their Role in Assisted Phytoremediation (Angélica Leonor Guerrero-Zúñiga, Eugenia López-López, Aída Verónica Rodríguez-Tovar, Angélica Rodríguez-Dorantes)....Pages 237-255
Toxic Metals in Industrial Wastewaters and Phytoremediation Using Aquatic Macrophytes for Environmental Pollution Control: An Eco-Remedial Approach (Mansi Rastogi, Meenakshi Nandal)....Pages 257-282
Microalgae: An Eco-friendly Tool for the Treatment of Wastewaters for Environmental Safety (Jae-Hoon Hwang, Anwar Sadmani, Seung-Jin Lee, Keug-Tae Kim, Woo Hyoung Lee)....Pages 283-304
Phycoremediation: An Integrated and Eco-friendly Approach for Wastewater Treatment and Value-Added Product Potential (J. Umamaheswari, D. Saranya, S. Abinandan, Mallavarapu Megharaj, Suresh R. Subashchandrabose, S. Shanthakumar)....Pages 305-331
Pulp and Paper Mill Wastewater: Ecotoxicological Effects and Bioremediation Approaches for Environmental Safety (Izharul Haq, Abhay Raj)....Pages 333-356
Cadmium as an Environmental Pollutant: Ecotoxicological Effects, Health Hazards, and Bioremediation Approaches for Its Detoxification from Contaminated Sites (Sushila Saini, Geeta Dhania)....Pages 357-387
Cyanobacteria: The Eco-Friendly Tool for the Treatment of Industrial Wastewaters (Sharma Mona, Virendra Kumar, Bansal Deepak, Anubha Kaushik)....Pages 389-413
Plant-Microbe Interactions for Bioremediation and Phytoremediation of Environmental Pollutants and Agro-ecosystem Development (Akash Mishra, Shraddha Priyadarshini Mishra, Anfal Arshi, Ankur Agarwal, Sanjai Kumar Dwivedi)....Pages 415-436
Molecular Technologies for Assessment of Bioremediation and Characterization of Microbial Communities at Pollutant-Contaminated Sites (Sudhir Kumar Shekhar, Jai Godheja, Dinesh Raj Modi)....Pages 437-474
Biochar: A Sustainable Tool in Soil Pollutant Bioremediation (Chhatarpal Singh, Shashank Tiwari, Jay Shankar Singh)....Pages 475-494
Bioremediation of Melanoidins Containing Distillery Waste for Environmental Safety (Vineet Kumar, Ram Chandra)....Pages 495-529
Progresses in Bioremediation Technologies for Industrial Waste Treatment and Management: Challenges and Future Prospects (Ram Naresh Bharagava, Gaurav Saxena)....Pages 531-538

Citation preview

Ram Naresh Bharagava · Gaurav Saxena Editors

Bioremediation of Industrial Waste for Environmental Safety Volume II: Biological Agents and Methods for Industrial Waste Management

Bioremediation of Industrial Waste for Environmental Safety

Ram Naresh Bharagava • Gaurav Saxena Editors

Bioremediation of Industrial Waste for Environmental Safety Volume II: Biological Agents and Methods for Industrial Waste Management

Editors Ram Naresh Bharagava Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Microbiology (DM) Babasaheb Bhimrao Ambedkar (Central) University Lucknow, Uttar Pradesh, India

Gaurav Saxena Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Microbiology (DM) Babasaheb Bhimrao Ambedkar (Central) University Lucknow, Uttar Pradesh, India

ISBN 978-981-13-3425-2 ISBN 978-981-13-3426-9 https://doi.org/10.1007/978-981-13-3426-9

(eBook)

© Springer Nature Singapore Pte Ltd. 2020 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, express or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Singapore Pte Ltd. The registered company address is: 152 Beach Road, #21-01/04 Gateway East, Singapore 189721, Singapore

This book is truly dedicated to my parents for their unfailing patience, contagious love, forgiveness, selflessness, and endless support; my wife for trusting and believing in me; and my kids for always being a reminder that there is hope to move forward in life. Ram Naresh Bharagava This book is truly dedicated to my parents for their unfailing patience, contagious love, forgiveness, selflessness, and endless support and for nurturing and educating me to date. Without them I would not be the person I am today. Gaurav Saxena

Foreword

Safeguarding the environment and a sustainable future in the wake of rapid industrialization is a key challenge worldwide. Industries are the fundamental drivers in the world economy, but they can also be major polluters discharging potentially toxic and hazardous wastes into the environment. Conventional physico-chemical remediation techniques are known to be costly and environmentally destructive and can result in secondary pollution and disturb the natural environment. In contrast, bioremediation is a low-cost and eco-friendly alternative to conventional remediation technologies for the treatment and management of industrial wastes. Bioremediation has been identified by the US Environment Protection Agency (USEPA) to promote the sustainable development of our society with low environmental impact. The applications and scope of bioremediation technologies are everexpanding. For instance, the development of constructed wetlands (CWs) has revolutionized the way contaminated wastewater is treated and managed and its reuse in wastewater treatment facility. The Putrajaya wetland in Putrajaya city of Malaysia represents an excellent example of the commercial success of CWs in developing countries. Bioremediation of Industrial Waste for Environmental Safety: Biological Agents and Methods for Industrial Waste Management (Volume II), edited by Dr. Ram vii

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Foreword

Naresh Bharagava and Mr. Gaurav Saxena, is particularly useful to researchers working in the fields of bioremediation, phytoremediation, and waste management. All the chapters are contributed by leading experts in the field. This book provides excellent source materials on the different aspects of biodegradation and bioremediation of industrial wastes. It introduces the readers to the potential toxic effects of various pollutants in industrial wastes and their environmental impacts. It examines the role of biological agents (such as bacteria, fungi, algae, plants, microalgae, and genetically modified organisms) in biodegradation and biodetoxification of a variety of organic and inorganic pollutants (e.g., phenols, chlorophenols, petroleum hydrocarbons, polychlorinated biphenyls, organic solvents, azo dyes, pesticides, recalcitrant compounds, toxic metals, to name but a few), as well as their potential in electricity production, biofuel generation, and phytomining (where plants are used to recover valuable metals from contaminated sites). I appreciate the editor’s effort in editing this valuable book, which will go some way to make our planet greener and cleaner. I congratulate the book editors for bringing out this valuable compilation with up-to-date knowledge in the field of industrial waste bioremediation. I wish this book a great success as it is of great value to the stakeholders, including researchers, academicians, students, environmentalists, and policy makers. It also makes a fine companion to the volume I of this book title. Dr. Diane Purchase, Ph.D., FHEA, FIES, FIEnvSci Honorary Secretary, Committee of Heads of Environmental Sciences, UK Editor, Environmental Science and Pollution Research, a Springer Nature Journal Coordinating Editor, Environmental Geochemistry and Health, a Springer Nature Journal Professor of Environmental Biotechnology Department of Natural Sciences School of Science and Technology Middlesex University The Burroughs, Hendon, London NW4 4BT, England, UK

Preface

Environmental issues have been always at the forefront of sustainable development and have become a serious matter of concern in the twenty-first century. Environmental sustainability with rapid industrialization is one of the major challenges of the current scenario worldwide. Industries are the key drivers in the world economy, but these are also the major polluters due to the discharge of partially treated/untreated potentially toxic and hazardous wastes containing organic and inorganic pollutants, which cause environmental (soil and water) pollution and severe toxicity in living beings. Among the different sources of environmental pollution, industrial wastes are considered as the major source of environmental pollution because industries use cheap and poorly or nonbiodegradable chemicals to obtain good quality of products within a short period of time and in an economic way; however, their toxicity is usually ignored. Ensuring the safety of chemicals used in many industrial processes is a major challenge for environmental safety. The governments around the globe are also strictly advocating for the mitigation of environmental pollution due to industrial wastes to promote the sustainable development of our society with low environmental impact. Being a low-cost and eco-friendly clean technology, bioremediation can be an eco-sustainable alternative to conventional technologies for the treatment and management of industrial wastes to protect public health and the environment. Bioremediation is a waste management approach that utilizes microorganisms, plants, or their enzymes to degrade/detoxify organic and inorganic pollutants such as phenols, chlorophenols, petroleum hydrocarbons, polychlorinated biphenyls, organic solvents, azo dyes, pesticides, recalcitrant compounds, toxic metals, etc. from contaminated soils and wastewaters. There has been an increasing concern regarding the release of various hazardous chemicals along with industrial wastes, which are considered highly toxic for the environment and living beings. Some of these chemicals are classified as priority pollutants by the United States Environmental Protection Agency (USEPA) and other environmental pollution control agencies. The biological removal of a wide range of pollutants from contaminated sites requires our increasing understanding of different degradation pathways and ix

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regulatory networks to carbon flux for their degradation and detoxification, which is utmost important for environmental safety. Therefore, this book provides a comprehensive knowledge of the fundamental, practical, and purposeful utilization of bioremediation technologies for the treatment and management of industrial wastes. The book describes the microbiological, biochemical, and molecular aspects of biodegradation and bioremediation, including the use of “omics” technologies for the development of efficient bioremediation technologies for industrial wastes/pollutants to combat the forthcoming challenges. This book Bioremediation of Industrial Waste for Environmental Safety: Biological Agents and Methods for Industrial Waste Management (Volume II) describes the toxicity of various organic and inorganic pollutants in industrial wastes, their environmental impact, and bioremediation approaches for waste treatment and management. For this book, many relevant topics have been contributed by experts from different universities, research laboratories, and institutes from around the globe in the area of biodegradation and bioremediation. In this book, extensive focus has been relied on the recent advances in bioremediation and phytoremediation technologies, including the use of an array of microbes for environmental remediation, aquatic macrophytes for phytoremediation of toxic metals from contaminated industrial wastewaters, constructed wetlands for degradation and detoxification of industrial wastewaters, genetically modified organisms (GMOs) for degradation and detoxification of environmental pollutants, bioremediation of toxic metals and radionuclides in contaminated environments, algae as a phycoremediation tool for the removal of heavy metals from industrial wastewaters, electro-bioremediation and nano-bioremediation for the treatment and management of contaminated soil, application of microbial fuel cell (MFC) for the treatment and remediation of highly polluted wastewaters and power generation, application of plant-microbe interactions in the remediation of environmental pollutants and agro-ecosystem development, the role of endophytes in heavy metal (HM) phytoremediation, microalgae for the treatment of industrial wastewaters with value-added product potential, phytotechnologies for wastewater treatment and management, the role of cyanobacteria in the remediation of wastewaters, bioremediation of pulp paper mill wastewater and heavy metals like cadmium, etc. Researchers working in the field of bioremediation, phytoremediation, waste management, and related fields will find a compilation on the progress made in bioremediation technologies for industrial waste treatment and management for environmental sustainability. To get richer in the knowledge on the topic, readers may visit our first volume of this book series, i.e., Bioremediation of Industrial Waste for Environmental Safety: Industrial Waste and Its Management (Volume I). At the end, we hope that the book will be of great value to researchers, environmental chemists and scientists, microbiologists and biotechnologists, eco-toxicologists, waste treatment engineers and managers, environmental science managers, administrators and policy makers, industry persons, and students at

Preface

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bachelor’s, master’s, and doctoral levels in the relevant field. Thus, in this book, readers will find updated information, as well as the future direction for research in the field of bioremediation. Lucknow, Uttar Pradesh, India Lucknow, Uttar Pradesh, India May 2018

Ram Naresh Bharagava Gaurav Saxena

Acknowledgments

The edited book Bioremediation of Industrial Waste for Environmental Safety: Biological Agents and Methods for Industrial Waste Management (Volume II) is the outcome of long dedicated efforts of many individuals who directly or indirectly supported us during the compilation and upbringing of this valuable edition, many of whom deserve special mention. Editors are thankful to all the national and international contributing authors for their valuable submissions and cooperation and for providing most up-to-date information on the diverse aspects of the subject regardless of their busy schedules; Prof. Diane Purchase, Middlesex University, London, England (United Kingdom), for writing an opinion foreword for the book; Dr. G. D. Saratale, Dongguk University-Seoul, Seoul (Republic of Korea), and Dr. Sikandar I. Mulla, Chinese Academy of Sciences (CAS), Xiamen (People’s Republic of China), for the meaningful collaboration, cooperation, and support; Dr. Jay Shankar Singh, Department of Environmental Microbiology (DEM), Babasaheb Bhimrao Ambedkar (Central) University (BBAU), Lucknow (India), for the better advice and helpful discussion on the subject, and Mr. Surya Pratap Goutam and Mr. Rajkamal Shastri, Doctoral Fellow, Department of Applied Physics; Roop Kishor, Doctoral Fellow, DEM, BBAU, and Mr. Akash Mishra, Doctoral Fellow, Defence Research and Development Organization (DRDO)-Defence Institute of Bioenergy Research (DIBER), Haldwani (India), for helping us in various ways during the book project. We are extremely thankful to our publishing editors, Ms. Aakanksha Tyagi and Dr. Mamta Kapila, Springer Nature (India), for the encouragement, support, and valuable advice and skillful organization and management of the entire book project; Ms. Raman Shukla, for the skillful management of book production, Mr. N. S. Pandian and Mr. Daniel Ignatius Jagadisan, for moving the book through the production process in an efficient and professional manner and R. Rathika for generating and managing the book proof. We are also heartily thankful to the Almighty God for helping us through the entire journey and making the experience enjoyable. Further, we hope that the book volume will be of great value to researchers in the area of bioremediation of xiii

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industrial wastes and will go some way to make our planet safe and greener. At the end, we seek to learn more on the subject through the valuable comments, reviews, and suggestions from our readers, which can be directly sent to our e-mails: [email protected] (Ram Naresh Bharagava) and [email protected] (Gaurav Saxena). May 2018

Ram Naresh Bharagava Gaurav Saxena

Contents

1

2

3

4

Genetically Modified Organisms (GMOs) and Their Potential in Environmental Management: Constraints, Prospects, and Challenges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Gaurav Saxena, Roop Kishor, Ganesh Dattatraya Saratale, and Ram Naresh Bharagava Advances in Bioremediation of Toxic Heavy Metals and Radionuclides in Contaminated Soil and Aquatic Systems . . . . . . . Evans M. Nkhalambayausi-Chirwa, Pulane Elsie Molokwane, Tshilidzi Bridget Lutsinge, Tony Ebuka Igboamalu, and Zainab S. Birungi Phycoremediation: Algae as Eco-friendly Tools for the Removal of Heavy Metals from Wastewaters . . . . . . . . . . . . . . Shamshad Ahmad, Arya Pandey, Vinayak Vandan Pathak, Vineet Veer Tyagi, and Richa Kothari Recent Advances in Phytoremediation of Toxic Metals from Contaminated Sites: A Road Map to a Safer Environment . . . . . . . Mukesh Kumar Awasthi, Di Guo, Sanjeev Kumar Awasthi, Quan Wang, Hongyu Chen, Tao Liu, Yumin Duan, Parimala Gnana Soundari, and Zengqiang Zhang

1

21

53

77

5

Emerging and Ecofriendly Technologies for the Removal of Organic and Inorganic Pollutants from Industrial Wastewaters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113 Gaurav Saxena, Surya Pratap Goutam, Akash Mishra, Sikandar I. Mulla, and Ram Naresh Bharagava

6

Constructed Wetlands: A Clean-Green Technology for Degradation and Detoxification of Industrial Wastewaters . . . . 127 Sardar Khan, Javed Nawab, and Muhammad Waqas xv

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Contents

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Nano-bioremediation: An Innovative Remediation Technology for Treatment and Management of Contaminated Sites . . . . . . . . . 165 Ritu Singh, Monalisha Behera, and Sanjeev Kumar

8

Electro-bioremediation: An Advanced Remediation Technology for the Treatment and Management of Contaminated Soil . . . . . . . 183 Sivasankar Annamalai and Maruthamuthu Sundaram

9

Microbial Fuel Cell (MFC): An Innovative Technology for Wastewater Treatment and Power Generation . . . . . . . . . . . . . 215 Mostafa Rahimnejad, Maryam Asghary, and Marjan Fallah

10

Functional Diversity of Plant Endophytes and Their Role in Assisted Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 237 Angélica Leonor Guerrero-Zúñiga, Eugenia López-López, Aída Verónica Rodríguez-Tovar, and Angélica Rodríguez-Dorantes

11

Toxic Metals in Industrial Wastewaters and Phytoremediation Using Aquatic Macrophytes for Environmental Pollution Control: An Eco-Remedial Approach . . . . . . . . . . . . . . . . . . . . . . . 257 Mansi Rastogi and Meenakshi Nandal

12

Microalgae: An Eco-friendly Tool for the Treatment of Wastewaters for Environmental Safety . . . . . . . . . . . . . . . . . . . . 283 Jae-Hoon Hwang, Anwar Sadmani, Seung-Jin Lee, Keug-Tae Kim, and Woo Hyoung Lee

13

Phycoremediation: An Integrated and Eco-friendly Approach for Wastewater Treatment and Value-Added Product Potential . . . 305 J. Umamaheswari, D. Saranya, S. Abinandan, Mallavarapu Megharaj, Suresh R. Subashchandrabose, and S. Shanthakumar

14

Pulp and Paper Mill Wastewater: Ecotoxicological Effects and Bioremediation Approaches for Environmental Safety . . . . . . . 333 Izharul Haq and Abhay Raj

15

Cadmium as an Environmental Pollutant: Ecotoxicological Effects, Health Hazards, and Bioremediation Approaches for Its Detoxification from Contaminated Sites . . . . . . . . . . . . . . . . 357 Sushila Saini and Geeta Dhania

16

Cyanobacteria: The Eco-Friendly Tool for the Treatment of Industrial Wastewaters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 389 Sharma Mona, Virendra Kumar, Bansal Deepak, and Anubha Kaushik

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17

Plant-Microbe Interactions for Bioremediation and Phytoremediation of Environmental Pollutants and Agro-ecosystem Development . . . . . . . . . . . . . . . . . . . . . . . . . . 415 Akash Mishra, Shraddha Priyadarshini Mishra, Anfal Arshi, Ankur Agarwal, and Sanjai Kumar Dwivedi

18

Molecular Technologies for Assessment of Bioremediation and Characterization of Microbial Communities at Pollutant-Contaminated Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . 437 Sudhir Kumar Shekhar, Jai Godheja, and Dinesh Raj Modi

19

Biochar: A Sustainable Tool in Soil Pollutant Bioremediation . . . . . 475 Chhatarpal Singh, Shashank Tiwari, and Jay Shankar Singh

20

Bioremediation of Melanoidins Containing Distillery Waste for Environmental Safety . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 495 Vineet Kumar and Ram Chandra

21

Progresses in Bioremediation Technologies for Industrial Waste Treatment and Management: Challenges and Future Prospects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 531 Ram Naresh Bharagava and Gaurav Saxena

Editors and Contributors

About the Editors Ram Naresh Bharagava Dr. Ram Naresh Bharagava was born in 1977 and completed school education from Government Schools at Lakhimpur Kheri, Uttar Pradesh (UP), India. He received his B.Sc. degree (1998) in Zoology, Botany, and Chemistry from the University of Lucknow, Lucknow, UP, India, and M.Sc. degree (2004) in Molecular Biology and Biotechnology from Govind Ballabh Pant University of Agriculture & Technology (GBPUAT), Pantnagar, Uttarakhand (UK), India. He received his Ph.D. degree (2010) in the area of Microbiology jointly from the Environmental Microbiology Division, Council of Scientific and Industrial Research (CSIR)-Indian Institute of Toxicology Research (IITR), Lucknow, and Pt. Ravishankar Shukla University, Raipur, Chhattisgarh, India. He was a Junior Research Fellow (JRF) during his Ph.D., qualified twice (2002 & 2003) in the CSIR-National Eligibility Test (NET) and once in the Graduate Aptitude Test in Engineering (GATE) in 2003. His major research work during his Ph.D. degree focused on the bacterial degradation of recalcitrant melanoidin from distillery wastewater. He has authored one book entitled Bacterial Metabolism of Melanoidins from Distillery Effluent and edited five books entitled Bioremediation of Industrial Pollutants, Environmental Pollutants and Their Bioremediation Approaches, Emerging and Ecofriendly Approaches for Waste Management, and Recent Trends in xix

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Environmental Management and Bioremediation of Industrial Waste for Environmental Safety: Industrial Waste and Its Management (Volume I). He has published more than 117 research items including research/review papers/two book reviews in national and international journals of high impact factor published by Springer Nature, Elsevier, and Taylor & Francis Group. He has also written more than 40 book chapters for national and international edited books. He has published many scientific articles in national and international science magazines. He has presented many papers relevant to his research areas in national and international conferences. He has served as a potential reviewer for various national and international journals in his respective areas of research. He has also been awarded a postdoctoral appointment at CSIR-IITR, Lucknow, but left the position after a while and subsequently joined (2011) Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, UP, India, where he is currently Assistant Professor of Environmental Microbiology, and actively engaged in teaching at postgraduate and doctoral levels and research on various Government of India (GOI) sponsored research projects in the area of environmental toxicology and bioremediation at the Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Environmental Microbiology (DEM). His research has been financially supported by the University Grants Commission (UGC) and Department of Science and Technology (DST) India. He has been the advisor of 60 postgraduate and five doctoral students and currently the mentor of one project fellow and one doctoral student. His major thrust areas of research are the biodegradation and bioremediation of environmental pollutants in industrial wastewaters, metagenomics, and wastewater microbiology. He is life member of the Academy of Environmental Biology (AEB), Association of Microbiologists of India (AMI), and Biotech Research Society (BRSI), Indian Science Congress Association (ISCA), India. He can be reached at [email protected].

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Gaurav Saxena was born in 1989 and completed school education from Government Schools at Shahjahanpur, Uttar Pradesh (UP), India. He received his B.Sc. degree (2010) in Industrial Microbiology, Zoology, Botany, and Chemistry from Hemwati Nandan Bahuguna Garhwal (Central) University (HNBGU), Srinagar (Garhwal), Uttarakhand (UK), India, and then moved to Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, UP, India, where he received his M.Sc. degree (2013) in Environmental Microbiology. Because of his interest in environmental safety and sustainability, Mr. Saxena joined Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, UP, India, to further pursue doctoral research in environmental microbiology at the Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Microbiology (DM). Currently, he is a senior doctoral student and investigating the role of microbes in biodegradation/bioremediation of industrial effluents containing highly toxic organic and inorganic pollutants and their ecotoxicological assessment and engaged in developing eco-friendly treatment solutions for hazardous tannery effluent. His research work was financially supported by the University Grants Commission (UGC) and the Department of Science and Technology (DST), Government of India (GOI), New Delhi, India. He has qualified (2015) in the National Eligibility Test (NET) and also received the Junior Research Fellowship (JRF) of DST, GOI, New Delhi, India. He has been awarded honorary “Young Environmentalist Award-2018” by Agro-Environmental Development Society (AEDS), India, in recognition of his scientific work. He is the editor of two books, entitled Bioremediation of Industrial Pollutants and Bioremediation of Industrial Waste for Environmental Safety: Industrial Waste and Its Management (Volume I). He is also on the editorial board of Frontiers in Microbiology journal and serving as a review editor. He has published many scientific papers and one book review relevant to his field in high impact national and international journals published by Springer Nature, Elsevier, and International Water Association (IWA). He has also written several chapters on the relevant topics of his field for national and international edited books published by Springer Nature,

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Taylor & Francis, and Elsevier. He has published many scientific articles in national and international science magazines and also presented his research works in national and international conferences. He is also serving as a potential reviewer for various scientific national and international journals in his research areas. He has received several short-term trainings on various subjects like bioremediation, various analytical techniques for pollutant detection and characterization in environmental samples, molecular techniques, biological treatment of waste, nanotechnology, and research methodology from national laboratories, institutes, and universities in India. He is always interested in collaborating with national and international researchers relevant to his research area, and his major thrust areas of research are environmental microbiology, biotechnology, and toxicology, bioremediation/phytoremediation of environmental pollutants/industrial effluents, metagenomics, wastewater microbiology, and wastewater treatment. He is a life member of the Association of Microbiologists of India (AMI), Indian Science Congress Association (ISCA), and Agro-Environmental Development Society (AEDS), India. He is a nature lover and has keen interest in environmental protection. He can be reached at [email protected]. For more updates on his research, please visit ResearchGate profile (www.researchgate.net/profile/Gaurav_Saxena2).

Contributors S. Abinandan Global Centre for Environmental Remediation (GCER), Faculty of Science and Information Technology, University of Newcastle, Newcastle, New South Wales (NSW), Australia Ankur Agarwal Defence Institute of Bio-Energy Research (DIBER), Defence Research and Development Organization (DRDO), Ministry of Defence, Government of India, Haldwani, Uttarakhand, India Shamshad Ahmad Bioenergy and Wastewater Treatment Laboratory, Department of Environmental Sciences, Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India

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Sivasankar Annamalai Microbial Corrosion and Bio-Environmental Engineering Group, Corrosion and Materials Protection Division, CSIR – Central Electrochemical Research Institute, Karaikudi, Tamil Nadu, India Anfal Arshi Defence Institute of Bio-Energy Research (DIBER), Defence Research and Development Organization (DRDO), Ministry of Defence, Government of India, Haldwani, Uttarakhand, India Maryam Asghary Department of Analytical Chemistry, Faculty of Chemistry, University of Mazandaran, Babolsar, Mazandaran Province, Iran Mukesh Kumar Awasthi College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Sanjeev Kumar Awasthi College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Monalisha Behera Department of Environmental Science, Central University of Rajasthan, Ajmer, Rajasthan, India Zainab S. Birungi Water Utilization and Environmental Engineering Division, Department of Chemical Engineering, University of Pretoria, Pretoria, Gauteng, South Africa Ram Chandra Department of Environmental Microbiology (DEM), School for Environmental Sciences (SES), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Hongyu Chen College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Bansal Deepak JBM Group, Gurugram, Haryana, India Geeta Dhania Department of Environmental Sciences, Maharshi Dayanand University, Rohtak, Haryana, India Yumin Duan College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Sanjai Kumar Dwivedi Defence Institute of Bio-Energy Research (DIBER), Defence Research and Development Organization (DRDO), Ministry of Defence, Government of India, Haldwani, Uttarakhand, India Marjan Fallah Biofuel and Renewable Energy Research Center, Faculty of Chemical Engineering, Babol Noshirvani University of Technology, Babol, Mazandaran Province, Iran Jai Godheja School of Life and Allied Sciences, ITM University, Naya Raipur, Chhattisgarh, India

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Editors and Contributors

Surya Pratap Goutam Advanced Materials Research Laboratory, Department of Applied Physics (DAP), School for Physical Sciences (SPS), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Angélica Leonor Guerrero-Zúñiga Gerencia de Transformación de Biomasa, Instituto Mexicano del Petróleo, Ciudad de México, Mexico Di Guo College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Izharul Haq Environmental Microbiology Laboratory, Environmental Toxicology Group, CSIR-Indian Institute of Toxicology Research (CSIR-IITR), Lucknow, Uttar Pradesh, India Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, Assam, India Jae-Hoon Hwang Department of Civil, Environmental and Construction Engineering, University of Central Florida, Orlando, FL, USA Tony Ebuka Igboamalu Water Utilization and Environmental Engineering Division, Department of Chemical Engineering, University of Pretoria, Pretoria, Gauteng, South Africa Anubha Kaushik University School of Environment Management (USEM), Guru Gobind Singh Indraprastha University, New Delhi, India Sardar Khan Department of Environmental Science, University of Peshawar, Peshawar, Pakistan Keug-Tae Kim Department of Environmental & Energy Engineering, University of Suwon, Hwaseong-si, Gyeonggi-do, Republic of Korea (South Korea) Roop Kishor Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Microbiology (DM), Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, Uttar Pradesh, India Richa Kothari Department of Environmental Sciences, Central University of Jammu, RahyaSuchani (Bagla), Samba, Jammu, Jammu and Kashmir, India Sanjeev Kumar Centre for Environmental Sciences, Central University of Jharkhand, Ranchi, Jharkhand, India Vineet Kumar Department of Environmental Microbiology (DEM), School for Environmental Sciences (SES), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Virendra Kumar Department of Environmental Sciences, Central University of Haryana, Pali, Haryana, India

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Seung-Jin Lee Department of Computer Science, Engineering, and Physics & Department of Geography, Planning, and Environment, University of MichiganFlint, Flint, MI, USA Woo Hyoung Lee Department of Civil, Environmental and Construction Engineering, University of Central Florida, Orlando, FL, USA Ronghua Li College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Tao Liu College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Eugenia López-López Laboratorio de Evaluación de la Salud de los Ecosistemas AcuÃticos, Departamento de Zoología, Escuela Nacional de Ciencias Biológicas, Prol. Carpio y Plan de Ayala, S/N, Unidad Profesional Lázaro Cárdenas, Instituto Politécnico Nacional, Ciudad de México, México Tshilidzi Bridget Lutsinge Water Utilization and Environmental Engineering Division, Department of Chemical Engineering, University of Pretoria, Pretoria, Gauteng, South Africa Mallavarapu Megharaj Global Centre for Environmental Remediation (GCER), Faculty of Science and Information Technology, University of Newcastle, Newcastle, New South Wales (NSW), Australia Akash Mishra Defence Institute of Bio-Energy Research (DIBER), Defence Research and Development Organization (DRDO), Ministry of Defence, Government of India, Haldwani, Uttarakhand, India Shraddha Priyadarshini Mishra Defence Institute of Bio-Energy Research (DIBER), Defence Research and Development Organization (DRDO), Ministry of Defence, Government of India, Haldwani, Uttarakhand, India Dinesh Raj Modi Department of Biotechnology (DBT), School for Bioscience and Biotechnology (SBBT), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Pulane Elsie Molokwane Water Utilization and Environmental Engineering Division, Department of Chemical Engineering, University of Pretoria, Pretoria, Gauteng, South Africa Oloenviron Consultancy, Johannesburg, South Africa Sharma Mona Department of Environmental Sciences, Central University of Haryana, Pali, Haryana, India Department of Environmental Science and Engineering, Guru Jambheshwar University of Science & Technology, Hisar, Haryana, India

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Sikandar I. Mulla CAS Key Laboratory of Urban Pollutant Conversion, Institute of Urban Environment, Chinese Academy of Sciences, Xiamen, People’s Republic of China Meenakshi Nandal Department of Environmental Sciences, Maharshi Dayanand University, Rohtak, Haryana, India Javed Nawab Department of Environmental Sciences, Abdul Wali Khan University, Mardan, Pakistan Department of Environmental and Conservation Sciences, University of Swat, Mingora, Pakistan Evans M. Nkhalambayausi-Chirwa Water Utilization and Environmental Engineering Division, Department of Chemical Engineering, University of Pretoria, Pretoria, Gauteng, South Africa Arya Pandey Bioenergy and Wastewater Treatment Laboratory, Department of Environmental Sciences, Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Vinayak Vandan Pathak Department of Chemistry, Manav Rachna University, Faridabad, Haryana, India Loganathan Praburaman College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Mostafa Rahimnejad Biofuel and Renewable Energy Research Center, Faculty of Chemical Engineering, Babol Noshirvani University of Technology, Babol, Mazandaran Province, Iran Abhay Raj Environmental Microbiology Laboratory, Environmental Toxicology Group, CSIR-Indian Institute of Toxicology Research (CSIR-IITR), Lucknow, Uttar Pradesh, India Mansi Rastogi Department of Environmental Sciences, Maharshi Dayanand University, Rohtak, Haryana, India Xiuna Ren College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Angélica Rodríguez-Dorantes Laboratorio de Fisiología Vegetal, Departamento de Botánica, Escuela Nacional de Ciencias Biológicas, Prol. Carpio y Plan de Ayala, S/N, Unidad Profesional Lázaro Cardenas, Instituto Politécnico Nacional, Ciudad de México, México Aída Verónica Rodríguez-Tovar Laboratorio de Micología, Departamento de Microbiología, Escuela Nacional de Ciencias Biológicas, Prol. Carpio y Plan de Ayala, S/N, Unidad Profesional Lázaro Cardenas, Instituto Politécnico Nacional, Ciudad de México, México

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Anwar Sadmani Department of Civil, Environmental and Construction Engineering, University of Central Florida, Orlando, FL, USA Sushila Saini Department of Botany, Janta Vidya Mandir Ganpat Rai Rasiwasia (JVMGRR) College, Charkhi Dadri, Haryana, India D. Saranya Department of Environmental and Water Resources Engineering, School of Civil and Chemical Engineering, VIT University, Vellore, Tamil Nadu, India Ganesh Dattatraya Saratale Department of Food Science and Biotechnology, Dongguk University-Seoul, Goyang-si, Gyeonggi-do, Republic of Korea S. Shanthakumar Department of Environmental and Water Resources Engineering, School of Civil and Chemical Engineering, VIT University, Vellore, Tamil Nadu, India Sudhir Kumar Shekhar Department of Biotechnology (DBT), School for Bioscience and Biotechnology (SBBT), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Chhatarpal Singh Department of Environmental Microbiology (DEM), School for Environmental Sciences (SES), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Jay Shankar Singh Department of Environmental Microbiology (DEM), School for Environmental Sciences (SES), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India Ritu Singh Department of Environmental Science, Central University of Rajasthan, Ajmer, Rajasthan, India Parimala Gnana Soundari College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Suresh R. Subashchandrabose Global Centre for Environmental Remediation (GCER), Faculty of Science and Information Technology, University of Newcastle, Newcastle, New South Wales (NSW), Australia Cooperative Research Centre for Contamination Assessment and Remediation of the Environment (CRC CARE), C/- Newcastle University, Newcastle, New South Wales (NSW), Australia Maruthamuthu Sundaram Microbial Corrosion and Bio-Environmental Engineering group, Corrosion and Materials Protection Division, CSIR – Central Electrochemical Research Institute, Karaikudi, Tamil Nadu, India Shashank Tiwari Department of Environmental Microbiology (DEM), School for Environmental Sciences (SES), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India

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Editors and Contributors

Vineet Veer Tyagi Department of Energy Management, Shri Mata Vaishno Devi University, Katra, Jammu and Kashmir, India J. Umamaheswari Department of Environmental and Water Resources Engineering, School of Civil and Chemical Engineering, VIT University, Vellore, Tamil Nadu, India Meijing Wang College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Quan Wang College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Muhammad Waqas Department of Environmental and Conservation Sciences, University of Swat, Mingora, Pakistan Zeng Qiang Zhang College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Zengqiang Zhang College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China Junchao Zhao College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China

Chapter 1

Genetically Modified Organisms (GMOs) and Their Potential in Environmental Management: Constraints, Prospects, and Challenges Gaurav Saxena, Roop Kishor, Ganesh Dattatraya Saratale, and Ram Naresh Bharagava

Abstract Increasing environmental contamination with highly toxic chemicals is warning us to find sustainable technologies to protect the environment and human health, which is a key challenge of the current scenario. A variety of physicochemical technologies are currently being applied presently to decontaminate the environment to safeguard the environment and human health. However, these technologies are costly and chemical-consuming, thus causing secondary pollution and, hence, are not environmental-friendly. As an alternative approach, bioremediation technologies using microbes and plants and their enzymes are currently viewed as eco-friendly and most sustainable technologies due to their self-sustainable and low-cost nature. But sometimes bioremediation technologies are get limited by low degradability/accumulability of microbes and plants, respectively. To overcome these limitations, genetic engineering approaches are highly decisive to design the transgenic microbes and plants for the enhanced biodegradation and biodetoxification of environmental pollutants for sustainable development. Genetically modified organisms (GMOs) offer great potential for environmental remediation, and hence, in this chapter, we focused on the applications of GMOs in the environmental management with risks involved, constraints, and challenges faced by researchers in the release of GMOs for field applications. Keywords Environmental pollutants · Genetically modified organisms · Environmental remediation · Transgene · Genetic engineering

G. Saxena · R. Kishor · R. N. Bharagava (*) Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Microbiology (DM), Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, Uttar Pradesh, India G. D. Saratale Department of Food Science and Biotechnology, Dongguk University-Seoul, Goyang-si, Gyeonggi-do, Republic of Korea © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_1

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1 Introduction In the last few decades, due to industrialization, increase in population, and daily life requirements, harmful chemicals have been released into the earth’s air, soil, and water (Goutam et al. 2018; Gautam et al. 2017; Bharagava et al. 2017a, b; Saxena et al. 2016; Olugbenga 2017). Excessive mining, agriculture waste, and burning of fossil fuels consequently release enormous amounts of toxic heavy metals like Hg, Pb, U, Cd, Zn, Cr, Ni, Co, and Cu and metalloids (As) into the environment which create mutagenic and carcinogenic effect (Wernick and Themelis 1998; Wijnhoven et al. 2007). Several chemical industries use and produce wide varieties of hazardous compounds like benzene, toluene, polychlorinated biphenyls (PCBs), polyaromatic hydrocarbons (PAHs), dioxins, nitro-aromatics, dyes, polymers, pesticides, explosives, chlorinated organic, and pharmaceuticals (Meagher 2000; Pilon-Smits 2005). Moreover, many of these substances are non-biodegradable and persistent in nature that stay long in our natural environment. Many of these substances are toxic and cause a harmful effect on human health and damage the ecological balance. However, there is an urgent need to remove these compounds for environmental and public health safety. The remediation and restoration of sites contaminated with highly toxic and hazardous pollutants requires eco-friendly and effective approach for environmental sustainability and to safeguard the public health. Microbial bioremediation is a waste management technology which uses microorganisms like bacteria, algae, and fungi to degrade and transform hazardous compounds of soil and water, while phytoremediation is cost-effective and environmental-friendly technology that has a potential application to efficiently degrade and transform organic and inorganic pollutants (Kishor et al. 2018; Saxena and Bharagava 2016; Bharagava et al. 2017c, 2019; Meagher 2000). Eventually, naturally occurring microorganisms are incapable of degrading all toxic compounds, especially xenobiotic. To overcome this, serious efforts have been done to create genetically engineered microorganisms (GEMs) to enhance bioremediation approaches besides degrading xenobiotic (Sayler and Ripp 2000). Thus, biotechnology is a most important technique that has been applied in different areas especially in remediation to neutralize various unfit complex environmental pollutants into nontoxic or simple form and to completely remediate organic wastes (Iwamoto and Nasu 2001). Recombinant DNA technology has been studied intensively to improve the biodegradation of hazardous pollutants in lab conditions (Dua et al. 2002). In the late 1970s and early 1980s, the cloning and characterization of bacterial genes that code for catabolic enzymes for the biodegradation of recalcitrant pollutants has started. The organism whose genetic material, i.e., DNA, has been modified/altered in such a way so as to get the required traits is often called as genetically modified organism (Shukla et al. 2010; Liu et al. 2011). This technology is often called “gene technology,” or “recombinant DNA technology” (RDT), or “genetic engineering,” and the resulting organism is said to be “genetically modified,” “genetically engineered,” or “transgenic.”

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In addition, the leakage and industrial discharge of petrol and their associated chemicals like polycyclic aromatic hydrocarbons (PAH) pose a highly negative impact on aquatic and terrestrial ecosystems. Genetically modified organisms (GMOs) have a capability to clean up and remove industrial waste and pollutants from the environment as well as reduce toxicity of elements (Liu et al. 2011). Genetic engineering is currently popular among researchers worldwide to develop new microbes with required traits as compared to its wild type for the degradation and detoxification of a wide range of xenobiotic compounds (Kumar et al. 2013). In 1970, the first GMOs called “superbug” were developed by genetic engineering through plasmid transfer that have ability to degrade a variety of petroleum chemicals such as xylene, camphor, hexane, naphthalene, and toluene. GMOs are capable for enhanced degradation and removal of a wide range of xenobiotic and also have potential application for bioremediation of environmental pollutants (Kulshreshtha 2013). Designing of GMOs primarily depend on the knowledge of genetic basis of interaction between microbes and xenobiotic compounds, structure of operon, molecular biology, biochemistry, and ecology (ref). Thus, GMOs can be potential molecular tools to degrade and detoxify the environmental pollutants in contaminated matrix to safeguard the environment and public health. Therefore, this chapter has mainly focused on the role of GMOs in the bioremediation of organic and inorganic pollutants, constraints in utilizing them in bioremediation, and limitations in field applications.

2 Genetically Modified Organisms Designing of suitable genetically modified organisms (GMOs) for enhanced bioremediation of environmental pollutants from contaminated matrix requires creation of new routes for metabolism, intensifying a range of existing degradation pathways, avoiding substrate misrouting into unproductive routes or to toxic metabolite generation, improving the substrate flux through degradation pathways to avoid the accumulation of toxic intermediates, enhanced stability of catabolic potential, enhanced bioavailability of hydrophobic pollutants, and enhanced catabolic potential of microbes (Timmis and Pieper 1999; Pieper and Reineke 2000; Furukawa 2003). Although an organism produced from genetic engineering techniques allows the transfer of specific functional genes into a particular organism genome (Tozzini 2000). A US definition of GMO, “genetically modified organisms,” refers to microorganism, plants, and animals containing distinctive genes transferred from other species to produce unique characteristics to completely clean up and mineralize hazardous waste material. Many bacterial strains such as Bacillus idriensis, Ralstonia eutropha, Sphingomonas desiccabilis, Pseudomonas putida, Escherichia coli, Mycobacterium marinum, etc. have been used to design genetically engineered microbes with insertion of a functional gene into other species which capable for the bioremediation of heavy metals and non-biodegradable compounds of contaminated

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environment (Valls et al. 2000; Ackerley et al. 2004; Kube et al. 2005; Parnell et al. 2006; Schue et al. 2009; Liu et al. 2011). Moreover, the genetic engineering of plants also performed to enhance the accumulation and tolerance capacity as well as detoxification potential for heavy metal pollutants and to increase the biomass and growth of plants in metal contaminated sites (Hassani 2014). Metallothioneins (MTs) are the unique cysteine-rich peptides that are relevant to higher metal-binding capacity in hyperaccumulating plants and have been cloned to develop the genetically engineered plants for phytoremediation of organic and inorganic pollutants. Tobacco plant was the first genetically engineered plant for the phytoremediation of explosives and halogenated organic pollutants (Doty et al. 2000). Many reports have been published on the genetic engineering of plants and their role in the phytoremediation of contaminated soil and water environment (Cherian and Oliveira 2005; Pilon-Smits 2005; Eapen et al. 2007; Doty 2008; Macek et al. 2008; James and Strand 2009; Kawahigashi 2009; Van 2009). Recently, James and Strand (2009) reported the dehalogenation of tetrachloroethylene (PCE) by hybrid poplar trees under controlled field conditions. Genetically modified organisms can be also used as biosensors for related mixures of agrochemicals, petroleum products, metals, and toxins that are found in the environment, but cannot be directly in soil or water (Ozcan et al. 2011).

3 Environmental Bioremediation Technologies Environmental bioremediation technologies broadly can be classified into two major categories: bioremediation and phytoremediation.

3.1

Bioremediation

Bioremediation is the eco-friendly technique wherein biological agents (microbes and plants or their enzymes) are used to degrade and detoxify the organic and inorganic pollutants to safeguard the environment and public health in low-cost and efficient manner (Azubuike et al. 2016; Bharagava et al. 2018; Kishor et al. 2018). A range of bioremediation techniques have been developed by researchers to date; but due to diverse characteristics of pollutants and merits and demerits, no single bioremediation technique can provide full-scale solution to contaminated environment (Verma and Jaiswal 2016). Microbes that are involved in the degradation and detoxification of organic and inorganic pollutants are Mycobacterium, Acinetobacter, Flavobacterium, Actinobacteria, Alcaligenes, Beijerinckia, Arthrobacter, Methylosinus, Bacillus, Micrococcus, Serratia, Nitrosomonas, Rhizoctonia, Pseudomonas, Nocardia, Phanerochaete, Penicillium, Xanthobacter, and Trametes. Bioremediation involves three main processes: biotransformation (conversion of organic and inorganic pollutants into less or nonhazardous molecules),

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biodegradation (breakdown of complex organic pollutants into simple and smaller unit molecules), and mineralization (complete biodegradation of organic matter into inorganic constituents such as CO2 or H2O) (Saxena and Bharagava 2017; Saxena and Bharagava 2015; Pilon-Smits 2005). On the basis of application potential, bioremediation can be applied as ex situ and in situ. In situ bioremediation technologies involve treatment of pollutants at the site of pollution, do not require any excavation means, do not pose any disturbance to soil environment, and require continuous oxygen supply for proper aeration to support the microbial growth for degradation of contaminants (Vidali 2001). In situ bioremediation technologies are cost-effective as these uses microbes for pollutant removal from contaminated matrix and for the degradation and detoxification of polyaromatic hydrocarbons, azo dyes, chlorinated solvents, and heavy metals (Kumar et al. 2011; Folch et al. 2013; Kim et al. 2014; Frascari et al. 2015; Roy et al. 2015). In situ bioremediation technologies are biosparging, bioventing, and phytoremediation. Ex situ bioremediation technologies involve the treatment of pollutants at any place other than the site of pollution and require excavation of contaminated soil or pumping of groundwater to enhance the microbial degradation process. These remediation approaches are costly, and their applicability depends on the pollutants type, pollution strength and depth, and geographic conditions of contaminated sites (Philp and Atlas 2005). These approaches are classified into two methods: solid phase system (including land treatment and soil piles) and slurry phase systems (including solid liquid suspensions in bioreactors) (Kumar et al. 2013).

3.2

Phytoremediation

Phytoremediation is an eco-friendly phytotechnology that involves the use of plants/ trees for the treatment and restoration of contaminated sites/wastewaters/groundwater (Saxena et al. 2019; Chandra et al. 2015). By using green plants, the pollutants such as metals, pesticides, herbicides, explosives, oil, solvents, and their derivatives can be removed and cleaned up from polluted and contaminated soil, streams, and groundwater (Meagher 2000; Pilon-Smits 2005). Phytoremediation technologies may be inexpensive and harmless process than traditional ones and offer easy plant control and re-use of valuable metals. Exudates released by roots in the rhizosphere of plants also support the growth of soil beneficial microbes that participate in the degradation and detoxification of pollutants (rhizoremediation), and chelating agents help to convert non-available elements into bioavailable forms for plant uptake for growth (Suresh and Ravishankar 2004; Abhilash et al. 2009). The genetically engineered plants have been developed through transgenic engineering to degrade and detoxify the organic and inorganic pollutants (Zhu et al. 1999; Abhilash et al. 2009). The increased accumulation of pollutants (in case of heavy metals) facilitates their removal from contaminated matrix and, thus, prevents their migration to other environments where these can create pollution and health

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hazards to living beings. However, phytoremediation has some disadvantages such as limitation to the surface area and depth occupied by the roots, slow plant growth, low biomass production, and contamination possibility of food chain by accumulated contaminants (Macek et al. 2008). Phytoremediation covers several different strategies such as phytoextraction, rhizofiltration, phytostabilization, phytovolatilization, etc. (Eapen and D’Souza 2005; Cherian and Oliveira 2005; Doty 2008; Macek et al. 2008).

4 Genetically Engineered Bacteria in Bioremediation of Heavy Metals and Organic Pollutants Water and soil are essential components of all living things on earth. But unfortunately these are contaminated by geogenic and anthropogenic activities like mining, volcanic eruption, heavy rainfall, industrializations, urbanization, and agriculture waste, which are liable for the pollution of our natural environment and toxicity in the living beings. Therefore, it is urgent need to adequately treat the contaminated water and soil to protect the environment and public health. There are several reports available on the bioremediation of heavy metals and organic pollutants by different microorganisms (Strong et al. 2000; Barac et al. 2004). Genetically engineered bacteria reported in the degradation and detoxification of organic and inorganic pollutants are listed in Table 1.1. A variety of potential strains of bacteria such as Bacillus idriensis, Ralstonia eutropha, Sphingomonas desiccabilis, Pseudomonas putida, Escherichia coli, Mycobacterium marinum, etc. have been genetically engineered for the enhanced bioremediation of toxic heavy metals in the contaminated matrix (Valls et al. 2000; Deng et al. 2003; Ackerley et al. 2004; Deng et al. 2005; Kube et al. 2005; Parnell et al. 2006; Singh et al. 2008; Schue et al. 2009; Liu et al. 2011). Bioremediation of Hg is mainly facilitated by transgene that confers arsenic resistance to microbes such as mer operon genes (Jan et al. 2009), mercuric ion transporter gene merC in Acidithiobacillus ferrooxidans (Sasaki et al. 2005), and mercuric ion transporter gene merH in Mycobacterium marinum (Schue et al. 2009). The genetically engineered radiation-resistant bacterium, Deinococcus radiodurans, also showed a great potential for the bioremediation of radioactive waste containing mercury ion (Brim et al. 2000). The genetically engineered mercury-resistant bacterium, Escherichia coli (merT-merP and MT genes), also showed a huge potential for the removal of Hg2+ from electrolytic wastewater (Deng and Wilson 2001). It has been also reported that the accumulation of Cd2+ was enhanced into Mesorhizobium huakuii when transformed with a gene that code for phytochelatins from Arabidopsis thaliana (Sriprang et al. 2003). Kang et al. (2007) reported that the recombinant E. coli can accumulate Cd up to 25-fold more than control strain. Wu et al. (2006, 2010) studied the alleviation of Cd toxicity using a metal-binding peptide (EC20) expressing rhizobacterium,

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Table 1.1 Genetically modified bacteria (GMBs) for enhanced bioremediation of organic and inorganic pollutants GMBs Pseudomonas putida PaW340(pDH5)

Introduced gene(s) pDH5 plasmid

Pollutants 4-chlorobenzoic acid

Escherichia coli JM109 (pGEX-AZR) Pseudomonas putida pnrA

Azoreductase gene Nitroreductase

Azo dyes, C.I. Direct Blue 71 TNT

Pseudomonas putida PaW85

pWW0 plasmid

Petroleum

Rhodococcus rhodochrous XplA, XplB Enterobacter cloacae NfsI

Cytochrome P450 monooxygenase

RDX

Nitroreductase

TNT

E. coli NfsA

Nitroreductase

TNT

B. subtilis BR151 (pTOO24 ) Sphingomonas desiccabilis and Bacillus Idriensis strains Methylococcus capsulatus (Bath)

Luminescent Cd sensors

Cd (Naturally polluted soils) As (Laboratory conditions)

Pseudomonas strain K-62

MerE protein encoded by transposon Tn21 (broad Hg transporter) mercuric ion binding protein (MerP)

Bacillus megaterium strain MB1

Over expression of arsM gene CrR genes for Cr (VI) reductase activity

Cr (VI) (Cell-associated Cr removal in laboratory conditions) Hg (Across the bacterial membrane) Hg

References Massa et al. (2009) Jin et al. (2009) Van Dillewijn et al. (2008) Jussila et al. (2007) Jackson et al. (2007) Hannink et al. (2007) Kurumata et al. (2005) Ivask et al. (2011) Liu et al. (2011) Hasin et al. (2010) Kiyono et al. (2009) Hsieh et al. (2009)

Pseudomonas putida 06909. Patel et al. (2010) studied that a recombinant bacterial strain, Caulobacter crescentus JS4022/p723-6H, expressing RsaA-6His fusion protein can remove up to 99.9% of the Cd as compared to control bacterium which can remove up to 37% of Cd. Arsenic removal from contaminated matrix has been also studied using recombinant microbes by several workers (Valls and de Lorenzo 2002; Qin et al. 2006; Yuan et al. 2008). A recombinant bacterium, E. coli (containing arsM gene from Rhodopseudomonas palustris), can transform highly toxic inorganic As into less toxic volatile trimethylarsine (Qin et al. 2006; Yuan et al. 2008). Further,

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a recombinant bacterium, E. coli SE5000 strain (containing nixA gene), can also accumulate Ni2+ from aqueous solution (Fulkerson et al. 1998). Further, it has been reported that the Ni resistance was enhanced in the recombinant E. coli when introduced with the serine acetyltransferase gene from Ni hyperaccumulating plant, Thlaspi goesingense (Freeman et al. 2005). Recently, Hasin et al. (2010) have characterized a methanotrophic bacterium, Methylococcus capsulatus, which can successfully bioremediate Cr6+ in a wide range of concentrations (1.4–1000 mgL 1 of Cr6+). However, a recombinant Cd-resistant rhizosphere bacterial strain, Pseudomonas putida 06909, could detoxify Cd due to its ability to produce metal-binding peptide (MBP)-EC20 that has high affinity for Cd (Lee et al. 2001). In 1970, the first GEMs called “superbug” were constructed to degrade oil by the transfer of plasmids which could utilize a number of toxic organic chemicals like octane, hexane, xylene, toluene, camphor, and naphthalene. Microorganisms that are well adapted to survive in the soil environment may not be able to survive in aquatic environment and hence cannot be used successfully. Therefore, aquatic microbes can be used to develop GEBs for bioremediation of aquatic sources. The use of such organisms would avoid the supplementation of nutrients to the inoculated environment, thereby reducing the costs incurred and maintenance required (Kulshreshtha 2013). Scientists have developed Anabaena sp. and Nostoc ellipsosporum by the insertion of linA (from P. paucimobilis) and fcbABC (from Arthrobacter globiformis), respectively. The gene linA responsible for the biodegradation of lindane (γ-hexachlorocyclohexane), and fcbABC confers the ability to biodegrade halobenzoates and can be used to remediate these pollutants from water sources. GEBs have been developed by hybrid gene clusters which alter their enzymatic activity and substrate specificities (Kulshreshtha 2013). These gene clusters encode the enzyme possessing improved transforming capability. E. coli strain is genetically modified to express a hybrid gene cluster for the degradation of trichloroethylene (TCE) (Kulshreshtha 2013). GEMs possess chemical sensors that allow the monitoring of contaminant bioavailability rather than just contaminant presence (Kumar et al. 2013). Bioluminescence-producing GEMs also help us to understand the spread of microbes in the polluted area and end point of the bioremediation (Kulshreshtha 2013). The genetically engineered Pseudomonas strains were the first microbe developed by Indian-born American scientist Dr. Anand Mohan Chakrabarty, with high catalytic potential to the subject of intellectual property right [US Patent #425944], which could degrade a variety of petroleum hydrocarbons such as naphthalene, camphor, xylene, octane, and salicylate. Following the seminal work of Chakrabarty and his colleagues on the degradation of petroleum and chloroaromatic compounds (Harvey et al. 1990; Haugland et al. 1990), the possibilities of using genetic engineering technique in biodegradation of organic pollutants had received a breakthrough with many papers published by the Timmis Laboratory in the mid- and late 1980s (Ramos et al. 1987; Rojo et al. 1987). Thus, genetic engineering techniques have been proved to be an efficient molecular approach for the microbial bioremediation of pollutants.

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5 Genetically Engineered Plants in Phytoremediation of Heavy Metals and Organic Pollutants Phytoremediation is the engineered use of green plants/trees with associated microbiota for the degradation and detoxification of organic and inorganic pollutants from the contaminated matrix (soil/water) to safeguard the environment and public health. Genetically engineered plants were first developed for the phytoremediation of heavy metals (Misra and Gedamu 1989; Rugh et al. 1996). However, the tobacco plants were the first genetically engineered plants for the phytoremediation of organic pollutants (explosives and halogenated organic compounds) (Doty et al. 2000). Genetically engineered plants are developed by introducing the transgene of interest that are responsible for the metabolism of xenobiotic compounds and offer increased resistance to pollutants (Abhilash et al. 2009). Due to the increased capacity to accumulate toxic metals from contaminated matrix, plants are chiefly preferred for the phytoremediation of heavy metals-contaminated sites. After phytoremediation, the aboveground harvestable plant biomass is safely disposed of or utilized to recover the valuable metals for future use (Salt et al. 1998). Genetically engineered plants used for the phytoremediation of environmental contaminants are listed in Table 1.2. Phytoremediation has several advantages over microbial bioremediation approaches such as high biomass of the remediating plants with less nutrient requirements, which prevent migration of pollutants from one place to another and greater acceptance among public (Alkorta et al. 2004). The best known metal hyperaccumulating plant is alpine pennycress, Thlaspi caerulescens, which hyperaccumulates Zn2+, Cd2+, and Ni2+ from contaminated matrix (Milner and Kochian 2008; Baker et al. 2000). Members of Brassicaceae, Alyssum sp. (a serpentine-endemic shrub), Astragalus racemosus, Leguminosae milkvetch, and Indian mustard Brassica juncea, are known to accumulate high concentration of heavy metals from contaminated environment (Reeves and Baker 2000). Recently, Asian stonecrop, Sedum alfredii of Crassulaceae, has gained more attention to researchers as it hyperaccumulates Pb2+ and Cd2+ and Zn2+ with more than 2% of shoot weight (Yang et al. 2003; Lu et al. 2008; Deng et al. 2008). Further, the genetically engineered, fast-growing, and high-biomass-producing metal hyperaccumulators with required genetic traits have been proved to be the suitable candidates for the phytoremediation of contaminants and include shrub tobacco Nicotiana glaucum, B. juncea, yellow poplar Liriodendron tulipifera, and sunflower Helianthus annuus (Eapen and D’Souza 2005). Several publications have reported the potential of phytoremediation to restore the polychlorophenolcontaminated soil/water (Newman and Reynolds 2004). Different plant-based remediation approaches are known including the rhizosphere biodegradation of chlorophenols inside the plant tissues (Van 2009). de Araujo et al. (2002) showed that Agrobacterium rhizogenes-transformed roots removed up to 90% phenolics, including phenol, 2-chlorophenol (2-CP), 2,6-dichlorophenol (2,6-DCP), and 2,4,6TCP, from culture medium within 120 h. Sandermann (1994) studied the plant

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Table 1.2 Genetically engineered plants (GEPs) for enhanced phytoremediation of organic and inorganic pollutants Gene AtACR2

Origin A. thaliana L.

StGCS-GS

Streptococcus thermophilus E.coli XL1-Blue Homo sapiens

MerE CYP2E1 and GST ScYCF1 YCF1

S. cerevisiae S. cerevisiae

tcu1

Neurospora crassa Neurospora crassa Pisum sativum L. Bacillus megaterium S. cerevisiae S. cerevisiae and A. sativum A. thaliana L.

tzn1 PsMTA1 TnMERI1 GSH1 GSH1 and AsPCS1 AtPCS1

Target plant Nicotiana tabacum Beta vulgaris L.

Pollutants As

References Nahar et al. (2017)

Cd, Zn and Cu

Liu et al. (2015)

Arabidopsis thaliana L. Homo sapiens Alfalfa (Medicago sativa) Populus alba X P. Brassica juncea L.

Methyl-Hg and Hg Hg and Trichloroethane

Sone et al. (2013)

Nicotiana tabacum L. Nicotiana tabacum L. Populus alba L.

Cu and Zn Cd, Fe, Ni, Cu, Mn and Pb Cu

A. thaliana

Hg

Balestrazzi et al. (2009) Hsieh et al. (2009)

A. thaliana L. A. thaliana L.

Cd and As Cd and As

Guo et al. (2008) Guo et al. (2008)

B. juncea L

Cd and As

Gasic and Korban (2007) Kawahigashi et al. (2008)

Cd, Zn and Pb Cd and Pb

CYP1A1, CYP2B6, CYP2C19 GstI-6His

Homo sapiens

Oryza sativa

Zea mays

N. tabacum

Herbicide (atrazine, metolachlor) Alachlor

TaPCS1

T. aestivum

N. glauca

Pb and Cd

P1A1, CYP2B6, CYP2C9, CYP2C19 atzA

Homo sapiens

Solanum tuberosum, Oryza sativa Medicago sativa, N. tabacum

Sulfonylurea and other herbicides Atrazine

Bacteria

Zhan et al. (2013)

Shim et al. (2013) Bhuiyan et al. (2011) Singh et al. (2011) Dixit et al. (2010)

Karavangeli et al. (2005) Gisbert et al. (2003); Martinez et al. (2006) Inui and Ohkawa (2005) Wang et al. (2005)

metabolism of 2,4-D, including hydroxylation of the aromatic ring (Phase I), conjugation with O-manolyl-glucoside (Phase II), and deposition into the vacuole (Phase III). Burken and Schnoor (1998) also studied the degradation of [14C]atrazine into less toxic metabolites inside hybrid poplar trees. Cytochrome P-450s have been reported to oxidize many chlorinated pesticides, including chlorotoluron, linuron, atrazine, and isoproturon (Kawahigashi et al. 2007).

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Banerjee et al. (2002) reported that the transgenic hairy root cultures of Atropa belladonna (developed by introducing rabbit cytochrome P-450 2E1) can metabolize trichloroethane at very fast rate as compared to its wild type. Doty et al. (2007) successfully performed the transgenic engineering of poplar plants (Populus deltoides  Populus alba) overexpressing mammalian cytochrome P450 2E1 (CYP2E1) for the enhanced degradation of trichloroethane, carbon tetrachloride benzene, and chloroform.

6 Constraints, Risks, and Challenges in the Release of Genetically Modified Organisms for Field Applications Genetically modified organisms (GMOs) can be produced by introducing the gene of interest into other organisms to accelerate their performance. A variety of GMOs have been developed through genetic engineering and utilized in the degradation and detoxification of organic and inorganic pollutants in lab conditions (Pieper and Reineke 2000; Furukawa 2003; Lovely 2003; Paul et al. 2005). The introduction of GMOs in field applications may interbreed with the wild type or sexually compatible relatives (Barac et al. 2004). The novel trait may disappear in wild types unless it confers a selective advantage to the recipient. However, tolerance abilities of wild types may also develop, thus altering the native species’ ecological relationship and behavior. Faster growth of GMOs can enable them to have a competitive advantage over the native organisms. This may allow them to become invasive, spread into new habitats, and cause ecological and economic damage. Pressure may increase on target and nontarget species to adapt to the introduced changes as if to a geological change or a natural selection pressure causing them to evolve distinct resistant populations. The effects of changes in a single species may extend well beyond to the ecosystem. Single impacts are always joined by the risk of ecosystem damage and destruction. Once the GMOs have been introduced into the environment and some problems arise, it is impossible to eliminate those (Prakash et al. 2011). One risk of particular concern relating to GMOs is the risk of horizontal gene transfer (HGT). HGT is the acquisition of foreign genes (via transformation, transduction, and conjugation) by organisms in a variety of environmental situations. It occurs especially in response to changing environments and provides organisms, especially prokaryotes, with access to genes other than those that can be inherited (Martin 1999; Ochman et al. 2000; Prakash et al. 2011). However, to overcome the associated constraint, researchers from around the globe have made several efforts to delimit the uncontrolled proliferations and survival of genetically engineered microbes (GEMs) and stop the horizontal gene transfer (HGT) to the native microbes (Kolata 1985; Atlas 1992; Paul et al. 2005). In addition, many of these risks are identical to those incurred with regard to the introduction of naturally or conventionally bred species (Sayler and Ripp 2000). But still the GMOs are neither safe nor they should be less scrutinized.

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7 Conclusion and Future Outlook Environmental contamination from around the globe has forced the scientific community to think about the environmental sustainability. Environmental sustainability and safety is a major issue in the world due to rapidly increasing pollution that create health hazards and toxicity in the environment. Environmental pollutants (organic and inorganic in nature) can be hazardous to living beings upon exposure and need to be remediated/detoxified using an array of microbes. Being of highly toxic nature, pollutants sometime can inhibit the growth of remediating microbes and, thus, halt the bioremediation processes. Therefore, genetic engineering can be a potential molecular technique to engineer the intended microbes to enhance their catalytic potential for bioremediation of environmental pollutants. However, the potential risks should also be considered before applying genetically engineered microbes in field. Acknowledgments Gaurav Saxena and Roop Kishor are thankful to the University Grants Commission (UGC) Fellowship from UGC, Government of India, New Delhi, India.

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Chapter 2

Advances in Bioremediation of Toxic Heavy Metals and Radionuclides in Contaminated Soil and Aquatic Systems Evans M. Nkhalambayausi-Chirwa, Pulane Elsie Molokwane, Tshilidzi Bridget Lutsinge, Tony Ebuka Igboamalu, and Zainab S. Birungi

Abstract Metals are used in several products essential to humans. However, processes for extraction of the metals generate effluents containing chemical by-products many of which are toxic to living organisms and are disruptive to ecosystems. Processes used in the creation of useful products from the metals leave a legacy of pollution that may take generations to clear. Metals such as mercury, cadmium, lead, chromium, and uranium, and a range of metalloids such as arsenic and selenium, are widely known for their acute toxicity at high doses and carcinogenicity at low doses. Several technologies for treatment of land and water that have been contaminated with toxic heavy metals have been proposed. Other metallic elements, although possessing no significant chemical toxicity to organisms, occur as radioactive isotopes that impart oxidative stress on organisms leading to increased incidence of mutations and carcinomas in animal tissue. The main difficulty in the treatment of metals is that the metals cannot be degraded or mineralized as is the case with organic pollutants. Metallic elements can only be oxidized or reduced to forms that are less mobile and easier to extract from the environment. This chapter is compiled from information from projects in which metals were either oxidized or reduced to less mobile and less toxic states using pure or consortium cultures of bacteria followed by immobilization or extraction using physical or biological media. The uptake of metals for reuse was attempted using bioengineered molecular adsorbents on cell surfaces. The latter process was developed to facilitate selective uptake of different metallic species as a low energy biorefinery. E. M. Nkhalambayausi-Chirwa (*) · T. B. Lutsinge · T. E. Igboamalu · Z. S. Birungi Water Utilization and Environmental Engineering Division, Department of Chemical Engineering, University of Pretoria, Pretoria, Gauteng, South Africa e-mail: [email protected] P. E. Molokwane Water Utilization and Environmental Engineering Division, Department of Chemical Engineering, University of Pretoria, Pretoria, Gauteng, South Africa Oloenviron Consultancy, Johannesburg, South Africa © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_2

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Keywords Toxicity pathways · Bioremediation · Bioseparation · Toxic metals · Bioengineering · Enzyme kinetics

1 Background Metals are used in industrial processes either as structural components such as steel or as additives for modification of strength properties, corrosion resistance, and control of chemical reactions and for ornamental purposes. Metals such as chromium (Cr) and radium (Rd) have been used in the making of paints and pigments, wood preservation, and luminescence in displays (Beszedits 1988). Addition of metals and carbon compounds is critical in the formation of stronger alloys that are less susceptible to corrosion. For example, metals such as chromium, silver, and tin have been widely used to give steel a polished silvery mirror coating or to galvanize it against corrosion. Ornamental use of Cr and Pb includes the production of emerald green (glass) and synthetic rubies in the case of Cr (Morrison and Murphy 2010) and giving glass a reddish tinge in the case of Pb (Francis and Croft 1849). The discovery and utilization of metals by humans is so important such that at least three of the human eras of civilization have been named after a metal, i.e., iron age, bronze age, and nuclear age. One evolutionary epoch, the Anthropocene age, is associated with the humans’ ability to harness the power of the atom through fission of the uranium atom which is also loosely associated with our ability to impact the global climate (Wilke and Johnstone 2017). The production of metals and utilization of dissolved metals typically produces contaminated wastewater (Goutam et al. 2018; Bharagava et al. 2017a; Gautam et al. 2017; Saxena and Bharagava 2015; Saxena et al. 2016). Metals can serve as oxidants or reductants in environmental systems resulting in damage to cells or inhibition of cellular processes (Jaishankar et al. 2014). At high concentrations, metals such as Cr, U, Pb, Hg, Se, and As can cause acute toxicity resulting in death of organisms (Chirwa and Molokwane 2011; Mtimunye and Chirwa 2013; Brink et al. 2017; Bansal 2010; Wessels 2017). At low concentrations, metals can cause mutagenicity, carcinogenicity in organisms, and teratogenicity in mammals (Flessel 1979). For this reason, all national and international environmental regulatory organizations place stringent regulations on the disposal of metals to the environment (Federal Register 2014; Hiz and Aki 2015; Gao et al. 2014; ACGIH 2004).

2 Mining of Ores and Environmental Impacts During the mining operation for any particular metal, the target metal is extracted together with several metal pollutants from geological seams. For example, chromium is mined from seams containing trace amounts of platinum group

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metals and other base metals. In South Africa, the country with the largest reserve of chrome ore, the main seams include: • Chromite: (Mg,Fe2+)(Cr,Al,Fe3+)2O4, the main component of Merensky Reef of the Bushveld Igneous Complex in South Africa (von Gruenewaldt and Hatton 1987; Wagner 1923) • Barbertonite: Mg6Cr2(CO3)(OH)16.4H2O, the major component of the Eastern Bushveld Igneous Complex (Meli 2009; Calas et al. 1984) • Nichromite: (Ni,CoFe2+)(Cr,Fe3+,Al,)2O4, which was discovered first in the Bon Accord nickel deposit in Barberton District, South Africa (Tredoux et al. 1989) In Brazil, the richest chrome ore seam is in the Campo Formoso layered intrusion which also contains chromite (Cramer et al. 2004; Boukili et al. 1984). Consequently, tailing dumps and process waste stockpiles at chrome mining and ferrochrome processing sites contain significant levels of other platinum group and rare earth metals (Chirasha and Shoko 2010; von Gruenewaldt and Hatton 1987). Chromium in the tailing dumps can exist either as Cr(III) or Cr(VI) depending on the environmental conditions within the dump. Interaction with other metals at different oxidation states can influence the stability of the oxidation state of the chromium species inside the waste dump and its ability to leach into the surrounding water bodies (Tiwary et al. 2005; Ma and Garbers-Craig 2006). In areas where the leachate water enters agricultural supply water, there is a high risk of contamination of food products and bioaccumulation into higher order organisms (Chirasha and Shoko 2010).

3 Metal Polluting Activities Due to the wide range of uses of metals, pollution comes in many forms. The most important feature is whether a metal poses a particular threat and how such a threat could be best dealt with. In the following subsections, we classify pollutants based on whether they are toxic to organisms due to chemical poisoning or radiation poisoning. Chemical toxicity is best dealt with by neutralization using reducing or oxidizing agents, whereas radiation toxicity, on the other hand, is remediated by isolation of the radiation source and confinement for a period equivalent to the time required for the radiation to dissipate.

3.1

Chemically Toxic Elements

There are various metallic elements that confer chemical toxicity. Technically, any element at excessive loading can cause toxicity. For some metals, the levels at which they produce observable negative impacts are very low. For metals such as Cr(VI),

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Hg(II), Cd(II), and Pb(II), the concentrations at which they produce observable negative effects are quite low. Their high toxicity thus serves as an example and warrants special consideration for treatment in this chapter.

3.1.1

Chromium

Chromium has been used extensively in industrial processes such as leather tanning, electroplating, negative and film making, paint and pigment processing, and wood preservation (Beszedits 1988). Additionally, chromium has been used as a metallurgical additive in alloys (such as stainless steel) and metal ceramics. Chromium plating has been widely used to give steel a polished silvery mirror coating. The radiant metal is now used in metallurgy to impart corrosion resistance. Its ornamental uses include the production of emerald green (glass) and synthetic rubies. Due to its heat-resistant properties, chromium is included in brick molds and nuclear reactor vessels (Dakiky et al. 2002). Through the above and many other industrial uses, a large amount of chromium (approximately 4500 kg/d) is discharged into the environment making it the most voluminous metallic pollutant on earth. Almost all chromium inputs to the natural systems originate from human activities. Only 0.001% is attributed to natural geologic processes (Merian 1984). Chromium from the anthropogenic sources is discharged into the environment mainly as hexavalent chromium [Cr(VI)]. Cr(VI) – unlike Cr(III) – is a severe contaminant with high solubility and mobility in aquatic systems. Cr(VI) is a known carcinogen classified by the US EPA as a Group A human carcinogen based on its chronic and subchronic effects (Federal Register 2004). It is for this reason that most remediation efforts target the removal of Cr(VI) primarily.

3.1.2

Uranium

Uranium exists in the environment mainly as oxides, organic or inorganic complexes, and rarely as a free metal ion (Mtimunye and Chirwa 2013). Free elemental uranium primarily exists in higher oxidation states typically bound to oxygen. The oxygen bound uranium exists mainly as uraninite (UO2), triuranium octaoxide also known as pitchblende (U3O8), and uranium trioxide (UO3) (Stefaniak et al. 2009). U3O8 is relatively insoluble in water and relatively stable over a wide range of environmental conditions. UO2 on the other hand is not as stable as U3O8 in the environment as it may undergo alteration under various environmental conditions (Senanayake et al. 2005). Upon exposure to air, UO2 is subjected to oxidation and as a result produces a secondary mineral (UO22+) which complexes easily with phosphates, carbonates, silicates, and sulfates (Senanayake et al. 2005; Stefaniak et al. 2009).

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Selenium

There are many natural pathways in which selenium is released into the environment. The processes range from volcanic activity; wildfires; volatilization from soils, plants, and water bodies; and weathering of rocks (Sandy and DiSante 2010). Selenium occurs naturally in soil, but at low concentrations, ranging between 0.1 and 2 μg/g soil, whereas in seleniferous soil the concentration can go up to 320–324 μg/g soil (Girling 1984). Of the selenium released into the environment, 37.5–40.6% can be ascribed to anthropogenic activities. It is released from the earth’s crust by the mining of coal, by oil production, through the use of agricultural products, as well as during the melting of nonferrous metals (Lenz and Lens 2009). Landfill ash disposal generating toxic leachate poses a risk of groundwater contamination (Lemly 2004). According to different studies, acid mine drainage waters contain selenium at concentrations ranging between 2.0
104 and 6.2
103 mM (reported as total selenium) (Lenz et al. 2008). Wastewater from oil refineries in the San Francisco Bay (USA) contains relatively low concentrations of selenium of about 50–300 μg/L (Lawson and Macy 1995). The wastewater from a selenium refinery plant in Japan contained an average of 30 mg/L selenium, with the most of it present as selenite (Satoshi et al. 2012).

3.1.4

Mercury

Mercury is released to the environment from activities such as alkali and metal processing, incineration of coal, and medical waste and other wastes. In the Southern hemisphere, a lot of mercury is released into the environment as effluents from the mineral purification processes. Mercury is used extensively in the separation of gold from other metallic impurities. The process involves extraction of gold in liquid mercury and boiling away the mercury to leave behind the pure gold. Fumes produced result in atmospheric mercury pollution. Once in the atmosphere, mercury is widely disseminated and can circulate for years. Other sources of atmospheric mercury include natural sources such as volcanoes, geologic deposits, and volatilization from the ocean. Although all rocks, sediments, water, and soils naturally contain small but varying amounts of mercury, scientists have found some local mineral occurrences and thermal springs that are naturally high in mercury. In many relatively pristine areas, however, mercury concentrations have actually increased because atmospheric deposition has increased. For instance, concentrations of mercury in feathers of fish-eating seabirds from the northeastern Atlantic Ocean have steadily increased for more than a century. In North American sediment cores, sediments deposited since industrialization have mercury concentrations about 3–5 times those found in older sediments. Some sites may have become methylmercury hot spots inadvertently through human activities (Gilmour and Henry 1991). Lake acidification, addition of substances like sulfur that stimulate methylation, and mobilization of mercury in soils in newly flooded reservoirs or constructed wetlands

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have been shown to increase the likelihood that mercury will become a problem in fish. Although scientists from USGS and elsewhere are beginning to unravel the complex interactions between mercury and the environment, a lack of information on the sources, behavior, and effects of mercury in the environment has impeded identification of effective management responses to the nation’s growing mercury problem (USGS 1995).

3.1.5

Lead

Natural lead occurrence is not common on earth surface. Anthropogenic lead is introduced through the use of products containing lead additives such as paints, gasoline, and metal works (Gioia et al. 2006; Boyle et al. 2014). Metallic lead and lead alloys are extensively used for storage battery plates, sheathing for electrical cables, small-caliber ammunition, shielding for X-ray apparatus and atomic reactors, type metal, bearing metal, and solder. Lead is also released as a waste product from extraction of metals such as silver, platinum, and iron from their respective ores. In car engines, the burning of fuel in the presence of other impurities (chlorines, bromines, oxides) produces lead salts (Wang et al. 2010). The lead salts enter the environment through the exhausts of the cars in particulate form. Lead is also released from combustion of coal during electric power generation in coal-powered plants. Typical concentrations of lead released from coal-powered plants have been reported at 7–110 mg/kg with release rates up to 11,200 tons lead/year (UNEP 2010). The increased non-localized outputs of lead to the environment have generated a lot of concern mainly due to its toxicity to living organisms. Lead contamination of water and land is known to produce sub-chronic effects at low concentrations (Supanopas et al. 2005), acute effects at moderately high concentrations (Papp et al. 2006), and catastrophic changes to ecosystems at very high concentrations (Eisler 1988). Lead concentration above 1000 ppm can wipe out almost the entire ecosystem, whereas, concentrations lower than 10 mg/L can result in endocrine disrupting activity in sensitive species. In many areas around the United States, accumulation of lead in aquatic systems has risen to levels detrimental to human inhabitants (Renner 2010).

3.2

Radioactive Elements

In high temperature gas-cooled reactors (HTGR), also known as fast reactors, graphite is utilized as the moderator of the nuclear reaction. The graphite is either used as part of the structural materials for the reactor core vessel or as fuel containment elements in the form of pebbles. The graphite used from natural sources contains non-carbon impurities within the carbon matrix. Among these impurities are oxygen and nitrogen from entrapped air, cobalt, chromium, calcium, iron, and sulfur (Khripunov et al. 2006).

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Fuel Grains Grain Graphite Matrix Boundaries Microscopic Cracks

Graphite Pebble

TRISO Coated Particle

Fuel Kernal

Porous Carbon Buffer

Pyrolytic Carbon/Silicon Carbide Coating

Fig. 2.1 Propagation of fission products and impurities in a graphite-regulated high temperature (gas)-cooled reactor fuel element

The radioactive fission products are created within the fuel grains and migrate through grain boundaries and then through microscopic cracks in the graphic matrix (Fig. 2.1). Most of the fission products are entrained in the matrix – a small proportion escapes through the outer layers into the gas phase. Reprocessing of irradiated nuclear graphite entails the separation of the metallic radionuclides from the graphite matrix and reducing the amount of C-14. Impurities in the fuel itself include (1) metallic fission products (Mo, Tc, Ru, Rh, and Pd) which occur in the grain boundaries as immiscible micron- to nanometer-sized metallic precipitates (ε-particles); (2) fission products that occur as oxide precipitates of Rb, Cs, Ba, and Zr; and (3) fission products that form solid solutions with the UO2 fuel matrix, such as Sr, Zr, Nb, and the rare earth elements (Buck et al. 2004; Bruno and Ewing 2006).

3.2.1

Uranium-238 and Transuranic Elements

Uranium is the heaviest naturally occurring element in the solar system. All other elements with higher molecular weight than uranium are referred to as transuranic (TU) elements and are synthetically produced. Uranium and all transuranic elements are highly fissionable, that means, they can be split into smaller elements releasing neutrons and large amounts of energy in the process. Among the milliard of fission products, there are a few common species that are persistent in nature due to their relatively long half-lives such as strontium-90 (Sr-90), cesium-120 (Cs-120), and radiocarbon-14 (C-14).

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Uranium used in electricity generation in nuclear reactors is generally sourced from ores with uranium oxide concentrations up to 10%. However, lower grade ores with uranium oxide concentration of 0.2% or less are the most common and the most mined (Ross 2015; Thomas 1981). The uranium concentrate used in nuclear reactors is typically at 75–95% uranium oxide (U3O8). Mining of such readily available lower-grade uranium ore to achieve about 75–95% U3O8 concentrate used in nuclear power plants results in the release of large amounts of waste rock tailings with significant environmental consequences due to the presence of residual uranium in the rocks (Thomas 1981). Uranium is sourced from rich ores with concentrations up to 10%. However, ores with uranium oxide concentration as low as 0.2% are also mined and are the most common (Sovacool 2008; Tudiver 2009). Uranium producers have been able to utilize ores with uranium oxide concentration as low as 0.0004%. Uranium is recovered from ore by comminution of the rocks followed by leaching using alternative solutions of acid and/or alkaline chemicals. The end product from ore milling and leaching results into a bright yellow powder called yellow cake (U3O8) which is about 75–90% uranium oxide (Sovacool 2008). Before this uranium oxide concentrate can be used in a reactor for generating electricity, it must first be converted into uranium hexafluoride (UF6), which is used in a gaseous diffusion enrichment process. During the uranium enrichment process, U-235 concentration is increased to least 3.5% for atypical commercial light-water reactor and up to 4–5% for other modern reactors, while at the same time, the U-238 isotope is decreased notably. Suffice to say, U-235 is the only natural occurring isotope that can sustain a fission chain reaction by capturing neutrons and splitting into two parts yielding large amount of energy (Soudek et al. 2006; WNA 2008). On average, the specific radioactivity of natural uranium is 25 kBq/g, double that of U-238. During its decay process, uranium may generate 0.1 watts/tonne which is enough to warm the earth’s mantle (WNA 2008). After the enrichment process, about 85% of oxide comes out as waste in the form of depleted UF6, and the remaining 15% emerges as enriched uranium and is converted into ceramic pellets of UO2. Fresh UO2 which contains up to 5% of U-235 is then packaged in zirconium alloy tubes and bundled together to form fuel rod assembles for reactors. Thereafter, the used reactor fuel which contains up to 95% U-238, 3% fission products and transuranic isotopes, 1% plutonium, and 1% U-235 is removed and stored to be reprocessed prior to disposal (Soudek et al. 2006; WNA 2008). During the reprocessing stage, uranium (U-235) and plutonium (Pu-239) are separated from the spent fuel using the PUREX method and then reused as mixed-oxide (MOX) fuel in the reactor. This process is referred to as the closed fuel cycle (Fig. 2.2). The majority of radioactive organic waste is produced in the enrichment and reprocessing operations. All values in the tables are reported in cubic meters per Gigawatt electricity-year (m3/GWe-year).

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Waste Rock

Open pit/ underground Mining

Uranium Mill Tailings

Milling

Uranium ore

29

Uranium Ore Concentrate (U3O8) Conversion to Uranium hexafluoride (UF6)

UF6 Enrichment (eUF6)

Fuel fabrication

Uranium Reprocessing

MOX Plutonium

Enriched UO2, increase 0.7%U235 to 3-5% U235 Reactor Spent Fuel Interim storage

Disposal

Fig. 2.2 Nuclear fuel closed cycle

3.2.2

Sr-90and Cs-120

The radioactive isotope of strontium (Sr-90) is of great concern, and its toxic effects on human beings are well documented (Chen 1997; Kossman and Weiss 2000; Noshchenko et al. 2001; Greve et al. 2007). Strontium can be highly mobile in both soil and groundwater systems, and it has a half-life of 28 years. Due to its chemical similarity with calcium, it is easily incorporated into bone material in mammals. When incorporated in the organisms in this manner, it continues to irradiate localized tissues with the eventual development of bone sarcoma and leukemia (Chen 1997). The main disadvantages of using conventional adsorbents, such as zeolites and synthetic organic ion exchangers, for strontium removal from radioactive waste is their unsuitability at high pH, high sodium concentrations, and in irradiated environments (Chaalal and Islam 2001). However, the main disadvantages with the above methods is their unsuitability at high pH, high sodium concentrations, and in irradiated environments.

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Table 2.1 Carbon-14 production mechanisms and thermal cross sections Target isotope N 12 C 13 C 17 O 14

Mechanism N(n, p)14C 12 C(n, γ)14C 13 C(n, γ)14C 17 O(n, α)14C 14

Thermal cross section (barns) 1.81 n/k 0.0009 0.235

Isotopic abundance (%)* 99.6349 n/k 1.103 0.0383

Adapted from International Union of Pure and Applied Chemistry n/k Not known

3.2.3

Radiocarbon-14 and Irradiated Organic Pollutants

In nuclear fuel processing, the problem is the generation of large volumes of the partially water-soluble wastes that often contain toxic organic compounds. The use of decontamination reagents such as carbon tetrachloride (CCl4) together with phenolic tar results in wastewater with high content of chlorophenols (Makgato and Chirwa 2015). Chlorophenols are compounds of serious environmental concern due to their toxic and carcinogenic impact on living organisms (Olaniran and Igbinosa 2011). Operation of reactors at high neutron influx can result in the transmutation of C-12 to C-14, a common problem in the graphite-moderated generation-IV nuclear reactors. Additionally, upon exposure to high neutron flux, most of the impregnated nonmetallic impurities are expected to transmute to unstable radioactive forms. For example, experimental exposure of graphite in nuclear reactors has shown that the stable forms of oxygen, nitrogen, and C-12 are converted to radiocarbon-14 (C-14) as shown in Table 2.1. This table shows the available natural process that lead to production of C-14. Although not a heavy metal, the removal of C-14 can be managed together with the remediation of other radioactive metallic impurities (Molokwane and Chirwa 2007).

4 Conventional Treatment of Toxic Metals Dissolved metallic species are commonly treated by chemical conversion to precipitable chemical species and followed by extraction at pH ranges where the metal exists as a solid precipitate. For example, a metal such as Cr(VI) could be reduced to Cr(III) at a relatively low pH through the following reduction-oxidation (redox) reaction: Cr2 O7 2 þ 14Hþ þ 6e ! 2Cr3þ þ 7H2 O þ 1:33v E0




ð2:1Þ

(Garrel and Christ 1965), followed by precipitation as chromium hydroxide (Cr (OH)3(s)) at a higher pH. Because of the difference in electric potential between the two states, substantial amounts of energy are needed to overcome the activation energy for the reduction process to occur. It is therefore assumed that spontaneous reduction of Cr(VI) to Cr(III) never occurs in natural aquatic systems at ambient pH

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and temperature. Common reducing agents such as iron sulfide (FeS) and pyrite (FeS2) have been used to precipitate Cr economically since both are naturally occurring reductants. The resultant reactions using FeS and FeS2 are shown below: CrðVIÞðaqÞ þ 3FeðIIÞðaqÞ ! CrðIIIÞðaqÞ þ 3FeðIIIÞðaqÞ
xCrðIIIÞ þ ð1  xÞFeðIIIÞ þ 3H2 O ! Crx Fe1x ðOHÞ3 ðsÞ þ 3Hþ

ð2:2Þ ð2:3Þ

The shortcoming of chemical processes for the treatment of metals is that the process is environmentally intrusive due to the production of large volumes of toxic sludge which is difficult to dispose. Additionally, the treatment becomes increasingly expensive as the remaining concentration of the pollutant becomes smaller.

5 New Treatment Approaches In order to overcome the challenges posed by conventional methods of treatment of toxic metals, biological processes have been proposed (Shen and Wang 1993; Chirwa and Wang 1997, 2001; Igboamalu and Chirwa 2017). Biological treatment is based on the principle of emulating the natural occurring processes to treat waste. During 3 billion years of existence, microorganisms have evolved mechanisms to survive in hostile environments and to adapt to changes in the environment (Bush 2003). Environmental engineers around the world have undertaken to find ways to tap into the mysteries of nature by diligently studying the action of microorganisms as they adapt to extreme conditions. One of the most conserved mechanisms in the living cell is the biochemical pathway for electron transport through the cytoplasmic membrane to conserve energy through the oxidation of an electron donor and reduction of an electron acceptor such as oxygen. This process has been conserved over billions of years, such that, to this day, all life on earth depends on variants of this pathway (Bush 2003; Thomas et al. 1985; Nealson 1999; Kalckar 1974). Most biochemical processes for degradation and/or detoxification of compounds in are linked to the above process. Lately, microorganisms have been isolated that are capable of reducing the toxic forms of heavy metal and transitional metal elements to less mobile precipitable forms (Foulkes et al. 2016). Other researchers have found microbial cultures with the capability to resist high radiation doses (Battista 1997; White et al. 1999).

6 Bioremediation Processes for Removal of Toxic Metals Bioremediation is the use of microbes and plants to degrade/detoxify the organic and inorganic pollutants in contaminated matrix (Bharagava et al. 2017b, c; Saxena and Bharagava 2017; Chandra et al. 2015). As indicated above, the utilization of

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biological systems to remove toxic metals from water, soil, and environmental systems has been investigated extensively since the late 1980s. Metals can be removed through reduction to precipitable species and oxidation to precipitable species or by biosorption taking advantage of the ion-exchanging properties of cell surfaces of bacteria, fungi, and/or algae. The following are examples of processes that have been studied extensively by our research group at the University of Pretoria and our collaborators overseas.

6.1

Biological Reduction, Separation, and Recovery

Any highly oxidized metallic element such as Cr(VI), U(VI), Tc(VII), and Se(VI) can be reduced to a lower oxidation state utilizing the microbial cell’s NADHdehydrogenase (NADHþ-dh) (Fig. 2.3). NADH+ is readily oxidized to NAD, thereby donating two electrons to the membrane electron-transporting proteins such as NADH+-dh, ubiquinone, and cytrochromec-c3, which in turn channels electron to the target metallic species directly or through an enzyme typically known as a Mn+ reductase where Mn+ is the target metal of the valency state n (Cervantes et al. 2001; Barak et al. 2006). For hexa- and heptavalent metals, it was demonstrated that living cells of metal-reducing organisms can reduce the metals, either as a necessity to detoxify the cell’s immediate environment (Cervantes 1991) or as a source of energy for cell growth and maintenance (Horitsu et al. 1987). Energy for Mn+ reduction is

Fig. 2.3 Electron flow pathway resulting in reduction of a metal species. The n-valent metal is reduced to an (n-xe) valent species by receiving xe electrons from electron donors in the system

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derived from the oxidation of NADH+ to release electrons to the cytochrome c,c3 system which functions as a conduit for electron flow to Mn+ reductase (Lovley and Philips 1994; Chirwa and Wang 1997; Mtimunye and Chirwa 2014). The results obtained from the U(VI) reduction (Mtimunye and Chirwa 2014), Cr (VI) reduction (Chirwa and Molokwane 2011), and reduction of other metallic species (Cervantes and Silver 1992; Lloyd et al. 1999; Yong et al. 2002) showed that the reduction pathway for most metals is non-specific and is closely related to the sulfate reduction transmembrane electron shuttle.

6.2

Biological Oxidation

The detoxification of arsenic by oxidation of As(III) to As(V) is used here as an example of the bioremediation through an oxidation reaction process. In this example, As(III) served as an inorganic electron donor for beneficial oxidation to less toxic and immobile As(V). Thermodynamically, the conversion of As(III) to As (V) is an exothermic reaction and could generate considerable amount of energy ranging from – (254 to 468) KJ/mol for cell growth and metabolism (Dastidar and Wang 2010; Wang et al. 2013). The first heterotrophic As(III) oxidation was observed in a cow dip in South Africa in 1918 (Green 1918), whereas an autotrophic As(III) oxidation was observed in 1981 (Ilialetdinov and Abdrashitova 1981). Heterotrophic As (III) oxidation may represent a detoxification reaction on the cell’s cytoplasmic (inner) membrane, whereas autotrophic As(III) oxidation releases energy that is used for CO2 fixation and cell growth under both aerobic and anaerobic conditions (Santini et al. 2000). In 2010, Dastidar and Wang reported that about 256 KJ/mol energy can be generated during oxidation of As(III) to As(V) by a purified culture of Thiomonas arsenivorans strain b6 (Dastidar and Wang 2010). Further studies showed that about 467.95 KJ of energy could be generated in the process (Wang et al. 2013). The redox conversion of As(III) in aqueous environment was represented by the following equation: As3þ ! As5þ þ 2e þ energy  ð254 to 468ÞKJ=molð3Þ The oxidation of As(III) could operate in competition with the oxidation of the other organic and inorganic electron donors as energy sources. For example, nitrate (NO3) and chlorate (ClO3) in solution can act as competitive electron donors to As(III) (Sun et al. 2010). In the recent study, the use of As(III) as an electron donor for reduction of Cr(VI) was demonstrated using a consortium of As(III) oxidizing species isolated from a cow dip in Tzaneen, Limpopo Province (South Africa) (Fig. 2.4) (Igboamalu and Chirwa 2017). The predominance of As(III) as an electron donor for metabolic processes over other electron donors in the system has not been investigated.

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Fig. 2.4 Conceptual representation of biologically catalyzed Cr(VI) reduction with A(III) as an electron donor

6.3

Methylation

Partitioning of metallic species from the aquatic phase can be achieved through biologically mediated methylation followed by volatilization. Detoxification by methylation has been observed for mercury and selenium. For selenium, it is firstly necessary to reduce all forms of selenium to elemental selenium (Se(0)) in order to achieve methylation and volatilization. Methylation is identified as a critical step in the transport of selenium out of terrestrial and aquatic environments contaminated with selenium. Methylation occurs in water, sediment, and soil and is mostly a biotic process. A wide range of organisms from bacterial, algal, and fungal groups are capable of catalyzing the methylation process of selenium. Examples of the compounds that are formed from the methylation of selenium include dimethylselenide (DMSe), dimethyldiselenide (DMDSe), dimethylselenone, trimethylselenonium, selenous acid, selenium dioxide, and selenic acid, to name a few.

6.4

Biosorption

Traditional treatment technologies such as membrane filtration, GAC adsorption, precipitation-sedimentation, flotation, ion-exchange, and electrochemical deposition systems have been used for decades for the removal of toxic heavy metals from water (Veglio et al. 2003; Wang et al. 2004; Aziz et al. 2008). Unfortunately, these technologies are associated with high cost for processing of metal concentrations, lack of specificity, production of large volumes of sludge, and low performance at low metal concentrations (Kotrba 2011; Gaur et al. 2014). Adsorption processes are proving to be the most efficient for advanced wastewater treatment (Anastopoulos et al. 2015). Adsorption is a more advanced reaction-mediated process at the surface

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of a biosorbent such as granular-activated carbon (GAC), natural clay materials such as kaolinite and zeolite, biomass, and other natural and artificial materials (Borda and Sparks 2008; Gadd 2009; Gaur et al. 2014).

6.4.1

Conventional Adsorbents

Recent adsorption technologies which act as possible alternatives and/or complementary to conventional technologies include granular-activated carbon (GAC), nanotechnology, and biological adsorbent systems. GAC is the most commonly used adsorbent due to its large surface area and high adsorption capacity attaining high metal removal efficiency. However, the cost of production and regeneration of spent carbon is still high and has limited its use in wastewater treatment (Babel and Kurniawan 2004; Fu and Wang 2011). Nanotechnology offers great promise in treatment of wastewater containing inorganic pollutants but faces some challenges of cost-effectiveness and potential environmental and human risk (Qu et al. 2013). The current cost of most nanomaterials is significantly high and owing to the small size of these particles may facilitate transport of toxic materials in the environment causing harm to cellular materials. Biological adsorbent systems use materials of biological origin either dead or live for the treatment of wastewater containing heavy metals. The technology can be categorized under biosorption and bioaccumulation.

6.4.2

Novel Biosorption of Metallic Species

The first major challenge in the biosorption field is to screen and select the most promising biomass with high binding capacity for metals and possibility of reuse (Kratochvil and Volesky 1998). A broad range of biomass types have been tested including fungi, bacteria, yeast, and agricultural wastes such as cane molasses (Abdel-Rahman et al. 2016), maize stalks (Haryanto et al. 2017), wood chips, grass, and maize tassels (Guyo et al. 2015). The sorption of metals onto these biomaterials is attributed to the constituents of the cell wall which are mainly composed of carbohydrates, proteins, and phenolic compounds (Choi and Yun 2006). The biosorption isotherms are used for basic evaluation of sorption systems under optimal environmental conditions. Any comparison done at two different sorption systems can only be done at the same initial concentration for screening of sorbents with highest sorption capacity. These experiments are usually carried in batch reactors as an initial step before application into the dynamic continuous systems (Gadd 2009). Equilibrium isotherm models are classified into empirical and mechanistic equations. In the mechanistic models, mechanisms for biosorption are explained and can predict the experimental behavior (Pagnanelli et al. 2001; Volesky 2007).

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E. M. Nkhalambayausi-Chirwa et al.

Algae as Biosorbents

In our latest studies, we evaluated the performance of several species of algae in the bioflocculation and biosorption of toxic metals and rare earth metals (REMs). Empirical models were used to evaluate the biosorption kinetics and biosorption capacity of wet and dry cells. The Langmuir and Freundlich are the most widely used and accepted simplistic mathematical models in literature as they usually fit the experimental done relatively well (Volesky and Holan 1995; Arıca et al. 2001; Mehta and Gaur 2005). These isotherm models reflect metal binding as a function of equilibrium concentration of a known metal in solution. A Langmuir model assumes monolayer adsorption solutes onto surface containing a finite number of identical sites with homogenous distribution of adsorption energy and no transmigration of adsorbed ions. Freundlich model assumes the energy distribution is heterogeneous with binding sites having a higher affinity for the metal first occupied (Selatnia et al. 2004; Goksungur et al. 2005; Stumm and Morgan 1995). Other common isotherm models used include Temkin model whose consideration is given to the interaction between adsorbate and adsorbent; Dubinin-Radushkevich (D-R) model which expresses the adsorption mechanism with a Gaussian energy distribution on a heterogeneous surface, Table 2.2 (Gadd 2009; Pagnanelli 2011). Adsorption capacity and affinity of the different algal species were evaluated for removal/recovery of lanthanum (La), thallium (Tl), and cadmium (Cd). The Langmuir model performed better than the Freundlich model with a correlation coefficient of  0.90 (Table 2.2). The results indicated that Chloroidium saccharophilum had the highest potential for metal removal of La, Cd, and Tl with a sorption capacity (qmax) of 129.81, 1000, and 128.21 mg/g, respectively (Table 2.3). Tl adsorption capacity was significantly higher than other metals with a qmax in the range of 833 and 1000 mg/g. Desmodesmus multivariabilis showed a relatively higher affinity for La and Cd adsorption at 4.55 and 1.49 L/g, respectively. The algal species with both a higher qmax and higher affinity were considered as they indicate possible removal and recovery of metals (Birungi and Chirwa 2014, 2015).

6.5

Electrokinetic Mobilization

The first electrokinetic phenomenon was observed at the beginning of the nineteenth century when Reuss in 1808 applied a direct current to a clay water. However, Table 2.2 Most common equilibrium models for single metallic studies Isotherm type Langmuir Freundlich Dubinin-Radushkevich (D-R) Temkin

Linearized equation =q ¼ Ce =q þ 1=bqmax

Ce

e

max

logqe ¼ logk þ 1=n log C e Inqe ¼ In Xm + βε2 q ¼ B In AT + B In Ce

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Table 2.3 Langmuir parameters for La, Tl, and Cd adsorption using microalgae

Adsorbent Desmodesmus multivariabilis Chloroidium saccharophilum Scenedesmus acuminutus Stichococcus bacillaris

La adsorption qmax b (mg/g) (L/g) 100 4.55

Thallium adsorption qmax b (mg/g) (L/g) 909.1 0.524

Cd adsorption qmax b (mg/g) (L/g) 35.57 1.49

129.87

0.142

1000

1.667

128.21

0.016

111.1 51.02

0.12 4.56

833.3 833.3

0.290 0.293

– –

0.049

experimenting with the electroosmotic velocity of fluid and the zeta potential under an imposed electric gradient was first proposed by the two scientists, namely, Helmholtz and Smoluchowski [cica. 1879], as quoted in Walls (2010). The technology was demonstrated to be successful in removing large quantities of heavy metals from the soil by electric fields in laboratory studies. The technology involves the application of low-voltage direct current through electrodes that are placed across a section of contaminated ground, and the charge moves the contaminant. The principle of the technology is an electric current is used to mobilize ions. In order for electrokinetic remediation to be carried out, the pore fluid should be present as it has the following functions: conducting the electrical field, transporting species that are injected, and controlling and modifying the electrode reactions. Ions are transported by electromigration, electroosmosis, electrophoresis, and other modes of transport like diffusion. Electromigration of ionic species is defined as the movement of ions in the pore fluid of the soil under the influence of an electric current; cations move toward the cathode and anions move toward the anode. Cationic and anionic contaminants are both removed by electromigration. Compared to electroosmosis in terms of cation contaminant migration, electromigration has been reported to have greater charge of ionic species present, field strength, and ionic concentration which influences electromigration during electrokinetics (Page and Page 2002). The movement is described by the following equation: um ¼ vE

ð2:4Þ

where um is the velocity of an ion and v is ionic mobility (Page and Page 2002). Factors such as concentration, ionic charge, and temperature affect the electrical conductivity of the solution, and in turn this is related to the ionic mobility. The second transport mechanism is electroosmosis which is the movement of pore water under an electrical potential difference from the anode to the cathode. This process is affected by the soil porosity and the zeta potential of the soil medium. It occurs due to the drag interaction between the bulk of the liquid in the pore and a thin layer of charged fluid next to the pore wall. The ions move under the action of the electric field in a direction parallel (Probstein and Hicks 1993; Reddy and Parupudi 1997). The electroosmotic flow rate depends on the balance between the

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electrical force on the liquid and the surface of the soil particles (Page and Page 2002). Electroosmosis has been found to be effective in removing cation at low concentrations. The rate of electroosmotic flow in the soil is described by Darcy’s law for hydraulic flow (Page and Page 2002): qA ¼ k e EA

ð2:5Þ

where ke is the coefficient of electroosmotic permeability (or conductivity), E is the electric field strength or negative potential gradient, and A is the total cross-sectional area normal to the flow direction. Electrophoresis refers to the transport of charged particles under the influence of an electric current, these charged particles colloid in soil-liquid mixture, and it is an important mechanism in remediation of sludge. The drawback of this mechanism is that in compact systems like clay soil, movement of contaminants is restrained. However electrophoresis is important in remediating colloids that have contaminants adsorbed to them (Pamukcu and Wittle 1992; Reddy and Parupudi 1997). Other mechanisms that are involved in electrokinetics are diffusion which plays a significant role in cationic and anionic contaminant transport, advection which moves soil moisture or groundwater due to hydraulic forces, and finally convection which is responsible for the movement of soil moisture or groundwater due to buoyancy forces. Trivalent chromium migrates toward the cathode due to electromigration, and CrO42 and Cl- migrate toward the anode due to electromigration, and the negative-charged colloids move due to electrophoresis. The electrolysis reaction at the electrodes generates hydrogen ions (H+) and oxygen gas at the anode and hydroxyl ions (OH) and hydrogen gas at the cathode. The oxygen gas produced at the anode and the hydrogen gas produced at the cathode escape out of the soil. The hydrogen ions in the anode attempt to migrate through the soil toward the cathode, whereas the hydroxyl ions in the cathode attempt to migrate through the soil toward the anode. The degree at which the H+ and OH- ions migrate depends on the buffering capacity of the soil. An acid front is produced at the anode, and at the cathode a base front is produced, and these two fronts move toward opposite directions. Page and Page (2002) have found that the acid front moves faster than the base front due to the fact that the mobility of H+ exceeds that of OH, and electroosmotic flow is generally toward the cathode. During electrokinetics the pH of the soil becomes acidic with the reading at the anode dropping to around 2 and the pH at the cathode increases to above 10. The rate of acid and base production depends on the current density (Castillo et al. 2012). Precipitated hydroxides occur at the point where the pH change occurs as the solubility of metal ions is at a minimum. In order to enhance the electrodes, reduce the pH at the cathode, and increase the pH at the anode, an alkaline solution needs to be added at the anodic compartment, and an acid solution needs to be added to the cathode compartment. According to Acar et al. (1995), the application of electric current has the following effects:

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1. It produces an acid in the anode compartment that is transported across the soil and desorbs contaminants from the surface of soil particles. 2. It initiates electromigration of species available in the pore fluid and those introduced at the electrodes. 3. It establishes an electric potential difference which may lead to electroosmosisgenerated flushing of different species. In order to remove the contaminants from the soil by electrokinetics, the contaminants should exist in pore water in dissolved ionic form so that they are transported to either the cathode or anode. In a study by Reddy and Chinthamreddy (2003), they found that in order to improve the performance of electrokinetics, the following can be done: changing the operating conditions such as switching the electrodes, prolonging the processing time, and increasing electric gradient or by controlling the reservoir fluid pH. Trivalent chromium behaves differently under electrokinetics in different soil types. In a study conducted by Reddy and Chinthamreddy (1999), they found that during the electrokinetics of both Cr(III) and Cr(VI), the pH near the anode decreased to a value between 2 and 3, and at the cathode the pH increased between 11 and 12; due to the high pH at the cathode, a precipitate of chromium hydroxide was formed that clogged the pore space of the cathodic base front. The limitation of this process is the near-anode focusing effect which results in the formation of a precipitate layer block around the anode resulting in the reduction of efficiency with time (Shen et al. 2007; Li et al. 2011). The two most common occurring valence states of chromium trivalent exist in the form of cationic hydroxides such as Cr(OH)3 which will migrate toward the cathode during electrokinetic remediation (Fig. 2.5). However, chromium(VI) exists as CrO42 at high pH and as HCrO4 at low pH and other forms of oxyanions such as CrO42 which migrate toward the anode; however it is adsorbed by the soil in the low pH regions, and it stops the complete removal of Cr (VI) from the soil. Electrokinetics is highly dependent on the acidic condition which favors the re-solubilization of heavy metal precipitated contaminants into the solution phase which makes it easier to transport; this can be done by acidification. The problem encountered during electrokinetic mobilization and recovery of metallic species in soil is mainly due to accumulation of precipitates and increase Fig. 2.5 Movement of ionic species under an electrokinetic gradient

Cathode Cr3+

H+

Anode OH–

CrO4–

Pb++

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E. M. Nkhalambayausi-Chirwa et al.

in pH at the negatively charged electrode zone (anode zone) of the field. The accumulation of divalent metal precipitates reduces the permeability of the soil around the anode, thereby increasing the electric potential required to continue the mobilization and recovery process to the point where further increase of potential is not economically feasible. The inhibition of transport of dissolved charged species due to precipitate accumulation at the anode is called the “near-anode focusing” (Li et al. 2015). Strategies such as pulse charge (Reddy and Karri 2006), charge reversal (Hunter and James 1992), and acid enhancing (Maturi and Reddy 2008) have been applied in certain scenarios in the past to increase the recovery times by decreasing or all together eliminating the near-anode focusing effect.

6.6

Biomineralization and Biocrystallization

The process of biomineralization and biocrystallization attempts to reduce the metallic species to the zero-valent state and recover the metal in its pure form. The metal can be recovered by physical means such as scrapping or used as an in situ conductor collecting electrons (Yong et al. 2002; Foulkes et al. 2016; Lloyd et al. 2008). This biotechnological innovation has shown potential in the recovery of precious metals (Mabbett et al. 2006) or in the assembly of biosynthetic electrodes in microbial fuel cells (Yong et al. 2002). In the study by Yong et al. (2002), the researchers managed for the first to demonstrate the biological reduction of palladium from palladium (II) to palladium (0) which resulted in the production of the metallic form – Pd(0) – which was detected on cell surfaces using XRD and EDX methods. The crystalline nature of the precipitate was confirmed by SEM-XRD scans where the predominant element on the cell wall deposits was shown to be palladium (0). The process of biocrystallization will definitely be of economic importance in the field of bio-purification and biorecovery of several beneficial products from the energy conservation processes.

7 Environmental Application Strategies As indicated above, the utilization of biological systems to remove toxic metals from water, soil, and environmental systems has been investigated extensively since the late 1980s. Metals can be removed through reduction to precipitable species and oxidation to precipitable species or by biosorption taking advantage of the ion-exchanging properties of cell surfaces of bacteria, fungi, and/or algae. The following are examples of processes that have been studied extensively by the research group at the University of Pretoria and our collaborators overseas:

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In Situ Bioremediation

In situ biological permeable reactive barriers (BPRBs) have been used mainly for removal of toxic organic compounds by introducing organisms or by enhancing the activity of the portion of the indigenous community possessing inherent capability to degrade recalcitrant organic compounds (Ramsburg et al. 2004; Steffan et al. 1999). Specific application of BPRB systems for the removal of Cr(VI) in groundwater has only been attempted lately at pilot level by Jeyasingh et al. (2011). The slow progress toward full implementation of biological barriers for remediation of Cr(VI) and toxic metals has been both due to the unavailability of microorganisms capable of growing under nutrient-stressed conditions and lack of information on the speciation and mobility of the reduced metal ion species in the soil. During preliminary investigations in the laboratory at the University of Pretoria, microorganisms capable of reducing and immobilizing Cr(VI) were isolated from dried sludge and introduced in columns packed with aquifer media to remove Cr (VI) from water flowing through the system (Fig. 2.6). The isolated bacteria survived under minimal nutrient conditions and were able to reduce Cr(VI) as water flowed through the inoculated microbial barrier (Molokwane et al. 2008; Molokwane and Chirwa 2009). The findings implied that autotrophic organisms utilizing hypocarbonate as carbon sources can be used in in situ remediation of groundwater without the need of supplementing the groundwater with additional nutrients. However, in spite of these promising findings, the fate of reduced chromium species and the activity of the organisms inside the columns were not investigated.

Feed Reservoirs

Hydraulic head, h

Reactor Microcosm

Makeup Feed

Makeup Feed Pump

Waste

Fig. 2.6 Experimental setup for testing microbial inoculated aquifer medium using gravity fed reactor columns packed with aquifer material. (Adapted from Molokwane and Chirwa 2009)

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In Situ Bioaugmentation

This entails identifying indigenous species of bacteria within the vicinity of the contaminated site and determining the critical carbon sources and nutrients that could be supplied to encourage the growth of the target species. When the selected nutrients are introduced into the environment, either by injection into boreholes or by spreading on the ground, the target species will out-compete other species and will be able to degrade the contaminants. The potential problem with this form of bioaugmentation is that the nutrients may be viewed as pollutants in their own right especially at the beginning of the bioaugmentation process when microbial loading is very low. The nutrients such as NO3 and SO42 have an undesirable pollution effects on receiving water bodies such as eutrophication of streams receiving the base flow from the remediated areas. However, if applied successfully, in situ bioaugmentation is the only environmentally nonintrusive process that avoids the introduction of alien species into the environment. Most countries prohibit the introduction of organisms foreign to a particular region or country (Federal Register 1999).

7.3

Ex Situ Bioremediation

Ex situ methods of remediation are conventionally used in the treatment of contaminated groundwater through a pump-and-treat process which involves the extraction of contaminated water from the aquifer, treatment above ground, and injection of the treated water back into the aquifer (Milkey 2010; Wittbrodt and Palmer 1992). The treatment above ground can be conducted via both chemical and biological means. Chemical treatment processes employed in the remediation of toxic metals produce unwanted chemical by-products which result in the production of toxic sludge (Gonzalez et al. 2003; Shakoori et al. 2000). Therefore, the chemical processes are viewed as costly and environmentally intrusive. Alternative biological treatment methods have been suggested using biotransformation followed by bioseparation using biomass (Chirwa and Wang 1997; Chrysochoou et al. 2013; Jin et al. 2008). These methods may be applied ex situ (Colica et al. 2010; Zakaria et al. 2007). The ex situ remediation processes are used rarely for treatment of soil which requires excavation of large volumes of soil, treatment in off-site facilities, and backfilling. The whole process for treatment of contaminated soil is too expensive to be undertaken on a large scale. For the treatment of the aquatic phase, experience at sites where pump-and-treat remediation groundwater contaminated with metals and radionuclides showed that, although it is feasible to remove large volumes of the contaminant from the subsurface, it becomes more and more difficult to remove the remaining pollutants as concentration decreases (EPA-Odessa 2005). The biological removal by reactors that can be applied ex situ has been demonstrated at the laboratory scale for chromium, uranium, selenium, and lead (Hawley et al. 2004; Mtimunye and Chirwa 2014; Wessels 2017; Chirwa and Wang 2001; Brink et al. 2017).

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8 New Developments in Metal Bioremediation With the increasing global population, fast-tracked industrialization, and discovery of new technologies, the volume and complexity of metal pollution is bound to increase in the future. In addition to the increasing trajectory in the generation of conventional metallic pollutants, future energy demands may result in a transitional period where nuclear energy will become predominant. This will require environmental engineers and scientist to delve deeper into research on environmentally friendly processes to counter the effects of both conventional and nonconventional pollutants such as heavy metals and recalcitrant organic compounds in the environment. Advanced microbial cultures will be sought to treat a wider variety of recalcitrant pollutants. Discussions on the possibility of genetically engineering specialized cultures for the purpose are not new in the environmental engineering fraternity. However, the application of ideas bears a large ethical burden as it is forbidden in almost all countries in the world to introduce genetically engineered organisms into the environment. Less aggressive methods for dealing with the problem without violating ethics include in situ bioaugmentation and molecular bioaugmentation to a certain extent.

8.1

Molecular Bioaugmentation

The molecular bioaugmentation process utilized genetic carriers such as transposons and plasmids to shuttle genetic information for toxic metal remediation into native species to the environment or species already adapted to the target environment. Several species of bacteria are capable of picking up and retaining circular fragments of DNA called broad-host-range plasmids which may be engineered to carry specific genes for the degradation of xenobiotic compounds and transformation of toxic metals (Weightman et al. 1984; Vincze and Bowra 2006). The same process can be applied using genetically engineered linear DNA called transposons. Although studies have been conducted using these techniques in laboratory microcosms, the application in actual environments has not been attempted (Hill et al. 1994). In the future, it is foreseeable that these methods will find wide application for the new pollutant varieties that may be untreatable by conventional methods.

8.2

Biofractionation and Bioseparation of Radioactive Elements

A very little understood application of bioseparation involves using microorganisms to discriminate radioisotopes by size. So far, this application has remained conceptual due to limited understanding on the structure and function of organisms that are

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suspected to achieve biofractionation (Molokwane and Chirwa 2007). In the latter study, Molokwane and Chirwa observed with a modest degree of certainty that microbial cells previously isolated from a high radiation-exposed facility accumulated C-14 while growing on a C-14/C-12 carbon matrix from powdered nuclear graphite. The experiment was conducted in a closed loop chemostat system equipped with biofilters for collection of suspended matter for analysis. The observed metabolic activity in the cells indicated that the process was possible under very low dissolved oxygen suggesting that microorganisms preferred inorganic forms of carbon as the primary carbon source. Bacteria that utilize inorganic carbon sources such as CO2 and HCO3- as primary carbon sources – known as autotrophic organisms – favor anaerobic conditions for growth. However, in this preliminary study, the amount of C-14 remaining in solution was not measured which could be required to draw a mass balance on C-14 in the system. These preliminary results on C-12/C14 bioseparation hold promise for development of the decontamination and reuse process for graphite in model HTGR nuclear reactors that produce large amounts of radioactive nuclear graphite from expired fuel containment (pebbles). Success in the above process is also important for the decontamination and recovery of nuclear graphite decommissioned plants for reuse in new reactors.

9 Conclusion Since the first Cr(VI)-reducing bacteria were isolated in the 1970s and U(VI)reducing bacteria were isolated in the late 1980s, a lot of progress has been made in isolating and developing higher performing cultures adapted to various environments. New research using genetic tools has yielded new cultures and new understanding of the metal removal processes both at the molecular level [through genetic studies] and at culture community level [through genomics and proteomics]. Pure and mixed cultures of bacteria have been applied successfully in treating industrial effluents containing Cr(VI), U(VI), As(III), and many other metals. However, application of biological systems in the remediation of contaminated environments still faces a challenge. Although culture performance under natural conditions has been evaluated using laboratory microcosms, more research is still required to elucidate the fate and possibility of recovery of artificial microbial barriers. The question of the fate of reduced species and what to do about the foreseeable blockage by metal hydroxides, metal sulfides, and other complexing agents remains unanswered. In order for the in situ bioremediation technology to work for Cr(VI), U(VI), Se(VI), Tc(VII), and other toxic heavy metals, a solution must be found for feasible recovery of the barrier zones involving remobilization of reduced Cr species. So far, in the current studies, this was achieved by applying an electric potential across a section of contaminated soil. More work is required in the above process to prevent localized precipitation of metals around the electrodes which eventually stopped the migration of metals from the aquatic phase.

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Acknowledgment The research was funded by the National Research Foundation (NRF) of South Africa through the Incentive Funding for Rated Researchers Grant No. IFR2010042900080 and Competitive Programme for Rated Researchers Grant No. CPR2011060300001 awarded to Prof. Evans M.N. Chirwa of the University of Pretoria. Postdoctoral fellow working on this project – Dr Zainab Birungi – was funded by the National Research Foundation and the University of Pretoria Postdoctoral Fellowship Programme.

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Chapter 3

Phycoremediation: Algae as Eco-friendly Tools for the Removal of Heavy Metals from Wastewaters Shamshad Ahmad, Arya Pandey, Vinayak Vandan Pathak, Vineet Veer Tyagi, and Richa Kothari

Abstract Phycoremediation is a potential tool to eradicate the excess toxics (heavy metal and organic contaminants) from the industrial waste stream. The algal species are a promising, eco-friendly, and sustainable move toward a possible advantage to enhance the algal cultivation which in turn magnifies the economics of algal-based value-added products. Therefore, algae have been documented as a sustainable and inexpensive vector for detoxification of noxious waste-loaded industrial waste stream. Algal species may bind up to 10% of their biomass as metals. Various physical and chemical methods used for this purpose suffer from serious limitations like high cost, high energy input, alteration of basic properties, and disturbance in native flora. In contrast, phycoremediation provides a new insight/dimension for this problem by perceiving it as cost-effective, efficient, novel, eco-friendly, and solardriven technology with good public acceptance. The mechanism for the removal of heavy metal through alga works on the principle of adsorption onto the cell surface which is independent of cell metabolism and absorption or intracellular uptake which depends on cell metabolism. So, their ability to adsorb and metabolize is associated with their large surface/volume ratios; the presence of high-affinity, metal-binding groups on their cell surfaces; and efficient metal uptake and storage systems. Hence, the present review article deals with the basic mechanism of S. Ahmad · A. Pandey Bioenergy and Wastewater Treatment Laboratory, Department of Environmental Sciences, Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India V. V. Pathak Department of Chemistry, Manav Rachna University, Faridabad, Haryana, India V. V. Tyagi Department of Energy Management, Shri Mata Vaishno Devi University, Katra, Jammu and Kashmir, India R. Kothari (*) Department of Environmental Sciences, Central University of Jammu, RahyaSuchani (Bagla), Samba, Jammu, Jammu and Kashmir, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_3

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algal-based heavy metal removal strategies with the effect of physicochemical parameters. Use of transgenic approaches to further enhance the heavy metal specificity and binding capacity of algae with the objective of using these algae for the treatment of heavy metal-contaminated wastewater is also focused in this article. Keywords Phycoremediation · Heavy metals · Passive and active algal biomass · Mechanism

1 Introduction Heavy metals (HMs) are defined as metallic element with higher density and have toxicity at lower concentration. Heavy metals belong to the group of metal or metalloids with atomic density higher than 4 g cm
3 or five times or more greater than water (Kumar et al. 2015). However, in biochemistry HMs are defined as metallic elements with Lewis acid behavior, i.e., electron pair acceptor. About 53 chemical elements are considered as heavy metal, and most of these HMs are found as natural constituent of the earth crust and soil. From environmental perspective, any metals or metalloid that poses potential harmful effects to the living organism even at lower concentration can be termed as heavy metals. Most of these heavy metals (Zn, Cu, Mn, Ni, and Co) have vital function in plants, while other metals such as Cd, Pb, Hg, and Cr are found to cause toxic effects in biological system. The combined influence of urbanization, industrialization, and chemical consumption in agrarian practices has raised the heavy metal concentration up to the toxic level; hence its remediation is now considered as global concern. HM contamination in water is one of the most critical environmental problems, which include Cd, Cr, Cu, Hg, and Zn as common contaminants (Pathak et al. 2019; Ahmad et al. 2018; Kothari et al. 2012). Heavy metal contamination in wastewater depends on industrial processing and is stated an as anthropogenic activity. HMs are classified as conservative pollutants, and their accumulation in the environment causes various negative impacts such as inhibition of photosynthesis and seed germination, decreased enzymatic activity, reduction of chlorophyll production, etc. In order to prevent the negative impacts of heavy metals, adequate treatment of wastewater is desired prior to its disposal or discharge in receiving water bodies. Chemical and biological treatment methods are available for HM removal, but biological methods are preferred because of limitation and drawbacks in chemical treatments. Bioremediation is a key process that utilizes microbes to tackle heavy metal pollution (Pathak et al. 2015). Biological processes to remove heavy metals are well explored by various researchers as a part of phycoremediation. However, remediation of heavy metals via algae gained substantial attention due to its effectiveness and feasibility in implementation. Algae offer potential solution for treatment of industrial wastewater containing heavy metals in a natural way. Algal-based remediation can be termed as phycoremediation, which not only resolves the challenges associated with conventional treatment methods but is

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also considered as an economically viable and environment-friendly treatment option (Ungureanu et al. 2015). Phycoremediation involves the natural ability of alga to uptake the nutrient, accumulate the heavy metals, and degrade the organic contaminant via symbiotic interaction with aerobic bacteria. Algae resemble the pigment of higher plants with higher photosynthetic efficiency; hence algae released greater extent of oxygen in aquatic system and induce the aerobic degradation of organic compounds (Majumder et al. 2015). Alga is found to have the ability to utilize waste as nutritional source, and it reduces the pollutants through metabolic and enzymatic processes. The xenobiotics and heavy metal pollution can be detoxified, transformed, and volatilized through the algal metabolic pathways (Gautam et al. 2015). Therefore, biological method employing algae has various advantages such as (i) minimum capital and operating cost compared to physicochemical/ oxidation process (Mane et al. 2011), (ii) true destruction of organic and inorganic pollutants (Parameswari et al. 2010), (iii) oxidation of wide range of organic compounds, (iv) removal of reduced inorganic compound, i.e., sulfides and ammonia (Praepilas and Pakawadee 2011), etc. Biosorption is the dominant mechanism in uptake of heavy metals either by active algal biomass (AAB) or passive algal biomass (PAB) and found as a cost-effective solution to eliminate HMs from industrial effluent. In case of passive algal biomass, biosorption doesn’t involve in metabolic pathway; however, it entirely depends on interaction between the biomass and metal ion; hence it resembles with the binding of metal ions through ion-exchange resins. Contrary to the ion-exchange resins, biosorption involves various steps such as chelation, partial adsorption, complexation, micro-precipitation, etc. On the other hand, biosorption in active algal biomass is carried out through energy-mediated transport of metal ions through the cell membrane. The ability of metal sorption through various organisms has been widely reviewed by researchers and concluded that PAB have massive potential to bind metal ions from very low concentration in the external solution. It has also been reported that biosorption is significant to remove toxic metal and at the same time, recovery of valuable metals such as gold, silver, and radionuclides is also possible. Regardless of various advantages with algal-based metal uptake, several drawbacks (low biomass generation, cost-effective biomass production less effective to remediate various industrial wastewaters) are also associated with phycoremediation. Therefore, phycoremediation coupled with effective cultivation system (solar-driven open pond or close photobioreactor) requires extensive investigation (Kothari et al. 2017; Ahmad et al. 2017). The present review deals with the various aspects of HM removal as well as factors affecting the metal removal efficiency.

2 Hazardous Effects of Heavy Metals Heavy metals are considered to have hazardous effect on the flora and fauna. Key source of heavy metals involving anthropogenic activities such as extraction, excavation, etc. is depicted in Fig. 3.1. Nonpoint source pollution such as haphazard

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Fig. 3.1 Source of HMs in the environment Table 3.1 Maximum contaminant level (MCL) standards for the most hazardous HMs HMs Arsenic (As) Cadmium (Cd) Chromium (Cr) Cupper (Cu) Nickel (Ni) Zinc (Zn) Lead (Pb) Mercury (Hg)

Toxicities Skin infection, visceral organ cancers, vascular disease

Unit (mg/L) 5  10
2

Renal failure, kidney dysfunction

10
2

Headache, diarrhea, nausea, vomiting, carcinogenic in nature

5  10
2

Hepatic damage, Wilson disease, insomnia

2.5  10
2

Skin infection, chronic asthma, carcinogenic in human Neurological disorder, loss of hunger Renal disease, circulatory and nervous system infection Rheumatoid arthritis and diseases of the kidneys, circulatory system, and nervous system

0.2 0.8 6  10
3 3  10
5

Adapted from Babel and Kurniawan 2003

disposal and release of industrial and domestic wastewater in aquatic ecosystem, with even trace concentration, threatens the aquatic flora and fauna. Even trace concentrations of heavy metal pose significant problems to flora and fauna given in Table 3.1. Heavy metals don’t participate in the metabolism of the body, but it is accumulated through different mechanisms which include bioaccumulation, bioconcentration, and biomagnifications illustrated in Fig. 3.2. Organisms exposed in heavy metals tend to have protective mechanism against heavy metal toxicity. The HM concentration beyond a threshold limit causes direct toxicity to the aquatic flora and fauna.

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Fig. 3.2 Effect of HMs on living organism

2.1

Biosorption of HM Ions Using Algae

Biosorption process involves sorption of material in contact via biopolymer or biomaterial. It is found effective in detoxifying heavy metals in lower concentration even with less biomass supplementation with no additional nutrient requirement. Presence of organic ligands or the functional groups (carboxyl, hydroxyl, sulfate, phosphate, and amine group) in structural components of algal cell makes it as a potential biosorbent. Moreover, studies have shown that inactive biomass may be even more effective than active (living) algal cell for removal of heavy metals (Gautam et al. 2015). Active algal biomass-based heavy metal removal is based on the efficacy of algal growth in heavy metal containing aqueous solution, which may pose toxic effects to the algal cells resulting in variation in heavy metal removal capacity. Heavy metal uptake by active algal biomass is more complicated than the inactive biomass as metals are absorbed and involved in intracellular pathway of living algal cells (Misbah et al. 2014). In contrast, PAB cells adsorb HM ion on the surface of the cell wall. PAB can be observed as an aggregation of polymers (carbohydrates, cellulose, pectin, glycoprotein, etc.) that is capable of binding with HMs as adsorbents with the efficient and cost-effective wastewater treatment.

2.2

Cellular Sites Involved in HM Binding

HM ions bind to the AAB and PAB cell surface and are also transported within the cell, whereas the adsorptions process does not depend on metabolic process, requiring several metal transporters (Barakat 2011). Several AAB have metal efflux metabolism-driven systems for maintaining the HM concentration in intracellular

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Fig. 3.3 Cellular mechanism of HMs: (a) inside the living cell; (b) on the cell surface

space avoiding HM toxicity. Inside the cell HM ion can be distributed in cell vacuoles and organelle given in Fig. 3.3. Several cell-derived biomolecules (polyphosphates, phytochelatins, metallothioneins, metalloproteins, etc.) help in the sequestration of HMs from the wastewater. Besides this, the enzymatic reaction

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Table 3.2 Functional groups involved in metal ion biosorption and adsorption Binding group Hydroxyl Sulfhydryl (thiol) Sulfonate Amine Imine Imidazole Phosphodiester

Functional group -OH -SH O¼S¼O -NH2 -NH -C-NH > CH >P¼O-OH

Ligand atom O S O N N O O

Occurrence Polysaccharides, uric acid, amino acid Amino acid Sulfated Amino acid Amino acid Amino acid Teichoic acid, lipo PS

Chowdhury et al. 2015; Zhang et al. 2015; He and Chen 2014

can alter the oxidation number of HMs and change into less toxic forms. Microprecipitation of HM removal in the form of phosphates and sulfates by AAB is a potential approach to remove HMs from wastewater (Ungureanu et al. 2015). In case of PAB, drying and crushing should increase the metal-adsorbing capacity. The algal cell wall made up of microfibrillar exo-polysaccharides has typical chemical composition and contains functional groups as shown in Table 3.2 such as –COOH
, -OH
, -PO4
3, -RSH, SO4
2, etc. These functional groups produce anionic nature to the cell wall and microfibrils. Since HM ion in wastewater is cationic in nature, they are adsorbed by the cell and microfibril surface. Cyanobacterial cell wall consists of mainly peptidoglycan, polymer of N-acetylglucosamine and β-1,4-N-acetylmuramic acid, which provides mostly –COOH functional group for HM adsorption. Few cyanobacterial cell walls bear capsule wall which is anionic in nature due to acid in nature and thus help in metal adsorption. Eukaryotic algae cell wall contains heteropolysaccharides, which offer -COOH and SO4
2 groups for HM adsorption (David et al. 2012). Plasma membrane and membranes of organelles, consisting of lipopolysaccharides and lipoprotein, contribute significantly to metal sorption by algae and cyanobacteria.

2.3

Ion-Exchange Concept

Microalgal cell wall plays significant role in an ion exchange as it is composed of polysaccharides, proteins, and lipids. These constituents contribute to various functional groups (carboxyl, hydroxyl, phosphate, amino, sulfhydryl, amide, alkyl, and aromatic compound) and hence possess an overall negative charge to the cell surface. The cell surface of alga acts as a strong binding site for metal cations and is involved in metal exchange through the ion-exchange mechanism (Monteiro et al. 2011). During the interaction between metal ion and protein on biological surface, metal ions coordinated in formation of complex groups. However, in marine system, a major part of active sites are bonded with protons at low pH or with alkaline earth metals (Ca, Na, and Mg) at higher pH. In the presence of cations such as Cu+2, Mn+2, Zn+2, Ni+2, Cd+2, Fe+3, and Pb+2, the previously bind protons and metals are

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Fig. 3.4 Basic principle of ion exchange through algae

released, and these cations are sorbed on cell surface. But in the case of anions, adsorption characteristics of algal significantly change toward the competitive binding of metal ions to the cell surface as shown in Fig. 3.4.

2.4

Physical Adsorption

Physical adsorption refers to a phenomenon in which aqueous metal ion binds to polyelectrolytes of algal cell wall through weak force of attraction such as van der Waals force, covalent bonding, redox interaction, biomineralization, etc. (Perpetuo et al. 2011). The pH of the adsorbing media has strong influence on the adsorption of the metal ion. It has been found that alkaline pH increases the attraction of metal cations and thus improves their adsorption on cell surface by replacing the functional groups containing negative charge such as polysaccharides, phosphate, amino group of nucleic acid, and amino and carboxyl group of protein (Majumder et al. 2015). Furthermore, electrostatic attraction has been found as the main mechanism for adsorption of metals such as uranium, cadmium, zinc, copper, and cobalt through passive algal biomass (Fig. 3.5).

3 Factor Affecting Uptake of HM Ions Biosorption has been regarded as the main mechanism for metal ion removal via algal biomass, but various factors are found to have influence on biosorption potential of alga such as pH, contact time, and temperature. In this context, the following section provides a detailed perspective on influence of these variables on biosorption of heavy metals (Ahalya et al. 2003).

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HM+ Cell surface adsorption +

HM

Electrostatic Interactions HM

+

HM+

HM+

Vacuole

Efflux

HM+ Bioaccumulation

Intracellular Ligands HM+

HM+ Ion exchange HM+ Cation exchange

HM+

Surface C complexation SO42RCOOR2OSO3–NH2 –SH

Precipitation

Fig. 3.5 Mechanism of physical adsorption of heavy metals. (Ayansina and Olubukola 2017)

3.1

pH

The pH of the aqueous solution (wastewater) is one the most imperative factors that directly influences the biosorption process to remove heavy metal from wastewater. pH affects the dissociation of functional groups of the active sites of biosorbent as well as chemistry of ionic solution. pH optima for biosorption via algae vary from metal to other metals. Ajayan et al. (2011) reported a significant decrease in pH ranging from 5.6 to 8.3 while removing heavy metal from tannery wastewater. Ritixa and Monika (2013) reported that pH optima for iron and copper were 8 for both with removal efficiency of 92–93%, respectively, in biosorption process. Dominic et al. (2009) reported that pH level of industrially polluted wastewater shows a drift from acid to alkaline, i.e., 6.0–8.1, after treatment with Chlorella vulgaris. Wastewater treatment with Synechocystis salina shows a slight drift in pH from 6.0 to 8.0, while the same wastewater shows variation in pH decrease (from 6.0 to 7.9) treated with different algal species Gloeocapsa gelatinosa. Therefore, it is clear from the above explanation that changes in pH have substantial potential to alter the biosorption potential through various processes such as affecting ionic chemistry and metal availability in medium and affecting algal growth in case of active algal biomass.

3.2

Temperature

The biosorption of heavy metals through algae is unaffected within the temperature ranging from 20 to 35  C, while at 40–50  C, biosorption efficiency increased, but such high temperatures may be responsible for permanent structural damage to the

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Table 3.3 Metal uptake through various passive algal biomasses at different temperatures Algal species Cystoseira barbata Cystoseira barbata Cystoseira barbata Lessonia nigrescens Sargassum muticum Spirogyra sp.

Metal Cd+2

Temp ( C) 20

Initial HMs conc. (mg/L) 117

Metal uptake (mg/L) 37

% Removal ~68

Ni+

20

224

78

~65

Pb+2

20

414

196

~52

As+

20

200

45

~77

Sb+2

23

10

5

50

Pb+2

25

200

140

30

References Yalçın et al. (2012) Yalçın et al. (2012) Yalçın et al. (2012) Hansen et al. (2006) Ungureanu et al. (2015) Gupta and Rastogi (2008a, b)

algal cells. As a consequence, it decreases in metal uptake. Biosorption is mainly based on adsorption reaction which is an exothermic process. The extent of adsorption of heavy metals through algae increases with decrease in temperature. It has been reported that the temperature optimum for S. cerevisiae was 25  C for maximum heavy metal (Ni and Pb) biosorption. Ali et al. (2013) reported that metal uptake by S. platensis (PAB) increased gradually with increasing temperature, and it was found that metal (Cu) uptake was maximum (90.61%) at temperature 37  C. Hence, temperature plays a vital effect on metal uptake through algal biomass as shown in Table 3.3.

3.3

Contact Time

PAB adsorb passively HMs on the surface of cell wall rapidly within few minutes, while in living algal biomass metal sorption is a gradual process and follows the life cycle of alga (Vogel et al. 2010). Tuzen and Sari (2010) observed that PAB Chlamydomonas reinhardtii biomass adsorbs Hg+2, Cd+2, and Pb+2 and equilibrium is achieved within 60 min. According to Mata et al. (2009), PAB biomass of Fucus vesiculosus (macroalgae) removes Au+3 28.9 mg/g and 74.1 mg/g after 1 and 8 h; this process suggest that biosorption of HM ion is a passive process that occurs relatively on a rapid scale. But, in AAB the biosorption rate of Cd+2 by Cladophora fracta decreased by increasing time (Wang et al. 2010), but greater absorption capacity is found in old culture (Ozer et al. 2000). The issue with older culture gradual depletion of cell surface by nutrient; this will affect the biosorption capacity of HMs on the algal cell surface.

3 Phycoremediation: Algae as Eco-friendly Tools for the Removal of Heavy. . .

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Thus, pH, contact time, and temperature determine the sorption capacity of heavy metal ions. Researchers have optimized these variables and achieved maximum biosorption potential of various algal strains, which are shown in Table 3.4. Temperature ranges (20–96  C) with contact time (30–240 min) with optimal pH ranges were experimentally investigated by various researchers in the last 15 years, using different algae.

4 Algal Biomass-Based Remediation Approaches for Heavy Metals: Traditional vs. Advanced Conventional practices for metal ion removal were dominated by chemical methods such as chemical-mediated precipitation, redox reactions, ion-exchange resins, organic polymers (starch, poly ions, and xanthate), coagulation, osmosis, chemical-induced extraction, adsorption via activated carbon, electroprecipitation, and electrodialysis (Lezcano et al. 2010). But these methods are found to be costly and less effective for HM removal (Plaza et al. 2013). Contrary to these methods, application of inorganic adsorbents such as clay, mud, ash, alum, and other organic adsorbents (waste biomass, agricultural residues, plant leaf, etc.) was found to be less expensive, but most of them resulted in incomplete remediation (Zhang et al. 2016; Lee et al. 2016). These conventional methods demand large amount of energy and chemicals (Majumder et al. 2015). So, there have been developed, formulated, modern, economical, and sustainable adsorbents for the removal of HMs and toxic substances from wastewaters. Most of the researchers have been escalating their efforts in developing suitable adsorbents for the complete removal of HMs. Generally, algal-based HM remediation is considered as a part of bioremediation and involves biosorption either by passive sorption of pollutant independent of metabolic process or active sorption of pollutant depending on metabolic pathway. In case of active sorption process, energy generated by respiration is consumed in metal sorption; hence this process depends on the efficacy of physiological process of the living algal biomass. In addition to this, environmental variables such as pH, temperature, contact time, etc., nature of ionic species, biomass concentration, contact time, and nature of the adsorbent also affect the biosorption capacity.

4.1

Microalgae Potential in HM Remediation

Microalgae belong to the group of photosynthetic organisms and are found in freshas well as marine water environment. These organisms have tremendous photosynthetic efficiency, and about 32% of the global photosynthesis is carried out by microalgae (Priyadarshani et al. 2011). Microalgae perform the specific mechanism to uptake the essential heavy metals required to their cell growth. The benefits of

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Table 3.4 Factor influencing the HM removal efficiency by algal biomass

References Aksu (2001) and Aksu and D€onmez (2006) Mehta and Gaur (2001) Mehta et al. (2002) Mehta et al. (2002) Vogel et al. (2010) Zeroual et al. (2003) El-Sikaily et al. (2007) Lee et al. (2004) Lee et al. (2004)

Cd+2

Akhtar et al. (2004) Rincón et al. (2005) SrinivasaRao et al. (2005) Arıca et al. (2005) Tüzün et al. (2005) Mata et al. (2009)

Ni+2 Ni+2 Cu+2

Wong et al. (2000)

Ni+2

Han et al. (2006) Romera et al. (2007)

Cr+3 Cd+2

Temp ( C) 20

Time (minute)

89.19

3.5

25

30

“ “ “ U. lactuca “

420.67 714.89 26.6 149.25 10.61

3.5 3.5 4.4 7 1

25 25 96 25 25

180 120 – 60 120

Laminaria japonica Laminaria japonica C. sorokiniana F. vesiculosus Sphaeroplea sp. C. reinhardtii “ Fucus vesiculosus Chlorella miniata “ Ascophyllum nodosum “ Asparagopsis armata “ C. sorokiniana Rhizoclonium heiroglyphicum Ascophyllum nodosum Nostoc linckia Spirogyra sp. Cystoseira barbata Sargassum bebanom

75.27

4.5

30

60

136.1

4.5

30

120

48.08 46.95 140.43

7.4 6 4

25 – 33

– 75 105

25.6 72.2 74.05

2 6 7

25 25 23

120 120 120

1.36

5

24

32

41.12 87.7

4.5 6

25 –

– 120

58.8 21.3

4 5

– –

120 30

32.3 58.8 11.81

6 4 4

– 25 –

120 120 120

–

–

96

–

41.1 265 37.6

3.5 4 4

25 30 20

– 120 60

14.24

3

25

110

Algal species Chlorella vulgaris

Cu+2

“

Cu+2 Cu+2 U+4 Hg+2 Cr+4 Al+3

Cu+2 Cd+2

Akhtar et al. (2008) Onyancha et al. (2008) Vogel et al. (2010)

Optimal pH 4

Metal Cd+2

Cr+4 Hg+2 Au+3

Cu+2 Cr+3 Cr+3 Zn+2

Mona et al. (2011) Yaqub et al. (2012) Yalçın et al. (2012)

CO+2 Cr+3 Cd

Javadian et al. (2013)

Cr+6

Max. sorption (mg g
1)

(continued)

3 Phycoremediation: Algae as Eco-friendly Tools for the Removal of Heavy. . .

65

Table 3.4 (continued)

References Hammud et al. (2014) Ungureanu et al. (2015) Sargin et al. (2016)

Metal Pb Sb Cd+2

Algal species Enteromorpha Sargassum muticum Cladophora sp.

Cr+3

“

Cu+2

“

Max. sorption (mg g
1) 83.3 5.5 0.240 mmol/ gram 1.128 mmol/ gram 1.059 mmol/ gram

Optimal pH 3 5

Temp ( C) – 23

Time (minute) – 240

–

25

–

–

25

–

–

25

–

microalgae include rapid capacity of metal uptake, reduced time and energyefficient, eco-friendly, polynomial, recyclable, economical, highly efficient, large surface/volume (S/V) ratio, high selectivity (which enhances their performance), no synthesis required, and useful in all types of system (Cristina et al. 2012). Apart from possessing greater HM ion removal efficiency, microalgae perform easy recovery of HMs involving a few simple desorption physical and chemical methods. AAB requires minimum nutrients (nitrogen and phosphorus) and climatic condition, while PAB does not require nutrients. Moreover, they could also remove HM ion from wastewater and aqueous solutions too. Microalgae can effectively remove HMs, and use of transgenic approaches enhances the HM binding efficiency. Microalgae have the ability to bind polyvalent metal ions; thus they can be effectively applied to treat the wastewater contaminated with polyvalent metallic ion. With affinities for polyvalent metals helping to establish their potential application in cleansing of wastewater containing dissolved metallic ions (De-Bashan and Bashan 2010), particularly, Chlorella and Scenedesmus are microalgae of choice for metal removal. Passive algal biomass has been found to uptake a variety of heavy metals such as Fe, Co, Cu, Mn, Ni, V, Zn, As, Cd, Mo, Pb, and Se. Brinza et al. (2007) explored the potential use of PAB of Chlamydomonas reinhardtii, C. sorokiniana, C. vulgaris, C. miniata, Chlorella salina, Chlorococcum spp., Phaeodactylum tricornutum, Scenedesmus abundans, S. quadricauda, S. subspicatus, Spirulina platensis, (Gokhale et al. 2008) and Spirogyra sp. for biosorption of heavy metal ions. Microalgal biomass specially produces peptide bond during the photosynthesis which is capable of binding HM ion and forming organometallic complexes, which are further reached inside the vacuoles to maintain the cytoplasmic concentration of HM ion, which neutralizes the toxic effect of the HMs. Most studies of HMs focus on Cu, followed by Cd, Ni, Pb, Zn, Hg, and Cr by microalgae. The efficiency to absorb metal was found to be different in macro- and microalgae strains. Metal uptake capacity in macroalgae is found to be directly related to the extent of alginate, its availability to provide sorption site, and specific macromolecular conformations. Despite having similar functional groups in Spirogyra and Cladophora sp., Lee and

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Chang (2011) found that Spirogyra has higher adsorption capacity for Pb(II) and Cu (II) than Cladophora sp. Hence, capacity for bio-removal of metal in macro- and microalgae, clear differences have been observed in accumulation.

4.2

Active Algal Biomass vs. Passive Algal Biomass

The difference of metal biosorption via AAB and PAB has been clearly explained in the above sections. However, ion exchange is the common process found in both biosorbents which has a large contribution in biosorption potential. Most of the researchers preferred PAB for biosorption process due to its possibility to recycle and reuse. In addition to this, PAB don’t require additional nutrient source, and its application to remove heavy metal ions under extreme environmental conditions was found feasible in comparison to the living algal biomass. PAB biomass can be pretreated by physical and chemical methods to improve biosorption efficiency, while in living algal cell, sorption potential is limited and depends on growth capacity of algal strain. Acidic and alkaline condition of growth medium can affect the growth rate of algae and causes metal ion precipitation, respectively. Table 3.5 summarizes the efforts of various researchers toward the algal-based metal biosorption.

4.3

Immobilized Algae

Initially, immobilization of algal cells was proposed to deal with the challenges associated with harvesting and dewatering of algal cells. However, immobilization offers several advantages over the algal cells grown in free suspension such as (1) immobilized algal cells occupy less surface area; (2) immobilized algal cell has been found with increased photosynthetic activity, biosorption capacity, and bioactivity; and (3) immobilized algal cells are found to be resistant to harsh environmental condition and less exposure to toxicity. Immobilization increases the applicability of entrapped algal cells for repetitive biosorption process (Eroglu et al. 2015). Researchers have developed various techniques (adsorption on surface, flocculation, liquid-liquid emulsion, covalent coupling, etc.) for entrapment of algal cells, but application of synthetic polymer (poly acryl amide) or natural polymers (agar, cellulose, and alginate) is the most preferred technique. Immobilization of algal cell using polysaccharide gels has often been used for the purpose of wastewater treatment (nutrient uptake and metal ion removal). Entrapment of algal cells in alginate has been found to have sufficient immobilization and improved removal efficiencies from aqueous environment (Table 3.6). Maznah et al. (2012) reported higher biosorption capacity (Cu, 33.4 mg/g; Zn, 28.5 mg/g) in algal cells (Chlorella sp.) immobilized by sodium alginate than that of the free biomass. Recently, it has been found that incorporation of polyethyleneimine in alginate algal

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Table 3.5 HMs removal efficiency using active algal biomass and passive algal biomass Ionic Metal state Algae biomass Active algal biomass Cd Cd2+ Chaetoceros calcitrans Desmodesmus pleiomorphus Isochrysis galbana Planothidium lanceolatum Scenedesmus abundans Synechocystis sp. Tetraselmis chuii

Cr Cu

Cr3+ Cr6+ Cu2+

pH

Maximum uptake capacity (mg/g)

8

1055.27

4

61.2

7

0.02 275.51

8

574 199.83

8

13.46

References Sjahrul and Arifin (2012) Monteiro et al. (2010) Sbihi et al. (2012) Sbihi et al. (2012) Monteiro et al. (2009) Tiantian et al. (2011) Sjahrul and Arifin (2012) da Costa and de Franca (1998) Doshi et al. (2007) Doshi et al. (2007) Tien et al. (2005) Tien et al. (2005) Tien et al. (2005) Tien et al. (2005) Tien et al. (2005)

Tetraselmis chuii

292.6

Spirulina sp. Spirulina sp. Anabaena cylindrica Anabaena spiroides Asterionella formosa Aulacoseira varians Ceratium hirundinella Chlorella fusca

4–5 4–5 4–5 4–5 4–5

304 333 12.62 8.73 1.1 2.29 2.3

6

3.2

7

2.4

Dönmez et al. (1999) Yan and Pan (2002)

220 0.5 3.96 0.11 134.32

Doshi et al. (2006) Yan and Pan (2002) Tien et al. (2005) Sbihi et al. (2012) Sbihi et al. (2012)

3–7 5 7

7.74 14.57

6

42.6

Sandau et al. (1996) Singh et al. (2007) Inthorna et al. (2002) Tüzün et al. (2005)

6

22.24 58.6

Chlorella pyrenoidosa Chlorella sp. Closterium lunula Eudorina elegans Isochrysis galbana Planothidium lanceolatum Passive algal biomass Cd Cd2+ AER Chlorella Aulosira fertilissima Calothrix parietina Chlamydomonas reinhardtii Cyclotella cryptica Desmodesmus pleiomorphus

7 4–5 7

Schmitt et al. (2001) Monteiro et al. (2011) (continued)

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Table 3.5 (continued) Metal

Co

Cr

Ionic state

Co+

Cr3+

Cr 6+

Cu

Cr+3 Cu 2+

Fe

Fe3+

Hg

Hg2+

Algae biomass Hydrodictyon reticulatum Pithophora odeogonia Spirulina platensis TISTR 8217 Spirulina sp.

pH 5

Maximum uptake capacity (mg/g) 7.2

References Singh et al. (2007)

5

13.07

Singh et al. (2007)

7

98.04

7.5

0.46

Tolypothrix tenuis TISRT 8063 Oscillatoria angustissima Spirogyra hyalina

7

90

4

15.32

Spirulina sp.

7.5

0.01

Chlorella miniata Chlorella sp. Spirulina sp. Chlamydomonas angulosa Chlamydomonas reinhardtii Nostoc muscorum

3

8.2

14.17 98 167 5.32

2

18.2

Rangsayatorn et al. (2004) Chojnacka et al. (2005) Inthorna et al. (2002) Mehta and Gaur (2005) Kumar and Oommen (2012) Chojnack et al. (2005) Han et al. (2006) Doshi et al. (2008) Doshi et al. (2007) Dwivedi et al. (2010) Arıc et al. (2005)

3

22.92

Phormidium bohneri

8.2

8.55

4–5 4–5 5 5

104 0.53 3.03 21.77 8.72

Gupta and Rastogi (2008a, b) Dwivedi et al. (2010) Doshi et al. (2008) Tien et al. (2005) Tien et al. (2005) Singh et al. (2007) Singh et al. (2007)

4–5

2.47

Tien et al. (2005)

6

1.67

Schmitt et al. (2001)

2 9.2 6

24.52 0.03 72.2

Romera et al. (2006) Singh et al. (1998) Tüzün et al. (2005)

7

18

4

11.92

Inthorna et al. (2002) Schmitt et al. (2001)

Chlorella sp. Asterionella formosa Aulacoseira varians Aulosira fertilissima Hydrodictyon reticulatum Microcystis aeruginosa Phaeodactylum tricornutum Chlorella vulgaris Microcystis sp. Chlamydomonas reinhardtii Chlorella vulgaris BCC 15 Cyclotella cryptica

12.82

(continued)

3 Phycoremediation: Algae as Eco-friendly Tools for the Removal of Heavy. . .

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Table 3.5 (continued) Metal

Ni

Pb

Zn

Ionic state

Ni2+

Pb+2

Zn2+

Algae biomass Phaeodactylum tricornutum Porphyridium purpureum Scenedesmus subspicatus Arthrospira (Spirulina) platensis Aulosira fertilissima Chlorella sp. Hydrodictyon reticulatum Pithophora odeogonia Spirogyra neglecta Arthrospira platensis Aulosira fertilissima Chlamydomonas reinhardtii Cyclotella cryptica Hydrodictyon reticulatum Arthrospira platensis Aulosira fertilissima Desmodesmus pleiomorphus Hydrodictyon reticulatum Phaeodactylum tricornutum Pithophora odeogonia

pH 4

Maximum uptake capacity (mg/g) 0.51

References Schmitt et al. (2001)

4

0.51

Schmitt et al. (2001)

4

9.2

Schmitt et al. (2001)

5 5.5

20.78

Ferreira et al. (2011)

5 5

4.16 183 13.86

Singh et al. (2007) Doshi et al. (2008) Singh et al. (2007)

5

11.81

Singh et al. (2007)

5 5–5.5 5 5

26.3 102.56 31.12 96.3

Singh et al. (2007) Ferreira et al. (2011) Singh et al. (2007) Tüzün et al. (2005)

6 5

36.68 24

Schmitt et al. (2001) Singh et al. (2007)

5–5.5 5 5

33.21 19.15 360.2

5

3.7

Ferreira et al. (2011) Singh et al. (2007) Monteiro et al. (2009) Singh et al. (2007)

6

14.52

Schmitt et al. (2001)

5

8.98

Singh et al. (2007)

beads substantially increases the sorption capacity for Pt (II) and Pd (IV) (Wang et al. 2017). In addition to high capacity for metal biosorption, Lopez et al. (2017) reported that algal cells encapsulated in alginate yielded more nucleic acid than the free-living cells; hence higher concentration of nucleic acid indicates the more active cells in alginate algal beads than the free-living cells. Limitations were also observed in cell entrapment such as poor diffusion of carbon dioxide and oxygen, larger volume-to-surface ratio of encapsulating materials, etc. However, to overcome these challenges, researchers (Wang et al. 2017; Mujtaba and Lee 2017) have made several attempts by introducing new entrapment matrices.

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Table 3.6 HM removal efficiency by immobilized algal biomass Metal species Cd2+

Cu2+

Hg2+

Ni2+ Pb2+

Zn2+

4.4

pH 6

Immobilized –

Maximum metal uptake (mg/g) 79.7

5

–

192

6

Alginate

70.92

4–7

Silica

36.63

7

–

33.4

Chlorella vulgaris

4.5

–

63.08

Chlamydomonas reinhardtii Chlamydomonas reinhardtii Chlorella vulgaris

6

Ca-alginate

35.9

6

Ca-alginate

106.6

4.5

–

111.41

6

Ca-alginate

230.5

6

Ca-alginate

380.7

7

–

28.5

Algae Chlamydomonas reinhardtii Chlorella sorokiniana Spirulina platensis TISTR 8217 Spirulina platensis TISTR 8217 Chlorella sp.

Chlamydomonas reinhardtii Chlamydomonas reinhardtii Chlorella sp.

References Bayramoğlu et al. (2006) Akhtar et al. (2004) Rangsayatorn et al. (2004) Rangsayatorn et al. (2004) Maznah et al. (2012) Mehta and Gaur (2001) Bayramoğlu et al. (2006) Bayramoğlu et al. (2006) Mehta and Gaur (2001) Bayramoğlu et al. (2006) Bayramoğlu et al. (2006) Maznah et al. (2012)

Metal Ion Biosorption Enhancement Using Molecular Tools

Investigating biological mechanisms at the molecular level to produce bioengineered organism with higher biosorption capacity can be used for effective bio-removal of heavy metal. The high cost of conventional technologies to reduce toxic metal ion concentrations in industrial wastewater to acceptable regulatory standards has prompted exploitation of genetic and protein engineering approaches to produce cost-effective “green” biosorbents (Valls and de Lorenzo 2002). Another emerging area of research is the design and development of novel algal strains with increased affinity, capacity, and selectivity for biosorption of heavy metal ions (Zvinowanda et al. 2009; Karthikeyan et al. 2007). Many genes are involved in metal uptake, detoxification, or tolerance (Ayansina and Olubukola 2017). Cysteine-rich peptides such as glutathione (GSH), some lipopolysaccharides, phytochelatins (PCs), and metallothioneins (MTs) bind metal ions (Cd, Cu, Hg, etc.) and enhance metal ion bioaccumulation (Zhang et al. 2015). Based on the above said technologies (AAB, PAB, immobilized algal cell, molecular tools), various efforts have been made to commercialize the algal-based

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heavy metal removal (Zhou and Haynes 2010). In this context, Alga SORB is a commercial product comprising of PAB to remove the HM ion from aqueous solution. Another biosorbent, BIO-FIX, consisting of Sphagnum peat moss, algae (Ulva sp.), bacteria, and fungus encapsulated in polysulfone has also been found to have the potential to remove a variety of heavy metal ions. Thus available commercial technologies are dominated with adsorption process. Although microalgal potential to absorb a variety of heavy metal ions is known, sustainability of process is still under concern.

5 Conclusion This review revealed the contribution of algal biomass for heavy metal removal from wastewater. Low-cost cultivation and high HM ion uptake capacity, with suitable environmental conditions (pH, temperature, contact time, etc.), make algae biomass as a potential source for wastewater bioremediation. Microalgae, predominantly, possess numerous considerable sequestering mechanism HM ions and hence are demarcated as promising biosorbents. Several reports suggest their supremacy over various traditional and physiochemical methods and their usefulness in large-scale remediation of wastewater. Based on the biomass productivity of algae on wastewater is attractive dual-use algae cultivation coupled with other downstream or hybrid production systems. A suitable and sustainable approach needs to be developed in the field to select the most appropriate biosorbents and favorable physical conditions and to find out the major challenges involved. However, it is essential to deliberate on various avenues of microalgal remediation technologies as eco-friendly alternatives for a better environment. Acknowledgment Authors want to acknowledge the University Grants Commission for the necessary support in the accomplishment of the present work.

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Akhtar N, Iqbal J, Iqbal M (2004) Removal and recovery of nickel (II) from aqueous solution by loofa sponge-immobilized biomass of Chlorella sorokiniana: characterization studies. J Hazard Mater 108(1–2):85–94 Akhtar N, Iqbal M, Zafar SI, Iqbal J (2008) Biosorption characteristics of unicellular green alga Chlorella sorokiniana immobilized in loofa sponge for removal of Cr (III). J Environ Sci 20:231–239 Aksu Z (2001) Equilibrium and kinetic modeling of cadmium (II) biosorption by C. vulgaris in a batch system: effect of temperature. Sep Purif Technol 21:285–294 Aksu Z, D€onmez G (2006) Binary biosorption of cadmium (II) and nickel (II) onto dried Chlorella vulgaris: co-ion effect on mono-component isotherm parameters. Process Biochem 41 (4):860–868 Ali H, Khan E, Sajad MA (2013) Phytoremediation of heavy metals-concepts and applications. Chemosphere 91(7):869–881 Arıca MY, Tüzün I, Yalçın E, Ince €O, BayramoGlu G (2005) Utilization of native, heat and acidtreated microalgae Chlamydomonas reinhardtii preparations for biosorption of Cr (VI) ions. Process Biochem 40(7):2351–2358 Ayansina SA, Olubukola OB (2017) A new strategy for heavy metal polluted environments: a review of microbial biosorbents. Int J Environ Res Public Health 14:94. https://doi.org/10.3390/ ijerph14010094 Babel S, Kurniawan TA (2003) Low-cost adsorbents for heavy metals uptake from contaminated water: a review. J Hazard Mater 97:219–243 Barakat MA (2011) New trends in removing heavy metals from industrial wastewater. Arab J Chem 4:361–377 Bayramoğlu G, Tuzun I, Celik G, Yilmaz M, Arica MY (2006) Biosorption of mercury(II), cadmium(II) and lead(II) ions from aqueous system by microalgae Chlamydomonas reinhardtii immobilized in alginate beads. Int J Miner Process 81:35–43 Brinza L, Dring MJ, Gavrilescu M (2007) Marine micro and macro algal species as biosorbents for heavy metals. Environ Eng Manag J 6(3):237–251 Chojnacka K, Chojnacki A, Górecka H (2005) Biosorption of Cr3+, Cd2+ and Cu2+ ions by bluegreen algae Spirulina sp.: kinetics, equilibrium and the mechanism of the process. Chemosphere 59:75–84 Chowdhury P, Elkamel A, Ray AK (2015) Photocatalytic processes for the removal of toxic metal ions. In: Sharma SK (ed) Heavy metals in water: presence, removal and safety. The Royal Society of Chemistry, Cambridge, UK, pp 25–43 Cobbett C, Goldsbrough P (2002) Phytochelatins and metallothioneins: roles in heavy metal detoxification and homeostasis. Annu Rev Plant Biol 53:159–182 Da Costa ACA, de França FP (1998) The behavior of the microalgae Tetraselmis chuii in cadmiumcontaminated solutions. Aquacult Int 6:57–66 David SD, Marina C, Jonatan UF, Maria Dalgaard M, Peter U, William GTW (2012) The cell walls of green algae: a journey through evolution and diversity. Front Plant Sci 3:82–81 De-Bashan LE, Bashan Y (2010) Immobilized microalgae for removing pollutants: review of practical aspects. Bioresour Technol 101(6):1611–1627 Dominic VJ, Soumya M, Nisha MC (2009) Phycoremediation efficiency of three micro algae Chlorella Vulgaris, Synechocystis Salina and Gloeocapsa Gelatinosa. Acad Rev 16 (1&2):138–146 Dönmez GÇ, Aksu Z, Öztürk A, Kutsal T (1999) A comparative study on heavy metal biosorption characteristics of some algae. Process Biochem 34:885–892 Doshi H, Ray A, Kothari IL, Gami B (2006) Spectroscopic and scanning electron microscopy studies of bioaccumulation of pollutants by algae. Curr Microbiol 53(2):148–157 Doshi H, Ray A, Kothari IL (2007) Bioremediation potential of live and dead Spirulina: spectroscopic, kinetics and SEM studies. Biotechnol Bioeng 96(6):1051–1063 Doshi H, Seth C, Ray A, Kothari IL (2008) Bioaccumulation of heavy metals by green algae. Curr Microbiol 56:246–255

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Chapter 4

Recent Advances in Phytoremediation of Toxic Metals from Contaminated Sites: A Road Map to a Safer Environment Mukesh Kumar Awasthi, Di Guo, Sanjeev Kumar Awasthi, Quan Wang, Hongyu Chen, Tao Liu, Yumin Duan, Parimala Gnana Soundari, and Zengqiang Zhang

Abstract Toxic contaminants, or metal and metal-containing compounds, that are released into the environment from various anthropogenic sources cause severe environmental problems by destroying soil fertility, causing scarcity of resources as well as affecting human health. Thus, remediating environmental pollution, especially heavy metal contamination, is necessary in overcoming negative impacts on ecosystem health. Heavy metal (HM) contaminants threaten both human and environmental health. One report states there are more than 1.7 million metalcontaminated sites in central and eastern European countries that require appropriate reclamation. It was also observed that severe soil and water pollution in developing countries such as China, Pakistan, India, and Bangladesh results from small industrial effluent outputs over and near agricultural areas. Phytoremediation, although it is not new, is an efficient method to clean up toxic contaminants by employing different plant species. Although this technology is successful at the laboratory level, reports underlining the unsuccessful and inconclusive attempts in its use at the field level encouraged us to critically access why it is not satisfactory in the field, and also to find evidence that it is a promising remedial strategy without emphasizing negative perceptions. Analyzing the previous reports suggests two main themes for our attention. (1) Plant stress factors pose challenges in field application, although such were negligible with laboratory and greenhouse acclimatization. (2) Methods of phytoremediation should be assessed because often the decrease in contaminants is not adequate to demonstrate the occurrence of active remediation. Keeping these points in mind, this chapter focuses on the challenges in remediation, emphasizing rhizoremediation with detailed approaches in advanced technologies to confer environmental safety and assure human health.

M. K. Awasthi · D. Guo · S. K. Awasthi · Q. Wang · H. Chen · T. Liu · Y. Duan · P. G. Soundari · Z. Zhang (*) College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province, People’s Republic of China © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_4

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Keywords Phytoremediation · Anthropogenic sources · Rhizoremediation · Toxic contaminants · Environmental safety

1 Introduction The environment all over the world faces severe pollution as a result of the rapid growth of industries and modern technological development, with the inevitable release of huge amounts of toxic metal effluents. This level of pollution kindles environmental and public health concerns about these metal contaminants. The heavy metals (HMs) have very important applications in industry, agriculture, and domestic and technical development. Effluents from geogenic, agricultural, industrial, pharmaceutical, and even domestic and atmospheric processes are the major sources of heavy metal contamination. Also, many of the point source areas such as mining, smelting, foundries, and other metal-based industrial operations are the signature source of heavy metals into the environment. The intact heavy metals that are naturally present in the Earth’s crust are not classified as a source, whereas those resulting from various anthropogenic activities such as mining and smelting, production, domestic use, and agriculture are considered to be contaminants. Various industrial plants such as refineries, power plants, petrol combustion, nuclear stations, high tension lines, plastics, microelectronics, paper processing, and wood preserving contribute greatly to HM pollution. The release of metals from the Earth’s crust caused by various natural processes such as weathering, soil erosion of metal ions, leaching, atmospheric deposition, corrosion of metals, sediment re-suspension, and evaporation also contributes significantly to HM pollution. Many of the metals such as magnesium (Mg), iron (Fe), chromium (Cr), copper (Cu), selenium (Se), nickel (Ni), manganese (Mn), zinc (Zn), cobalt (Co), and molybdenum (Mo) are necessary for various physiological and biochemical functions because these elements function as essential micronutrients, the inadequacy of which results in various diseases and syndromes. Hence, heavy metals should be removed from the environment in a safe manner without affecting soil bioavailability. Phytoremediation is a significant method employing plants and other vegetation for in situ treatment in various metal-contaminated sites such as soil, sediments, and water resources. The roots of these plants and vegetation access the organic nutrient or metal pollutant, which after absorption will be sequestered, degraded, immobilized or mobilized, and metabolized. Research in the past decade has helped us to understand the uptake and fate of many toxic metals by phytoremediation to a greater extent. Better understanding of the absorption, uptake, transformation, and fate of toxic metal contaminants is discussed in full detail in this chapter. For every technological achievement there are pros and cons; similarly, the advantages of phytoremediation are its eco-friendliness, aesthetics, cost-effectiveness, and longterm applicability. Its application mainly covers contaminated sites, where other

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treatments are ineffective in practical use as well as very expensive; it can be easily adopted for low-level contaminated sites, where only superficial treatment is required for the long term. Phytoremediation can also be employed as a final step in conjunction with other technologies. The use of phytoremediation is limited because of the chance of introducing a contaminant or its metabolites into the food web. The process is slow to achieve regulatory levels, and toxic establishment and acclimatization in encountered sites also can result.

1.1

Background

Following the rapid development of industry and intensive agricultural activities, potent hazardous HMs are released into nature; as a consequence, the very basic and essential part of the ecosystem, the soil, becomes heavily contaminated. Up to the present, approximately 16.7% of farmland all over the world has been polluted by HMs such as Cr, Zn, and Pb. In addition, most of the HMs in farmland soil has exceeded the environmental quality limit, where the percentage of the exceedance was even greater at 19.4% (Liu et al. 2005). HM pollution in soil causes millions of tons loss of crop production every year. In addition to developed countries (American and European Union), developing countries such as China and India also have large areas of heavy metals-contaminated soil. It was estimated that in the European Union more than 3.5 million sites are heavily contaminated and 0.5 million sites have been seriously polluted. In America, HM-contaminated sites cover 600,000 ha (De Sousa and Ghoshal 2012; Perez 2012; Orooj et al. 2015). Compared to organic pollutants, the HMs are indestructible, which can result in their accumulation in the ecosystem and subsequent contamination of the food web with consequent harm to human health (Ali et al. 2013). As already mentioned, remediation of HM-contaminated soil has become a pressing problem worldwide. In the past few years, many approaches such as biological, physical, and chemical processes have been used in remediating HM-contaminated soil; for example, the soil replacement and leaching method, chelate extraction method, animal remediation, and phytoremediation (Sarwar et al. 2017). In all available approaches, the physical and chemical methods revealed serious limitations such as high cost, manpower, potential concomitant pollution, and alteration of soil character and native microflora, whereas phytoremediation uses the plants, rhizosphere, and microbially assisted collaboration in reducing the toxicity from HMs in soil (Rajkumar et al. 2012). When compared to other remediation methods, phytoremediation seem to be better for solving this problem, being also costeffective, eco-friendly, and clean. Many researchers have discussed the mechanism of phytoremediation for HM removal in contaminated soils and its development (Mahar et al. 2016; Sarwar et al. 2017).

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Phytoremediation: Present Outcomes

Phytoremediation is a flourishing field of active research with its novelty and costeffectiveness; being eco-friendly, aesthetic, and efficient, phytoremediation can be applied onsite, with remediation by solar-driven technology (Ali et al. 2013). The phytoremediation concept, first coined by Chaney (1983), includes phytoextraction, phytofiltration, phytostabilization, phytoevaporation, and phytodegradation. Over the past 20 years, phytoremediation has been widely used to remove contaminants (i.e., HMs, organic and inorganic substrates) from polluted soil or water by rhizofiltration and accumulation in the aboveground biomass (Ali et al. 2013; Rezania et al. 2016). Many kinds of hyper-accumulator were found; Brassica spp., a bioenergy crop, is found to be efficient in accumulating HMs in their tissues and recovering the contaminants (Wu et al. 2010). To date, many researchers have used plant-assisted biosolids or microbes to enhance the efficacy of the phytoremediation process in removing HMs from contaminated sites (Kim et al. 2010; Rajkumar et al. 2012). The addition of these organic substrate and microbes could possibly assist the plants in their growth and their HM adsorption capacities (Gao et al. 2010). Other than natural plants, genetically engineered plants with expected outcomes improve the remediation of HMs from contaminated soil (Chanu and Gupta 2016; Gomes et al. 2016).

1.3

Market Demand for Phytoremediation

Soil remediation is a difficult task because of current technical and financial limitations. Phytoremediation is a useful technology to harvest HMs from polluted soil and has been certified as an effective and economical method (Figs. 4.1 and 4.2). Compared to other technologies, phytoremediation is more suitable for market demand, having low installation and maintenance costs at less than 5% of the cost required for alternative methods (Prasad 2003). The plants remediated the contaminants without affecting the topsoil, consequently conserving the utility and fertility of the soil. Also, HMs with market value (e.g., Ni, Tl, Au) could be retrieved by plants. Furthermore, rapidly growing plants with high biomass yield, such as poplar, willow, and Jatropha, not only remediate the soil but can also produce energy (Prasad 2003; Abhilash et al. 2012). Therefore, employing phytoremediation could reduce HM pollution and create economic benefits.

1.4

Global Overview of Toxic Metals Phytoremediation

In the development of phytoremediation in HM-polluted soil, many plants and toxic metals have been widely investigated (Prasad 2003; Kim et al. 2010; Ali et al. 2013;

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Volatilization Accumulation

Conversion of metals and metalloids to less toxic and volatile forms

Enhancing metals and metalloids ligand production (MTs, PCs, GSH)

Detoxification Manipulating metals/ metalloids transporter genes and uptake system (influx/ efflux)

Modification Translocation

metals/ metalloids Uptake Transport

Fig. 4.1 Potential biotechnological strategies for phytoremediation. Toxic elements can be mobilized and transported (influx) into roots through plasma membrane transporters. They can then be transported (efflux) out of the roots into the xylem and translocated into the shoots. At this stage, plant tolerance to toxic elements may be enhanced through manipulation of influx/efflux transporters or by increasing the levels of ligands/chelators. Volatilization of the toxic elements can be achieved through enzymes that modify these toxic elements. Chelators or efflux transporters can also be used to export the toxic elements out of the cytosol and into vacuoles or the cell wall. (Source: Dhankher et al. 2011)

Mahar et al. 2016). The negative effects of heavy metals such as Hg, Cd, Cr, Pb, and Ar on soil, water, plants, animals, and humans have been studied extensively (Salem et al. 2000; Tripathi et al. 2007; Gulati et al. 2010; Iqbal 2012). Moreover, many studies have explored the efficiency of different plant varieties for remediating different toxic metal contaminants (Kim et al. 2010; Gomes et al. 2016). For example, the efficiency of accumulating Cu, Zn, and Pb by the plant species Eleocharis acicularis was reported by Ha et al. (2011). Oosten and Maggio (2015) and Gomes et al. (2016) also reviewed the use of genetically modified halophytic plants in remediating HM-contaminated soils.

2 Source and Impacts of Toxic Metal Contamination in Soil With the rapid development of global industry, HMs pollutants enter the soil through a variety of paths. In natural sources, parent soil materials and the soil-forming process are the two most important factors. Contaminated soil not only affects the yield and quality of agricultural products, but also relates to the quality of the

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Fig. 4.2 Diagrammatic model shows the process for reducing ethylene levels in roots by using bacterial 1-aminocyclopropane-1-carboxylic acid (ACC) deaminase. ACC synthesized in plant tissues is believed to be exuded from plant roots and is taken up by rhizobacteria where ACC is hydrolyzed to ammonia and 2-oxobutanoate. (Source: Tak et al. 2013)

atmosphere and the aqueous environment, harming human health via the food chain. Human activity is regarded as the main cause of heavy metal pollution of soils. Various industrial operations (e.g., metal forging, mining, manufacturing batteries, combustion of fossil fuels, smelting), agrochemicals, and long-term release of

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sewage sludge to arable lands contribute to the load of metals in the soil (Giller et al. 1989; McGrath et al. 1995; Anderson 1997). Industrial and mining enterprises directly discharge liquid effluent and solid waste; these discharges are not managed strictly, leading to a decline of soil quality and HM pollution. Smoke and exhaust emissions from these factories also contain heavy metals, which eventually enter the soil through natural sedimentation and rain. The industrial effluents, both solid and liquid, are highly enriched with HMs when compared with other soil background values.

2.1

Coal Mining and Combustion

As one of the main sources of electric power, coal mining and combustion have promoted the economy to a great extent. At the same time, the procurement, machining, and combustion processes for electric power production from both raw coal and waste by-products release numerous contaminants into the ecosystem. After coal is combusted for electricity, the ash, enriched with toxic metals, could flow into the atmosphere and then fall to the soil when wet or dry deposition occurs. The residues of combustion will also pollute the soil if there are no measures for separating them. The power industry wastes are HM enriched because the raw materials have already undergone a natural concentration process in their formation.

2.2

Metals Mining and Smelting

The environmental problems associated with the mining and smelting of precious and semiprecious metals, especially soil HM pollution, have a long history. As a result of the constant increase in the market value of precious metals such as gold and silver and other semiprecious metals (e.g., nickel), new sites are explored; mining and smelting results in the addition of HM loads in China and other developing countries. In the exploitation of photogenic sulfide deposits, HMs are easily poured into the mine soil through natural oxidation, the long-term leaching of the discarded sulfide minerals. When mine drainage and rainfall occur, the stream is mixed with metal ions, which when released by acid wastewater will flow into the soil. Moreover, the dust particles produced in mining, transport, and the casting process cannot be ignored as a soil HM source. The physical and chemical treatments employed in the ore-smelting process to extract a specific metal frequently release other metals (especially cadmium, arsenic) and constituents into the water that is used during the process or other solid waste which results from the process. During heating (smelting), metals are volatilized and emitted into the air in gas form, then adhere to atmospheric dust particles when the temperature decreases.

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Municipal Solid Waste and Industrial Waste Landfills

Millions of tons of municipal solid waste and industrial waste are produced every year. During this solid waste processing and disposal, huge quantities of toxic metals are leached into and pollute water bodies. Although a small quantity of HMs is essential for plant growth, surface and groundwater bodies are said to be contaminated when there is a huge runoff of HMs from agricultural applications and excessive leaching of HMs. A few reports suggest that this has become a major issue, especially in the United States and China (Wu et al. 2010; Sarwar et al. 2017). Even the low concentrations of HMs from contaminated soils and water that are released into shallow lakes and estuarine and marine waters are very toxic, threatening the lives of fish and other wild animals, and also posing serious problems in the ecosystem (Shi 2013). This realization necessitated development of eco-friendly green technologies that improve the ecosystem with positive impacts while preventing toxic HMs from leaching into the environment, also assisting the efficient management of HM shortages. Plant-assisted remediation has been so far recognized to be the better option for positive impacts of HMs as well as ensuring ecosystem and human health (Song et al. 2003; Tandy et al. 2006).

2.4

Oil Transport, Refining, and Utilization

Oil refineries produce a variety of metal-laden wastes, which cannot be avoided with current technology. The fossil raw materials formed over the timescale for oils is reported to be the source of HMs similar to coal. When oil products are used for energy, pollution problems occur. For example, vehicular exhaust, a hotspot we have discussed, has brought many environmental problems. Actually, as gasoline is produced by using crude oil as the raw material, oil probably causes more serious impacts on soil quality than coal because oil has a higher concentration of HMs than does coal.

2.5

Agricultural Irrigation

Although pesticides, chemical fertilizers, and mulching film are important agricultural materials that strongly promote the development of agricultural production, they also lead to soil HM pollution if utilized incorrectly. Most pesticides are organic compounds; individual varieties contain HM elements such as mercury, arsenic, copper, and zinc. The HM content in nitrogen fertilizer and potash fertilizer is low compared with the higher content in phosphate fertilizer. Farmyard manure is sometimes chosen as the base fertilizer applied to farmland without considering that heavy metal elements have been added to fodder as original materials at the first

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step. The application of mulching film that contains metal elements is also a threat to the soil environment. Sewage irrigation generally refers to using disposed urban sewage to irrigate farmland, forest, and grassland. The rapid development of urban industrialization causes a large amount of industrial wastewater to flow into wastewater treatment plants, so that many of the HM ions contained in urban sewage then enter the soil. Generally, the closer to industrial areas the more serious is pollution in the soil environment. In a word, many avenues will contaminate the soil, either industry sources or agriculture sources. The entire world faces the challenge of HM contamination in soil. It is understood from previous reports that in the United States many fields, such as 600,000 brownfields, were severely polluted with HM and require proper reclamation (McKeehan 2000). Similarly, irrigation water in the country of Georgia has been contaminated over the decades by HM-laden mining wastes (Felix-Henningsen et al. 2010); more than 1,400,000 sites in Western Europe have become contaminated with HMs (Wei et al. 2005).

2.6

Biosolids and Livestock Manure Application

Wastewater treatment plants and livestock industries all over the world produce large amounts of biosolids and manures, which contain huge quantities of organic matter, nutrients, trace elements, and a variety of infective agents (Chen et al. 2010; Awasthi et al. 2016). The traditional application of biosolids and manures onto agricultural lands over a period of time leads to the accumulation of toxic metals and also affects the ecosystem with a range of contaminants as well as odor pollution (Lüo et al. 2014). However, many technological solutions have been tried to overcome these problems by effective methods for reducing the levels of HM toxicity in biosolids and manures (Külcü and Yaldiz 2014), and also by converting toxic organic matter to a stable, safe, and sanitary fertilizer (Bernal et al. 2009). The excessive application and presence of high concentrations of HMs during the past two decades suggest several disadvantages of amendments in the soil and organic farming.

2.7

Fertilizer and Pesticide Application

Various heavy metals such as Cd, Zn, Pb, Cu, and As have begun to accumulate in the soils because of the extensive use of fertilizers and pesticides over time (Ali et al. 2013). The quality and growth capacity of a plant are consequently influenced by the accumulation of HMs in the upper soil surfaces from the excessive application of chemical pesticides and fertilizers, and also, from their production, as runoff from agricultural fields and industries (Sarwar et al. 2017). Many of the agriculturally dependent developing countries such as China and India are severely affected by these problematic issues.

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Nuclear Weapons Trial

Nuclear weapons trials are also a source of HM pollution. Uranium and plutonium, the main components of nuclear weapons, present relatively strong radioactivity. After a nuclear weapons trial, these elements enter into the soil, and their high radioactive properties are harmful for all the living organisms.

3 Types of Toxic Metals About 45 HMs occur in the periodic table of elements, and the toxicity of these elements varies from one to another. Therefore, only several metal elements with high toxicity are a concern, including heavy metals such as copper (Cu), cadmium (Cd), chromium (Cr), zinc (Zn), mercury (Hg), lead (Pb), nickel (Ni), and metalloids such as arsenic (As) and selenium (Se). Actually, copper and zinc, and even ferrum (iron) (Fe) and manganese (Mn), are necessary for plant growth, being involved in the activation of plant enzymes and formation of carbohydrates, for example. Similarly, selenium is critical in forming seleno-proteins for humans and animals, such as glutathione peroxidase (GPx) and thio-redoxin reductases (TrxR) (Barcelo and Poschenrieder 2011; Kaur et al. 2014), which benefit DNA synthesis, thyroid hormone metabolism, reproduction, and protection from oxidative damage and infection. However, some HMs such as Cd, As, Hg, and Pb are very toxic even at low concentrations and are not essential for any system. When the concentration of HMs in soil is more than the crop needs and is outside the tolerable range, the crop demonstrates toxicity symptoms, and then HM contamination occurs. HM contamination is the main threat causing severe effects on arable lands worldwide (Lin et al. 2012), possibly by making the lands barren for agricultural production. The uptake of HMs by plant crops depends on their chemical form in the soil and their bioavailability to the plant (Richard et al. 2000; Rajakaruna and Boyd 2008). The absorption of HMs by plants is influenced by soil physiochemical properties, including concentrations of HMs and other trace elements in the soil, clay content, soil temperature and moisture, soil pH, redox potential, cation-exchange capacity (CEC), soil organic matter (SOM), and aeration (Neilson and Rajakaruna 2012; Gall and Rajakaruna 2013). Among these, pH and SOM are the two most important factors contributing to the bioavailability of HMs (Guo et al. 2011). Generally, most HMs becomes more bioavailable under acidic pH ranges, and other soil conditions (Kumar et al. 1995) including HM concentration. However, the ability to endure HMs varies from plant to plant. A high concentration of HMs in the soil promotes the growth of some plants whereas HMs are toxic even at low concentration levels for other plant varieties. The bioaccumulation of HMs in soil can occur in many ways: HMs enter the human body by the food web, such as soil–plant–human or soil–plant–animal– human pathways. HMs can react with a variety of substances in the human body,

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such as amino acids, nucleic acids, vitamins, fatty acids, and phosphoric acid, leading to physiological dysfunction of the body and eventually causing lesions (Shi 2013). Different forms of danger are posed by HMs by its reactivity in plants, animals, humans, and ecosystems. All need micronutrients such as copper and zinc in small quantities, but their excess leads to poisoning; exposure to even smaller quantities of As, Hg, Pb, and Cd may cause lethal toxicities as these metals are not required by biosystems. Cadmium is one of the most toxic and mobile HM elements, which ensures its availability for plants under acidic pH ranges (Felix-Henningsen et al. 2010). It is considered that cadmium can lead to damage of the kidney and even death if the intake is too great at one time. Zinc, with characteristics similar to cadmium, can also contribute to acute gastrointestinal and respiratory problems (Anderson et al. 2005). Lead, known to cause severe cognitive dysfunction, neurological damage, behavioral disorders, renal impairment, and also hypertension in humans, can remain in soil and bioavailable for thousands of years because of its stable nature (McLaughlin 2002; Flora et al. 2012). (i) Lead Lead (Pb) is ubiquitous by nature but is an element considered to be not biologically useful because the facile interference of Pb2+ with several biological enzymes can be toxic. Its toxicity thus poses severe health issues in many body functions such as reproduction and by causing muscular, neural, gastrointestinal, genetic, and even behavioral changes. It also threatens the environment by its unknown fate. (ii) Cadmium Cadmium is a metal well known to be carcinogenic, inducing mutations in humans. It influences the metabolism of calcium by altering its availability, causing problems associated with kidney functions such as hypercalciuria and kidney failure; it also alters the blood hemoglobin level, resulting in anemia and other related problems (Awofolu 2005). (iii) Chromium Complications of anosmia and alopecia are caused by exposure to chromium (Salem et al. 2000). (iv) Zinc Zinc is an essential element for organisms, but fatigue and even dizziness can be caused by exposure to a high concentration (Hess and Schmid 2002). (v) Copper Many of the metabolic enzymes require copper for activation; thus, copper is considered to be a vital metal for many living organisms. Although vital, when its limit is exceeded, impairment in the functioning of brain and kidneys results; this is also associated with blood hemoglobin level. It can even disturb the digestive function by causing severe intestinal irritation (Salem et al. 2000).

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(vi) Mercury Mercury can be absorbed by the skin, respiratory tract, and digestive tract, injuring the brain and liver (Mahar et al. 2016). (vii) Arsenic Arsenic metal by nature does not exist in a free form; it usually exists as a metalloid. Its arsenide form is commonly seen with its association in sulfurcontaining ores. When in arsenate form it disturbs the cellular functions of living organisms by affecting the process of ATP synthesis and oxidative phosphorylations. Arsenic toxicity can actually reduce the chance of pregnancy and also inhibits enzyme activity (Tripathi et al. 2007). (viii) Selenium All living beings depend on selenium, as it is one of the essential trace elements. At excessive concentrations (>2.0 mg/kg), however, it causes selenosis and can be lethal (Li et al. 2013).

4 Phytoremediation of Toxic Metals The possible green ways to abolish metal pollution include either removing it completely from the environment by the assistance of plants or making it unavailable to the crop, followed by a different process of immobilization. The former is called phytoremediation, which in detail is transferring the pollutants in the environment to the plant or decomposition, immobilizing the pollutants by rhizosphere processes, and achieving the removal of pollutants and restoration of the ecological environment through the recycling of plants. Phytoremediation, when compared to the traditional physical and chemical alternatives of remediation techniques, shows technological improvement with sustainable remediation effects. It has been largely applied to the restoration of underground water and surface water by removal of HMs and organic and inorganic pollutants from soil. When compared with HMs, other pollutants can be biodegraded and simplified to non-bioavailable, less toxic components from a mother compound, whereas it is very difficult for HMs to take a less toxic or unavailable form and to be degraded or removed from the environment, signifying their indestructible nature.

4.1

Types of Phytoremediation and Mechanisms

In theory, the types of phytoremediation and mechanisms include phytoextraction, rhizodegradation, phytodegradation, phytostabilization, and phytovolatilization.

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Phytosequestration

Phytosequestration is a phenomenon in which the bioavailable contaminants trapped from the root systems and their entry into the root are reduced by several phytochemical complexations. Inhibition by transport proteins on the root membranes successfully prevents contaminant entry. The vacuolar storage of root cells also enhances the sequestration of contaminants into the vacuoles.

4.1.2

Rhizodegradation

Rhizosphere degradation is a process in which the plants stimulate the activities of root-associated soil microbial communities for contaminant degradation. Many of the aquatic plants also stimulate the microbial degraders, as proved in the case of atrazine by hornwort plants (Salt et al. 1995a, b; Sarwar et al. 2017). This process may also be referred to as phytostimulation.

4.1.3

Phytohydraulics

In phytohydraulics, the plants capture, transport, and transpire water from the environment to control the pollutants by hydrology. This mechanism does not degrade the contaminant.

4.1.4

Phytoextraction

The phenomenon of transferring the contaminants from the soil to the aboveground level by employing plants is called phytoextraction. The plants may either uptake or absorb the contaminant via the root system, and then transfer it to the shoot system, from which it can either be extracted as metal or create energy by burning.

4.1.5

Phytovolatilization

After taking up the bioavailable contaminants through the root system, the plant transfers them to the body of the plant above ground level, where it is digested or converted into the volatile form and will soon be transpired out of the plant system, along with water vapor, through the leaves and into the atmosphere. This phenomenon is called the phytovolatilization process.

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Phytodegradation

The complex organic molecules of the contaminants are either broken down and released as simple molecules, or the disrupted molecules will be incorporated into the tissues of the plant for growth and other metabolic activities. As the transformation of pollutants occurs, this may also referred to as a phyto-transformation process. Polycyclic aromatic hydrocarbons, polychlorinated biphenyls (PCBs), and other inorganic pollutants such as oxynitride and oxysulfide can be absorbed and degraded by plants (root, stem, leaf).

5 Phytoremediating Plants Plants that absorb heavy metals 100 times more than ordinary species, also known as hyper-accumulators, are selected for phytoremediation. The seed selection of hyperaccumulators is the basis for phytoremediation. The earliest report about hyperaccumulators can be traced back to the year 1583. Cesalpino, an Italian botanist, found a special plant growing on a “black rock.” It was afterward confirmed as a kind of hyper-accumulator of Ni. Generally, the definition of hyper-accumulator refers to the standard confirmed by Baker and Brooks (1989). Hyper-accumulators are reported for the metal-accumulating capacity in their leaves: the dry leaf tissue of such plants may contain more than 100 μg/g Cd and Se, and comparatively more for other metals such as Cr, Pb, Ni, Cu, Mn, and Zn at about 1000 μg/g or more. In many hyper-accumulators the concentration of heavy metal was found to be greater in shoots than in the roots. The threshold of accumulation by these hyper-accumulators will be more than thrice than that of most species growing on normal soils and at least one order of magnitude higher compared to those growing on metal-enriched soils. The more than 500 species of metal hyper-accumulators already found are widely distributed in about 50 families of vascular plants. Brassicaceae, which contain many food crop species, make up the most, approximately 25%. Most belong to Ni hyperaccumulators, approximately 400 species. For Cu, cobalt (Co), Zn, Se, Pb, and Mn, the totals are, respectively, 37, 29, 21, 20, 17, and 13 species. The number of Cd and As hyper-accumulators is relatively less. Baker (1987) found that the shoots of Thlaspi caerulescens can accumulate 7000 μg g
1 Pb under hydroponics condition. With the discovery of Pteris vittata, a well-known As hyper-accumulator, phytoremediation of As-polluted soils has made considerable progress.

5.1

How Do Plants Adsorb Toxic Metals?

As in nutrient and water absorption from the environment by the roots of the plants, the heavy metals also absorbed and are translocated from the roots to the

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aboveground parts. The metals from the aboveground biomass of the plant are crucial, feasible, desirable, and effective when compared to its root counterpart (Ali et al. 2013).

5.2

What Is a Natural Hyper-accumulator Species?

Plants with a greater ability for absorbing a high content of metals from the contaminated environment and accumulating it in their aboveground biomass through their effective root systems are said to be natural hyper-accumulators. The process of absorption and accumulation of metals into the tissues of the aboveground biomass is also found be higher in rate when compared with other plant species (Wu et al. 2010).

5.3

How Do Plants Tolerate High Metal Concentration in Soil?

To grow in a metal-contaminated ecosystem, plants must possess two indispensable and inherent properties, avoidance and tolerance (Baker 1987). Avoidance means that the plants prevent the absorption of a high concentration of heavy metals from soils by some kind of external mechanism; in some, this is achieved by the prevention of uptake of toxic metals by the root systems. The root of these plants may produce organic compounds, and also, along with rhizosphere bacteria, transform toxic metals into unavailable forms. Otherwise, the selective permeability of the plasma membrane prevents the accumulation process. This phenomenon often happens in tolerant populations. However, accumulators (especially hyperaccumulators) usually do not prevent contaminant entry into their roots, and thus possessing tolerance to toxic metals is necessary. The tolerance depends on the ecological and genetic characteristics of plants, so it varies from species to species. These plants have evolved specific mechanisms that detoxify the effects of high levels of metals that are accumulated in their cells, which allows even extremely high levels of metal accumulation.

5.4

Mechanisms of Metals Uptake into the Root and Translocation to Shoots

The main point of HMs in soil extracted by plants is their bioavailability, which means the HM exists in an activated chemical form. The bioavailability of HMs in soil is to a large extent dependent on soil factors (as listed earlier); plant root exudates and even some microbial activities near the rhizosphere greatly influence

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this action (Brown et al. 1999). Actually, HM activity is often promoted in rhizospheric areas. As mentioned earlier, most metals or metalloids become more active under acidic condition because of their more soluble cationic forms. Further, oxidizing and aerobic conditions also contribute to the activation of HMs in soil compared to reducing and anaerobic conditions, under which heavy metals exist in sulfide or carbonate precipitate forms. The bioavailability sequence of the following metals in soil under the same conditions is Zn > Cu > Cd > Ni > Pb. The uptake of most metals into the roots takes place in the aqueous phase. There are many transport sites in the root cellular membrane; the roots absorb metal ions from the soil solution (Fig. 4.2). This process is divided into two steps, a fast linear dynamics stage and a following slower saturated adsorption stage, related to the adsorption of the root cellular wall and passage through the root cellular membrane, respectively (Lasat et al. 1996). The charge of the metal ion limits its free access across the cellular membrane, which hence needs the assistance of membrane proteins for effective transport inside the plant system (Fig. 4.3). The transport of absorbed metals from the roots to the shoots determines the effectiveness of the phytoextraction process. However, it is much more difficult for plants to transport metals from root to shoot than that from soil to root. Therefore, this process is the key of accumulation of HMs for plants. The long-distance translocation of metal Pb from the root to shoot of a plant was the limiting factor influencing the efficiency of phytoextraction that was reported by Blaylock and Huang (1999). Metal-containing sap can be lifted from root to shoot by plants with transpiration through water and nutrient transport channels; in some cases it may revert back, which also limits the efficiency, and root pressure is regarded as another factor that cannot be ignored. Metal-microbe interactions

Biosorption M

Bioaccumulation

2+ 2L

2+

M

–

2+

M (out)

2L– 2+

2+

M

M (in) 2L–

Bioleaching e.g. Heterotrophic leaching Insoluble + metal

Microbially-enhanced chemisorption of metals

MICROBIA CELL

Organic acid

e.g. Hydrogen uranyl phosphate

soluble metal-chelate 2-

2+

MO 2 Metal (oxidised soluble)

2-

CO3 + M

MO2 Metal (reduced insoluble) Metal-chelate

Biotransformation e.g. Bioreduction

2+

HPO 4 + M

e-

CO2 +M

2+

H2 S + M

2+

2+

MHPO4 MCO3

MS

Biomineralisation Biodegradation of chelating agents

Fig. 4.3 Interactions of microbial cells with metals. (Source: Mosa et al. 2016)

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Plant Mechanisms for Metal Detoxification

Plants, even hyper-accumulators, cannot grow well in metal-enriched soils if there is no internal detoxification. The primary condition for the enrichment of HMs by plants is that the root and stem cells of these plants can tolerate high concentrations of corresponding elements, called detoxification. One of the major mechanisms of detoxification is the compartment effect of vacuoles and their complexation on HMs. It has been confirmed that Zn in the vacuoles of Thlaspi caerulescens leaves was higher than that of the apoplast, and the same results are verified in Thlaspi caerulescens roots. Biotransformation, another mechanism of hyper-tolerance, reduces the toxicity of HMs through chemical reduction in plants or in combination with organic compounds. A report by Brooks et al. (1981) stated that organic acids were present at a higher concentration level in the plant Alyssum serpyllifolium when compared to other species. The cadmium that is accumulated by plants can be detoxified by a process of binding with thiol (SH)-rich peptides, phytochelatins (PC) (Vogeli-Lange et al. 1990; Salt et al. 1995a, b; Griller et al. 1989). Moreover, plants can eliminate the reactive oxide species brought by HM stress, shielding cells from toxicity. Also, hyper-accumulators require higher metal concentrations for normal growth than do other plants.

5.6

Plant Limitations

There are no plant species that can grow well in all environment conditions, even in sufficiently highly metal-enriched soils. Thus, the detoxification by plants is limited. Depending upon the content and type of metal at a specific site, 15 years or even more time is needed for successful remediation; which limits its practical applications. Actually, the absorbed dose of some hyper-accumulator plants is not that great because of their lesser biomass. In addition, soil pollution is sometimes a combination of organic and inorganic pollutants, so hyper-accumulators that have a role in a particular pollutant cannot meet the remediation requirements. For contamination under deep soils (deeper than 50 cm), phytoremediation seems to be useless. There is still a long way to go for researchers to spend time on finding strengthening measures that can enhance the bioavailability of HMs in soils and optimal hyper-accumulator plants.

5.7

Improving Phytoremediating Plants

The traditional methods in practice limit the usage of these plants as such on a large scale, so it is necessary to improve the plant species to make it more efficient for the remediation process. One report states many modern tools in the field of chemical,

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biological, and genetic level engineering of plants make it possible to enhance the efficiency of phytoremediating plants in remediation (Sarwar et al. 2017).

5.8

Natural Hyper-accumulator Plants

With the development of phytoremediation, hyper-accumulator plants have been found. For example, Pteris vittata can significantly adsorb As from the soil; Zn and Cu can be accumulated by Thlaspi caerulescens; and Sedum alfredii is also a hyperaccumulator plant for Zn, Cr, and Pb (Kim and Owens 2010).

5.9

Transgenic Plants

With the fast development of life science theory and molecular biology technology, genetic engineering is considered one of the most effective approaches for phytoremediation. Genetic engineering technology allows plant metallothioneins (MTs), phytochelatins (PCs), and HM transporter genes into hyper-accumulators. Genetic engineering technology promotes accumulation by plants: first, through increasing the plant biomass; second, through reducing the toxicity of HMs to plants; and third, through improving the tolerance and resistance of plants to HMs. Among these, the third point refers to bacterial gene and enzyme expression. For example, the incorporation of metal-binding protein genes such as mammalian metallothionein into the tobacco plant system improves its metal tolerance level (Maiti et al. 1991). To overcome the deleterious effect and to withstand the herbicide application without reducing yield capacity, a new soybean species was formed by incorporating the glyphosate-resistant gene from bacterial 5-enolpyruvylshikimate3-phosphate synthesis (Padgette et al. 1995; Delannay et al. 1995).

6 Role of Genetics for Advancement of Phytoremediation Technology To improve phytoremediation potential, one has to breed the potential with superior remediation along with high biomass productivity; commonly the aspect of productivity is difficult to access via particular single gene insertion because it is controlled by the combined effects of many genes. Many authors employed genetic engineering by inserting squirrel genes into the plants to achieve high efficiency (Cunningham and Ow 1996; Brown et al. 1995a, b; Chaney et al. 2000). The taller the plant, the higher the biomass; so the efficient accumulator genes can be incorporated into taller plants than would be wise in natural alternatives for enhancing remediation

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Table 4.1 Genetically engineered plant growth-promoting (PGP) bacteria that enhance phytoremediation Genetically engineered PGP bacteria Pseudomonas putida KT2440 P. putida 06909

Mesorhizobium huakuii subsp. rengei strain B3 M. huakuii subsp. rengei strain B3 M. huakuii subsp. rengei strain B3

M. huakuii subsp. rengei strain B3 Enterobacter cloacae CAL2

Modified gene expression Phytochelatin synthase (PCS)

Associated plant Triticum aestivum

Expression of metal-binding peptide (EC20) Tetrameric human metallothionein (MTL4) PCSAT

Helianthus annuus

MTL4 and ATPCS

–

Iron-regulated transporter 1 gene from Arabidopsis thaliana (ATIRT1) EC 4.1.99.4

–

Astragalus sinicus –

Brassica napus

Beneficial features Production of phytochelatins (PCs) Production of metal-binding peptide Production of metallothioneins (MTs) Production of phytochelatins (PCs) Production of metallothioneins (MTs) and phytochelatins (PCs) Production of MTs and PCs, enhances nodule formation IAA, ACC deaminase, siderophores, antibiotics

Heavy metal (s) Cd, Hg, Ag Cd

Cd2+

Cd2+

Cd

References Yong et al. (2014) Wu et al. (2006) Sriprang et al. (2002) Sriprang et al. (2003) Ike et al. (2007)

Cu, Cd, Zn, As

Ike et al. (2008)

As

Nie et al. (2002)

properties. The plant growth-promoting rhizobacteria (PGPR), which enhance and assist the phytoremediation process, can also be engineered genetically for preferred properties such as withstanding biotic and abiotic stress, improved bio-derivative enzymes, metal homeostasis, metal chelation and transport, regulation of metal uptake, and mitigating risks (Abou-Shanab et al. 2006; Singh et al. 2011). A few of the expression genes and the features of some engineered PGPRs are listed in Table 4.1.

6.1

Manipulating Metal/Metalloid Transporter Genes and Uptake System

Although employing plants in HM remediation seems to be a greener way, there are limitations such as the toxicity imposed by heavy metals in plant systems. Also, the process itself is very slow, so these limitations must be overcome to make the plant

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very efficient in remediating toxic metals and metalloids of affected ecosystems (Dhankher et al. 2011). Figure 4.1 depicts the achievement of improved tolerance and metal accumulation by plants through genetic modification in metal transporters. A few reports showed the engineering of yeast cadmium factor (YCF1) overexpression in the plant Arabidopsis thaliana improved tolerance to HMs such as Cd(II) and PB(II); the plants even started accumulating HMs higher in their vacuoles followed by the conjugating glutathione (GSH) (Crowley et al. 1991; Song et al. 2003). Although we can clearly understand many of the soil–microbe– plant interactions in uptake, translocation, and accumulation of different types of heavy metals in their system, the fate of metals and the biological mechanisms underlying the methods of plant decontamination procedures are poorly understood, presenting a limiting factor that seeks our further attention.

6.2

Enhancing Metals and Metalloids Ligand Production

The variety of plant species and the type of metal contaminant decide the uptake and translocation of metals from roots to shoots of the plants. Differences in metal mobility are noted inside the plant system; when compared with Cu and Pb, metals such as Cd and Zn had more mobility inside the plants. During transport, many of the metals started to bind to the cell walls of plant roots, leading to their higher accumulation. So, to overcome this, chelation ligands such as organic and amino acids and thiols are important in facilitating their movement to the aboveground levels of plants, thus increasing remediating efficiency (Grill et al. 1987; Zacchini et al. 2009). Xylem cells have a high capability of cationic exchange, resulting in retreat of metal movement inside the shoot if not ligand chelated.

6.3

Rhizoremediation: The Combinatorial Effects of Bioand Phytoremediation

Rhizosphere microorganisms are exploited for use in remediating polluted environments, referred to as rhizoremediation, the combined approach of bioaugmentation and phytoremediation. Rhizoremediation laid a platform for engineering more than one microbial culture for desired factors of remediating a mixed type of pollutants in the co-contaminated substrate, which has attracted research (Khan 2006); especially, the bacterial communities draw much attraction for easy engineering effects in remediating different pollutants of co-contaminant sites (Wu et al. 2006; Yang et al. 2009). It was reported that the selection of the organism to be used is very important for successful rhizoremediation. Different varieties of grass and many of the leguminous plants such as alfalfa were found to be suitable plants for rhizoremediation (Kuiper et al. 2001; Qiu et al. 1994; Gupta and Sandallo 2011).

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Integration of “Omics” Tools for Developing Plants for Phytoremediation

By their nature, plants display their importance in remediating a polluted environment via different mechanisms such pollutant uptake through efficient root systems, accumulating pollutants inside the system, metabolizing them either alone or by microbial assistance, stabilizing pollutants, and volatilizing them out by photorespiration. Before the arrival of advanced technologies such as omics, science considered plants more visible than the microbial systems; later, the mechanisms behind phytoremediation were revealed by studying their genetic mechanisms, and also their molecular variability even in the same individual. High-throughput comparative analysis reveals interplant variability without focusing on single plant characteristic features such as genes and protein metabolism, which may assist researchers in selecting plant varieties and cultivation (Schmitz et al. 2013).

6.5

Plant Mechanisms for Metal Detoxification

In microbe-assisted phytoremediation, plants and microbes must stabilize, resist, or tolerate the heavy metals encountered in their systems. It is a natural phenomenon of adaptation to a stressed environment such as HMs by the plants via triggering its physiological or molecular mechanisms when exposed. A few of the mechanisms for adaption were found to be plant cell wall binding, transporting metals into the active vacuolar systems and intracellular complexations via chelating ligands such as phytochelatins and metallothioneins, as well as metal–siderophore complex sequestration into the root apoplasm or rhizosphere (Miransari 2011). Low molecular weight organic acids (LMWOAs) are exudates by plants enhancing microbial growth, solubilizing insoluble metal nutrients such as P, Zn, and Fe, and also used to detoxify metals such as As, Pb, and Cd by many of the metal-accumulating plants (Tu et al. 2004; Li et al. 2013). It is one of the best strategies by a plant for tolerating or excluding metals and metalloids via chelation in the apoplast or rhizosphere, preventing their entry into the cell symplast (Lena and Rao 1997; Magdziak et al. 2011).

6.6

How Do PGPR Combat Heavy-Metal Stress

It is very difficult to remove HMs from polluted sites as these are ultimately not destructible or not degraded biologically, because speciation and bioavailability vary according to the environmental changes whereas their other counterparts may undergo biodegradation by being less bioavailable, less mobile, and less toxic. Zn, Cu, and Ni (Olson et al. 2001; Li et al. 2013) are micronutrients essential for plants, animals, and microbes, but other metals such as Cd, Hg, and Pb are found to not have

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biological activity (Gadd 1992). Hence, these toxic metal pollutants have to removed or be made unavailable for biological systems by rendering them in a safe way. PGPR encounters HMs in the ecosystem and make them inactive through various processes of mobilization, immobilization, and transformation, through which they are adopted to survive in such stressed conditions (Nies 1999). Some of the mechanisms are given here: (1) exclusion, a process by which metals are excluded from the targeted sites; (2) extrusion, in which absorbed metals are pushed out from the cells via chromosomal/plasmid functioning; (3) accommodation, or conjugation with metal-binding proteins such as metallothioneins and other low molecular weight proteins (Kao et al. 2006; Umrania 2006) or other components of the cell; (4) biotransformation, the conversion of toxic to less toxic forms; and (5) methylation and demethylation. These methods lay a platform for survival ability and being metabolically active in such stress conditions.

6.7

Synergistic Interaction of PGPR and Plants in HeavyMetal Remediation

Plant growth promotion by PGPR is well documented (Reed and Glick 2004; Babalola et al. 2007; Babalola 2010). Recent reports say this interaction not only enhances growth but also reduces environmental stress to the plant where it grows; achieved by the synergic association with the plant root beyond many beneficial effects such as improving soil quality, fixing atmospheric nitrogen, enhancing plant performance, and also nutrient availability (Tinker 1984; Babalola et al. 2007; Li et al. 2013).

6.8

ACC Deaminase and Plant Stress Reduction from Ethylene

The ACC (1-aminocyclopropane) deaminase of certain PGPR enhances the uptake of inorganic contaminants in a site by plants via modifying their root architecture and their uptake abilities; this was achieved through modulating stress-induced production of ethylene by plants (Macek et al. 2000; Arshad et al. 2007). Plant growth inhibition and reduction in biomass is effectively achieved by stress ethylene biosynthesis, especially for roots (Glick et al. 2007). The PGPR with ACC deaminase is the only successful strategy to overcome the challenges in remediation by plants (Glick 2003; Gerhardt et al. 2006; Meplan 2011), which hydrolyzes the precursor to ethylene, 1-aminocyclopropane-1-carboxylic acid, thus lowering the ethylene biosynthesis rate (Glick et al. 1998; Ma et al. 2001), without being affected by giving space for the plant for the key defense response through a little burst of

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ethylene production when exposed to stress; rather, it only reduces the deleterious levels of stress ethylene by plants (Glick et al. 2007; Nagajyoti et al. 2010).

7 Mechanisms of Plant–Metal Interaction in the Rhizosphere The plant–microbe interaction responsible for biogeochemical cycling of metals and applications of phytoremediation came to light after the discovery of some plant– microbe–metal interactions (Fig. 4.2). Microorganisms are ubiquitous in nature: found everywhere, they grow well in all ecosystems such as soil, atmosphere, and water, even in some extreme conditions (Vidali 2001). One of the important habitats of all kinds of prokaryotic and eukaryotic microorganisms is the rhizosphere, which seems to have a good interaction with the nearby vegetation in various ways (Azevedo et al. 2000; Hao et al. 2012). Various metals for plant bioavailability are restricted because of some hindrance such as solubility levels and high affinity toward the soil, but some PGPR, root-colonizing, and mycorrhizal microorganisms enhance the bioavailability of metals to the plant systems. For example, rhamnolipids, a biosurfactant released by some bacteria, modifies the hydrophobic pollutants to more hydrophilic forms that are known to be released (Qiu et al. 1994; Volkering et al. 1998). The absorption of various metal ions such as Fe2+, Mn2+, and Cd2+ by the root system is facilitated by organic exudates of microorganisms via improving the bioavailability. Root exudates in turn feed the microbes by being good sources of carbohydrates, lipophilic compounds, and natural chelators such as citric, acetic, and other organic acids, which are important in increasing the mobility of metal ions or promoting the bio-surfactant-producing class of microorganisms (Fig. 4.2). The many feedback mechanisms by plant roots and a rhizosphere microorganism allow them to adapt to the conditions of their environmental habitats; for example, the root exudates of plants in a phosphorus-deficient habitat contain large amounts of citric acid as an attempt for mobilizing the phosphorus present in soils. In some cases, a few of the rhizosphere-dwelling organisms secrete growth hormones for the plant to increase the plant root growth, which in turn provides its nutrient root exudates.

7.1

Metal Bioavailability for Plant Roots

To remediate a pollutant from an environment, the plant or microorganisms should establish a close contact and have the ability to act on it; hence, the bioavailability of the pollutant is important for its remediation. Soil and environmental conditions and the pollutant physiochemical properties decide bioavailability. The soil has more sites for metal ions, especially in CEC, if it has small particle size (clay) by which it

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can hold water better than other soil types (Salt et al. 1995a, b; Taiz and Zeiger 2002). The entry of metals into the root system is by symplastic or apoplastic pathways (Tandy et al. 2006; Lu et al. 2009). The symplast or active pathway, as the name suggests, is an energy-associated process by which entry is mediated by a specific ion carrier or channels, whereas in the passive or apoplastic pathway the metal or its complexes enter the root through intercellular spaces.

7.2

Effect of Soil/Rhizospheric Microorganisms on Metal Uptake

The rhizosphere soil nature and microbial communities are greatly influenced by the rooting or root growth of the plants; it was reported that the rhizosphere microbial population is of several orders of magnitude higher than that of nearer soil surroundings (Anderson 1997). These microbes symbiotically favor the root in metal uptake. In a few cases, some microorganisms excrete important organic compounds for facilitating metal bioavailability for roots and also for root absorption of essential micronutrients such as Fe (Crowley et al. 1991) and Mn (Barber and Lee 1974), as well as a few of the nonessential metals such as Cd (Salt et al. 1995a, b; Li et al. 2013); it sometimes alters the soil chemistry to make the metals more soluble.

7.3

Effect of Root Exudates on Metal Uptake

Root exudates have a significant role in phytoremediation. Some of the organic chemicals released by root systems in certain stages of growth induce plants to acclimatize to the stressed environment either by affecting rhizosphere microbial and other plant growth (allelopathic functions) or by inactivating the metal pollutant by root absorption, assimilation, chelation, and transformation (detoxification functions). In addition, root exudates, particularly organic acids, are able to bind metal ions, therefore influencing metal mobility, solubility, and bioavailability from soil (Chiang et al. 2011; Luo et al. 2014). Kim et al. (2010) suggest organic acids from root exudates act as natural chelators to enhance the phytoextraction process; the enhanced efficiency in translocation and bioaccumulation of Cd, Cu, and Pb is achieved with the help of citric acid and oxalic acid from Echinochloa crus-galli.

7.4

Enzymatic Transformations

The metabolism or transformation of metals by plants after absorption via root systems from contaminated sites is said to be a phyto-transformation. After entering the root system, the metals are translocated and undergo different phases of

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transformation (Ohkawa et al. 1999; Ma et al. 2001): (I) conversion (through oxidations, reductions, and hydrolysis), (II) conjugation (chelating with glutathione, sugars, and amino acids), or (III) compartmentation (by-products from the previous phase are converted, bonded to the cell wall or lignin, and deposited in some vacuoles of the plant system.

7.5

Phyto-Transformation Enzymology and Biochemistry

As other eukaryotes, the pathways for degradation of many organic pollutants are similar in plant systems (Lamoureux and Rusness 1986; Vogeli-Lange and Wagner 1990; Meplan 2011). Based on the activation and imposed toxicity, mutagenicity and carcinogenicity by chlorinated aliphatic compound degradation in human systems gain importance in research; the metabolism of such may vary even within the same homologous series, but often it releases free radicals after oxidization. This action was mainly observed with carbon tetrachloride metabolism when compared to a lesser extent in other chlorinated methanes (Henschler 1990). In the dechlorination pathway was observed the formation of epoxides when the glutathione is conjugated with polychlorinated ethenes such as trichloroethylene (TCE) and perchloroethylene (PCE) (Henschler 1990; Qiu et al. 1994), suggesting the suitability of microbes for phytoremediation purposes by holding the eminent capacity of transforming metal contaminates into a non-bioavailable state or reducing their toxicity to the plants and environment.

8 Optimization of Metal Phytoextraction with Agronomic Practices The extent of metal contaminant removal is greatly affected by the selection of a suitable plant variety for remediation. Ecosystem protection has to be considered before selecting the plants for remediation, although metal extraction is the main attraction; native plant species will be given importance excluding exotic plant usages because the latter may affect the harmony of the ecosystem by their invasion. The propagation of weedy crop species may be avoided by giving preference to the general crops, but the downside is the palatable risks it may pose to grazing animals after remediation.

8.1

Plant Selection

Plant selection is important for the phytoremediation process; to achieve good remediation of metal contaminants, hyper-accumulators may be given preference.

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The total biomass harvested and metal content in the same decide the rate of metal removal by the plants. The usage of nano-accumulator plants remediation of metals is limited by their slow growth and smaller size. The low potentiality of metal accumulation is often compensated by significant biomass production in common nano-accumulators (Ebbs et al. 1997); hence, it was always debated about choosing the remedial plant species between these two types (hyper-accumulator or nanoaccumulator).

8.2

Soil Fertilization and Conditioning

The practice of field management, which maintains soil structure and nutrient balance, is always necessary for making soil fertile and healthy. The capacity of water retention and infiltration with respect to water management and fertility previously was highly preferred. The land is considered to be viable when it has balanced nutrients to sustain microorganisms and plant growth; it should even support beneficial soil dwellers such as the earthworm, which help nutrient cycling, with other beneficial effects, and also restricting pathogenic microbial communities.

8.3

Enhancing Metal Bioavailability Using Soil Amendments

For enhancing plant growth, many materials are added to soils, called fertilizers or soil amenders. Whether inorganic (10-10-10 fertilizers) or organic (bone meal), many of these serve the purposes of both fertilizing and amending soils. To promote healthy plant growth, soil amendments are mixed in the top layer of soil, providing help in many ways such as changing soil pH or by improving nutrient supply. The utilization of organic wastes as soil amendments dates back for centuries (Sims and Pierzynski 2000; Burton and Turner 2003). Animal manures have drawn attention for many decades as a good source of essential nutrients beneficial for plant growth and maintaining soil fertility. Mostly organic wastes are significant pollutants that have needed remediation in recent years; many types of industry and even increased wastewater generation and livestock intensification result in the addition of some organic contaminants into the environment, which vary according to their constitution and components (Sims and Pierzynski 2000; Burton and Turner 2003). A few reports stated that some of the important contributors of organic wastes include paper mill factories and olive mills (Edwards and Someshwar 2000; Suresh and Ravishankar 2004).

8.3.1

Sowing

Sowing is the process of planting seeds. After the preparation of soil, the previously selected seeds are scattered in the field. This is called sowing, which should be done

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carefully and uniformly. If seeds are not sown uniformly, overcrowding of the crop can happen. For sufficient sunlight, water, and other requirements, congestion should be prevented. Traditionally, sowing was done manually, whereas nowadays seed drilling machines are used.

8.3.2

Traditional Method

Here seeds are sown either by hand or by using tools. Sowing of seeds by hand is called broadcasting. This is a cheap method but uniform distribution is not maintained. Alternatively, a funnel-like tool filled with seeds is used. Seeds are passed through pipes deep into the soil.

8.3.3

Drilling Machine

This machine is a modern method where sowing is done by iron drills connected to a tractor. Here also funnels filled with seeds are present at the top of the drill. When the plough moves, seeds are distributed into the furrow made by the plough and covered. This method is more advantageous than the traditional method. Seeds are distributed at regular intervals and depth, and the method is also profitable in terms of both labor and time. Selection and sowing of seeds are agricultural practices that demand extreme attention and care.

8.4

Crop Rotation

Growing different kinds of dissimilar crops on the same soil area in different seasons is known as crop rotation, done to prevent only one type of nutrient remaining in the soil, which also helps in increasing soil fertility and crop yield by reducing soil erosion. For phytoremediation strategies, this kind of remediating plant variety rotation should initially be considered under field conditions; which may likely prevent plant diseases caused by different kinds of pathogens (Fusarium, Rhizoctonia, Alternaria), weed buildup, and even insect attacks.

8.5

Crop Maintenance: Pest Control and Irrigation

With soil fertility, selection of crops, and crop rotation, effective pest control should be efficiently managed to obtain the expected incremental changes in yield using irrigation systems. This is an effective agronomic practice that gains importance in soil reclamation and erosion control using irrigation. When these conditions are disturbed or ignored unknowingly without proper understanding, the lesser the yield.

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Weed control and effective irrigation are followed successfully according to the history. Herbicides, which also effectively control weed spread, can be applied before or after the emergence of phytoremediating crops. Pre-emergent herbicides are significant weed controllers improving quick emergence and establishment of the plant of interest, whereas the post-emergent herbicides are used to control weeds after the selected plant stabilizes its growth. Maintaining adequate moisture of soils is important because the movement of soil solution to the root surface from the soil greatly influences metal uptake.

8.6

Handling and Eco-friendly Disposal of Toxic MetalsContaminated Waste

Proper handling and decontamination of remediated plants without re-affecting the environment is of great concern. Waste is a direct result of human activity, and the quantity of waste generated directly indicates the development of a country in various terms of modernization and industrialization. Management of waste is an important challenge to the modern world. Waste management should efficiently ensure the safe and proper management of hazardous wastes from their generation until disposal, including their storage, handling, and safe transport. This is the responsibility of individuals who are generating waste (laboratory) and environmental services (i.e., REHS–Rutgers Environmental Health & Safety). The management should abide by the rules of local, state, federal, and regulatory laws. The unwanted hazardous chemicals should be routinely collected by REHS, and recycling of the chemicals, which can be used again at the laboratory level, is encouraged.

8.7

Economically Feasible and Time Projection Technology Application

To date, there are no reports to support the complete remediation of a contaminated site through phytoremediation and the cost for these processes is unknown. Available data only state the cost required for the short term may be 2–3 years of field study, through which we cannot decipher the cost required for the long term as the remediation may take 15 or more years, when placed less applicably in fields. Without minding these limitations, a few researchers reported a desirable timeframe and cost required for phytoremediating metals. Brown et al. (1995a) considered a soil contaminated with 400 mg kg
1 Zn and a desired cleanup level of 40 mg kg
1. The use of intrinsic or engineered bioremediation showed several advantages that inspired many of the site owners, regulatory agencies, and even the public: (1) more cost-effective than that of conventional counterparts; (2) resulting in innocuous

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product through conversion; (3) allowing continued usage of site because of being nonintrusive; and (4) above all, relatively easy implementation.

8.8

Research Advancement

Recent active research in the field of plant biology is phytoremediation, which has gained importance by its promising advances in the area of environmental concerns. Many plants of different varieties have already been reported for efficiency in remediation; attempts to understand the mechanisms behind the combined effects of plants such as uptake, translocation, and metabolism of metals have been widely reported, mainly in hyper-accumulators. The field of biotechnology and its implications helps to achieve improved sustainable remediation abilities of plants in conjunction with traditional technologies. The new variety of transgenic plants can efficiently manage large quantities of metal uptake and successfully manage their own metabolism, making the metals nontoxic to the environment, as acknowledged by one report. Although advances in this field have occurred, better understanding is required of the actual process by which the plant–microbial interactions take place for efficient metal accumulation and ionic homeostasis. For better understanding, the following areas need focus. (1) Safe compartmentation of HMs inside the plant tissues may be found by manipulating metal transporters toward some specific sites such as vacuoles without disturbing other plant cells. (2) To avoid the risk of trans-gene escape through pollen, an attempt must be made to manipulate the genomes of the chloroplast, which may be a better alternative in some plants. (3) Subsequent transformation of improved metal tolerances in plants in needed, along with identifying the substance produced to deter herbivores feeding on them and thus preventing HMs from entering the food web. (4) The plant–microbe interaction must be enhanced by developing transgenic plants, which may be achieved by two approaches: to develop transgenic plants that have the ability to secrete metal-selective ligands, which helps in metal solubility to make it bioavailable; or to identify and improve the natural secretion of simple molecules with selective chelation abilities in rhizosphere. (5) It is equally important to address the mixed contaminants at the same pollution site, so transgenic plants are to be created by opting for multigene approaches of the simultaneous transfer of several genes to a suitable plant candidate for efficient remediating of mixed contaminants at the same site. (6) However, field performance of the transgenic plants is always unexplored; immediate field trial experiments are necessary for an acceptable as well as a commercially viable technology. Several collaborative studies involving botany, soil biogeochemistry, microbiology, genetic engineering, agricultural and bioengineering, plant physiology, and biochemistry may provide better understanding about phytoremediating prospects. As in many other fields, the influence of proteomics and genomics may someday take phytoremediation and their genetic engineering to the next level, imparting benefits to society.

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9 Challenges and Limitation for Phytoremediation There is immediate need for greener technologies without serious implications for the environment. Phytoremediation appears stronger in such cases, although we need to acknowledge its limitations: researchers should consider the drawbacks and challenges to be encountered when bringing it to field trials as well as its importance in ecosystem well-being. Many strategies have been considered to overcome the challenges such as plant stress and other environmental conditions. Rhizoremediation is also promising and widely studied to improve its performance in greenhouses in conjunction with field trials to address field adaptation and overcome the drawbacks. New protocols are needed for effective sampling, analysis, and interpretation to represent results acquired from rhizoremediation as well as phytoremediation. Particular efforts have been focused on cost-effective method development that distinguishes between peterogenic and phytogenic carbon compounds, which may bring technology acceptance and even motivate regulators in changing guidelines for organic compound remediation. Rhizoremediation now has gained in importance over phytoremediation as being a very compatible, effective, and cost-efficient method that actively removes HMs and other organic pollutants from affected ecosystems.

10

Concluding Remarks and Future Aspects

When compared to the past, phytoremediation now shows significant development in actively taking a part in environmental health. However, this technology, now extensively applied in pilot and field applications, always needs greater understanding of the interactions of rhizosphere factors such as soil, metal, pollutants, microorganisms, and roots to make it more successful. Plant tolerance to metal hyperaccumulation and the metabolism and fate of metal ions inside plant tissues can only be explored by technological developments in spectroscopy and chromatography to bring credit to this greener way of cleaning up contaminated environments. Environmental bioavailability of metals to root systems also kindles research focused on enhancing their solubility rates and conjugation levels with environmentally harmless chelating chemicals, highlighting cost-effectiveness and efficiency standards. Different plant species with effective rates of extraction of metals in the same environment should be identified and selected for crop rotation to make the process complete within a given timeframe; the process has to be optimized to reduce the time needed for various stages of the growth cycle and harvesting. For agro-mining, the most reliable sources are moderately polluted soil types. Phytostabilization is an alternative practical approach widely adopted worldwide for cleaning up abandoned and modern rural and industrial pollutant sites. Only a few types of hyperaccumulator plants react successfully with soils containing more than one type of toxic pollutant, suggesting supplementary remediation processes to remove

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co-pollutants from the environment. The effects of pH and exchange capacity (EC) in contaminated systems that govern transformations of HMs in both lowland and upland rice paddy fields need analysis. Solution-phase and solid-phase speciation of HMs in water and soil using advanced techniques must be further studied. Further research is required to identify biochemical mechanisms involved in the different HM uptake in plants, because little research has been conducted on the rhizosphere processes underpinning effective phytoextraction technology for HM cleanup from contaminated soils. Employing different beneficial microbial communities such as PGPR and mycorrhizal organisms enhances the remediation process effectively. Thus, selection and identification of new microbial strains with these qualities are required to better understand their functions and interactions in the rhizosphere, that is, the plant roots, among metals, microbes, and soil systems. Further analysis of the HM-enriched plant biomass is needed for reclamation or safe decontamination. Practical approaches are needed for successful application of the various remediation technologies, such as hyper-accumulators, chelators, microbes, and soil amendments suitable for effective remediation of different HM-polluted environmental sites. Acknowledgments The authors are grateful for financial support from a Research Fund for International Young Scientists from National Natural Science Foundation of China (Grant No. 31750110469), and The Introduction of talent research start-up found (No. Z101021803). We also thank our laboratory colleagues and research staff members for their constructive advice and help.

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Chapter 5

Emerging and Ecofriendly Technologies for the Removal of Organic and Inorganic Pollutants from Industrial Wastewaters Gaurav Saxena, Surya Pratap Goutam, Akash Mishra, Sikandar I. Mulla, and Ram Naresh Bharagava

Abstract Environmental pollution is one of the major problems of the current world, and providing a sustainable solution to manage pollution is a key challenge. Industries are mainly responsible for the environmental pollution as they discharge highly toxic pollutants in the receiving environment and provide chance for exposure to mankind and, thus, may create toxicity in humans and animals. The physicochemical methods used for the removal of a variety of organic and inorganic pollutants from industrial wastewater are costly and environmentally destructive and may create secondary pollution and, thus, ultimately deter the environmental quality. To overcome these problems, various emerging and ecofriendly technologies are becoming popular for the removal of various pollutants from industrial wastewaters. Therefore, this chapter provides an overview of the various emerging and ecofriendly technologies for the removal of organic and inorganic pollutants from industrial wastewaters with their merits and demerits.

G. Saxena · R. N. Bharagava (*) Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Microbiology (DM), Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, Uttar Pradesh, India S. P. Goutam Advanced Materials Research Laboratory, Department of Applied Physics (DAP), School for Physical Sciences (SPS), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India A. Mishra Defence Institute of Bio-Energy Research (DIBER), Defence Research and Development Organization (DRDO), Ministry of Defence, Government of India, Haldwani, Uttarakhand (UK), India S. I. Mulla CAS Key Laboratory of Urban Pollutant Conversion, Institute of Urban Environment, Chinese Academy of Sciences, Xiamen, People’s Republic of China © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_5

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Keywords Industrial wastewaters · Organic pollutants · Inorganic pollutants · Emerging and ecofriendly technologies · Waste treatment and management

1 Introduction Environmental pollution is a serious problem of the current world. The pollution of our natural environment is mainly caused by anthropogenic activities. Anthropogenic activities mainly include industrial operations. Industries are the key players in the national economy of every nation but are also the major polluters as they discharge a high-strength wastewater mainly characterized by high physicochemical parameters such as biochemical oxygen demand (BOD), chemical oxygen demand (COD), total dissolved solids (TDS), and a variety of organic and inorganic pollutants, which cause serious threats to public health and environment (Mendez-Paz et al. 2005; Chandra et al. 2008, 2011; Saxena and Bharagava 2015, 2016; Saxena et al. 2016; Gautam et al. 2017; Arora et al. 2014, 2018; Goutam et al. 2018). Physicochemical methods are currently used for the treatment and management of industrial wastewaters. However, these remediation methods require high cost to be applied, are environmentally destructive, and may create secondary pollutants that cause further pollution of the environment. To overcome these problems, various emerging and ecofriendly approaches are applied for the removal of various organic and inorganic pollutants from industrial wastewaters. Therefore, this chapter provides an overview on the various emerging and ecofriendly methods applied for the treatment and management of highly toxic and hazardous industrial wastewaters with their merits and demerits.

2 Ecofriendly and Emerging Technologies for Removal of Organic and Inorganic Pollutants from Industrial Wastewaters 2.1

Microbial Bioremediation

It is an ecofriendly remediation technique that uses the inherent ability of microbes such as algae, fungi, and bacteria to degrade/detoxify organic and inorganic pollutants from industrial wastewaters (Kishor et al. 2018; Bharagava et al. 2019). In bioremediation process, waste is converted into inorganic compounds such as carbon dioxide, water, and methane and, thus, leads to mineralization and detoxification (Reshma et al. 2011). Bioremediation chiefly depends on the metabolic capability of microbes to degrade/ detoxify or transform the pollutants, which is also affected by the accessibility of pollutants and their bioavailability (Antizar-Ladislao 2010). It can be applied as both in situ (remediation at the site) and ex situ (remediation elsewhere) remediation

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technology. Bioremediation involves (Maszenan et al. 2011) bioattenuation (natural process of degradation and can be monitored by a decrease in pollutant concentration with increasing time), biostimulation (intentional stimulation of pollutant degradation by addition of water, nutrients, and electron donors or acceptors), and bioaugmentation (addition of laboratory-grown microbes with potential for degradation). A great deal of literature can be found in the public domain on the biodegradation and bioremediation of industrial waste/pollutants using individual microbe and microbial consortia (see, Pandey et al. 2007; Maszenan et al. 2011; Singh et al. 2011; Megharaj et al. 2011; Paisio et al. 2012; Saxena and Bharagava 2017; Saxena et al. 2015; Bharagava et al. 2018). For instance, Kim et al. (2014) reported 98.3% of COD and 88.5% of Cr removal from tannery wastewater (TWW). Noorjahan (2014) reported 90% of COD, 90% of BOD, and 63.8% of Cr removal from TWW using E. coli and 95.4% of COD and BOD and 73.5% of Cr removal from TWW using Bacillus sp. Yusuf et al. (2013) reported 87.6% of COD from TWW using B. subtilis and 85.2% of COD from TWW using P. fragi. El-Bestawy et al. (2013) reported 79.16 of COD, 94.14 of BOD, and 93.66 of Cr from TWW using an optimized bacterial consortium containing Providencia vermicola W9B-11, Escherichia coli O7:K1 CE10, Bacillus sp. 58, Bacillus amyloliquefaciens T004, Pseudomonas stutzeri M15-10-3, and Bacillus sp. PL47. Sivaprakasam et al. (2008) also reported 80% of COD removal from TWW using a bacterial consortium (P. aeruginosa, B. flexus, E. homiense, and S. aureus).

2.2

Phytoremediation

Phytoremediation is a low-cost and eco-sustainable in situ remediation technology. It is advantageous over the conventional physicochemical cleanup methods that require high capital investment and labor, alter soil properties, and disturb soil microflora. Phytoremediation is a type of bioremediation wherein green plants with associated microbes are used for the removal of toxic metals from the contaminated matrix to safeguard the environment and public health. It involves different strategies such as phytoextraction, phytostabilization, phytodegradation, phytostimulation, phytovolatilization, and rhizofiltration to remove metal pollutants from the contaminated sites (Lee 2013; Chandra et al. 2015; Chirakkara et al. 2016). It can be commercialized, and income can be generated, if metals removed from contaminated sites could be utilized as “bio-ore” to extract usable form of economically viable metals (i.e., phytomining) (Chandra et al. 2015; Mahar et al. 2016). Bioenergy can be generated through the burning of plant biomass, and land restoration can be achieved for sustainable agricultural development or general habitation (Lintern et al. 2013; Stephenson and Black 2014; Mahar et al. 2016). The rationale, mechanisms, and economic feasibility of phytoremediation have been discussed elsewhere (Ali et al. 2013; Wan et al. 2016; Sarwar et al. 2017). A great deal of literature can be found in the public domain on the phytoremediation of heavy metals

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from contaminated matrix (Ali et al. 2013; Chandra et al. 2015; Mahar et al. 2016; Sarwar et al. 2017). However, extensive research is currently underway to testify the phytoremediation potential of hyperaccumulating plants in the field for the effective treatment and management of HM-contaminated sites. Further, exploiting plant-associated microbes with desired traits to enhance the phytoremediation efficiency of hyperaccumulating plants via increasing the bioavailability of metals in soil and plant growth promotion in the stressed environment is termed as microbe-assisted phytoremediation. Inoculation of plants with plant growth-promoting bacteria (PGPR) may be helpful in phytoremediation as they can suppress phytopathogens, tolerate abiotic stress and lowers the metal toxicity to remediating plants through biosorption/bioaccumulation as bacterial cells have extremely high ratio of surface area to volume as well as promote plant growth by secreting various hormones, organic acids, and antibiotics (Rajkumar et al. 2012; Ullah et al. 2015). Endophytes are also able to tolerate high metals concentration and hence, lower phytotoxicity to remediating plants and helps in growth promotion by various means and thus, enhance phytoremediation efficiency (Ma et al. 2011, 2015). In addition, arbuscular mycorrhizal fungi (AMF, colonize plant roots) have been also reported to protect their host plants against heavy metal toxicity through their mobilization from soil, thus helping in the phytoremediation (Marques et al. 2009; Khan et al. 2014). A great deal of literature can be found in the public domain on the microbe-assisted phytoremediation of heavy metals (Khan et al. 2014; Ma et al. 2011, 2015; Rajkumar et al. 2012; Ullah et al. 2015). Further, to ameliorate metal toxicity, plant growth promotion, and metal sequestration, extensive research efforts are also required to explore novel microbial diversity and their distribution, as well as functions in the autochthonous and allochthonous soil habitats for microbeassisted phytoremediation of HM-contaminated sites. Several examples are existing on the phytoremediation and microbe-assisted phytoremediation of pollutants from industrial wastewaters. For instance, Gupta et al. (2018) studied the microbe-assisted phytoremediation of tannery wastewater (TWW) contaminated agricultural soils. They isolated a Cr6+-resistant plant growthpromoting Pseudomonas sp. (strain CPSB21) from the tannery effluent contaminated agricultural soils and evaluated for the various plant growth-promoting activities, oxidative stress tolerance, and Cr6+ bioremediation. Further, they applied the isolated strain for microbe-assisted phytoremediation and reported that the inoculation of strain CPSB21 alleviated the Cr6+ toxicity and enhanced the plant growth parameters and nutrient uptake in sunflower plant during pot experiment. Kassaye et al. (2017) reported the phytoremediation potential of swamp smartweed (Polygonum coccineum), Para grass (Brachiaria mutica), and papyrus (Cyperus papyrus) for Cr-containing TWW. They reported that all the three plants exhibited a significant transfer of Cr from wastewater (phytoextraction) to roots and shoots, but removal efficiency of Cr for swamp smartweed was relatively low as compared to Para grass and papyrus and further suggested the use of Para grass and papyrus for effective phytoremediation of TWW. Gregorio et al. (2015) reported the bacterialassisted phytoremediation of organic pollutants in TWW received from a conventional tannery wastewater treatment plant. They bioaugmented a plant growth-

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promoting rhizobacteria (PGPR) belonging to the Stenotrophomonas species in the rhizosphere of P. australis and reported 88%, 84%, and 71 % degradation of 4-nnonylphenol, mono-ethoxylated nonylphenol, and di-ethoxylated nonylphenol as compared to control (simple phyto-based approach).

2.3

Electrobioremediation

It is becoming an increasingly popular hybrid technology that uses the combination of bioremediation and electrokinetics for the treatment of environmental pollutants (Maszenan et al. 2011). It involves the electrokinetics phenomena for the acceleration and orientation of transport of environmental pollutants and microbes for pollutants bioremediation (Li et al. 2010; Maszenan et al. 2011). Electrokinetics involves the use of several phenomenon like diffusion, electrolysis, electroosmosis, electrophoresis, and electromigration and uses weak electric currents of about 0.2 to 2 V cm
1 (Saichek and Reddy 2005; Maszenan et al. 2011). A number of studies are available on the use of electrobioremediation technology for pollutants/contaminated soils (Wick et al. 2007; Martinez-Prado et al. 2014; Yan and Reible 2015). For instance, electro-biodegradation of toluene has been studied at a variety of anode potentials, both with pure cultures and consortia (Daghio et al. 2016; Lin et al. 2014; Zhang et al. 2010). Benzene was degraded in the anode of an electrobioremediation system using mixed cultures enriched from contaminated sediments (Zhang et al. 2010), wastewater (Wu et al. 2013), and anaerobic sludge (Adelaja et al. 2015). Polycyclic aromatic hydrocarbons’ (PAHs) degradation has also been reported in several studies (Adelaja et al. 2015; Yan et al. 2012). Phenol has been bioelectrochemically degraded both by mixed cultures and a pure culture of Cupriavidus basilensis (Friman et al. 2013; Huang et al. 2011). Furthermore, the dechlorination of 1,2-DCA was achieved in a electrobioremediation reactor inoculated with a mixed culture enriched in Dehalococcoides spp. (Leitão et al. 2015). In addition, the applications, potentials, and limitations of electrobioremediation technology have been reviewed by many authors (Wick et al. 2007; Maszenan et al. 2011; Gill et al. 2014).

2.4

Electrokinetic-Phytoremediation

Combining phytoremediation with electrokinetic remediation could be an excellent strategy to enhance metal mobility in contaminated soil and facilitate their plant uptake and, thus, phytoremediation (Saxena et al. 2019). For instance, Mao et al. (2016) evaluated the feasibility of electrokinetic remediation coupled with phytoremediation to remove Pb, As, and Cs from contaminated paddy soil. Results revealed that the solubility and bioavailability of Cs and As were significantly

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increased by the electrokinetic field (EKF) and thereby lower the pH of contaminated soil. Furthermore, they observed that EKF significantly enhanced the bioaccumulation of As and Cs in plant roots and shoots and, thus, enhanced phytoremediation efficiency. However, the optimization of electrical parameters such as electrical field intensity, current application mode, distance between the electrodes, stimulation period, and their effect on the mobility and bioavailability of HMs are the associated key challenges (Mao et al. 2016). Further, the application of electrokinetic-phytoremediation for the mixed contaminants (organic and inorganic) is also not reported so far.

2.5

Constructed Wetlands

Constructed wetlands (CWs) are the eco-technological option for the treatment and purification of HM-rich wastewaters. These are the man-engineered systems constructed to utilize the natural processes of aquatic macrophytes with their associated microbial assemblages for wastewater treatment within a more controlled environment (Stottmeister et al. 2003; Khan et al. 2009). CWs are mainly vegetated with different wetland plants with high biomass, fast growth rate, and metal accumulation capacity such as Phragmites australis, Typha latifolia, Canna indica, Stenotaphrum secundatum, Scirpus americanus, Scirpus acutus, Iris pseudacorus, etc. for metal-rich wastewater treatment (Bharagava et al. 2017c). CWs have been proven to be successful in the removal of a variety of organic and inorganic pollutants such as metals, nutrients, fecal indicator bacteria, and pathogens and a wide range of micro-pollutants, such as pharmaceutical and personal care products (Zhang et al. 2015). However, the pollutant removal efficiency of CWs mainly depends on wastewater treatment rate, organic loading rate, hydrologic regime, hydraulic retention time, operational mode, and vegetation type (Zhang et al. 2015). The application of CWs in pollutants’ removal from wastewaters has been recently reviewed by many workers (see Vymazal 2010; Zhang et al. 2015; Bharagava et al. 2017a, b, c). For instance, phytoremediation potential of Pennisetum purpureum, Brachiaria decumbens, and Phragmites australis in CWs has been reported for the phytoremediation (phytoextraction) of Cr from TWW (Mant et al. 2004). Calheiros et al. (2007) has been also reported the phytoremediation potentials of Canna indica, Typha latifolia, P. australis, Stenotaphrum secundatum and Iris pseudacorus in CWs for the treatment of TWW under two different hydraulic loading rates at 3 and 6 cm/day. It was found that only P. australis and T. latifolia were able to establish successfully. Further, they also evaluated Arundo donax and Sarcocornia fruticosa in two series of horizontal subsurface flow CWs that are used to treat TWW received from a conventional biological treatment plant and reported the removal of COD (51 and 80%) and BOD5 (53 and 90%) for COD inlet (68–425 mg L
1) and for BOD5 inlet (16–220 mg L
1) (Calheiros et al. 2012). CWs may provide many ecological and economic benefits such as require low capital investment for construction, low electricity for operation, and less

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maintenance and provide wildlife habitat, as well as human recreational opportunities and a reuse and recycling option for wastewater treatment facility. CWs are more favored in developing countries due to easily available and less costly land and tropical environment, which help to flourish the microbial communities responsible for the degradation/detoxification of organic and inorganic contaminants in wastewaters and therefore high treatment efficiency (Zhang et al. 2015). Thus, increasing use of CWs can successfully remediate heavy metal pollution and solve various water quality issues in the world. In addition, integrating CWs with a microbial fuel cell (MFC) for wastewater treatment and electricity generation could be an innovative approach for the improved degradation of pollutants. According to a recent study, a maximum power density of 15.73 mW m
2 and maximum current density of 69.75 mA m
2 could be achieved during the treatment of synthetic wastewater containing methylene blue dye (1000 mg l
1765) with 75% COD removal in an integrated CW-MFC system planted with an ornamental plant, Canna indica (Yadav et al. 2012). Moreover, CWs may have great potential for bioenergy production and carbon sequestration, if planted with energy crops. According to a study, the incineration of harvested biomass (16,737 kg with C content, 6185 kg) of Ludwigia sp. and Typha sp. recovered from a subtropical CW could produce 11,846 kWh for 1 month (Wang et al. 2011). However, the future research should be focused on (a) understanding of microbiological dynamics and correlation of biological and non-biological processes in CWs, (b) knowledge of element cycle dynamics that will help to understand the fundamental processes of greenhouse gas emission in CWs, and (c) understanding of microbial community and plant-microbe interactions to know the underlying mechanism of pollutant removal in CWs (Carvalho et al. 2017). Furthermore, researches are underway to expand the scope and efficacy of CWs for treatment of metal-contaminated wastewaters.

2.6

Microbial Fuel Cells

A microbial fuel cell (MFC) is a bioelectrochemical device that harnesses the power of respiring microbes to convert organic substrates directly into electrical energy. MFC can be a suitable alternative to the conventional activated sludge process based-treatment systems in terms of energy consumption and excess sludge generation. MFCs offer several advantages over conventional treatment systems related to energy (like direct electricity generation, energy savings by anaerobic treatment due to elimination of aeration, low sludge yield), environmental (water reclamation, low carbon footprint, less sludge generation), economic (revenue through energy and value-added products (chemicals), low operational costs), and operational benefits (self-generation of microorganisms, good resistance to environmental stress, and amenable to real-time monitoring and control) (Li et al. 2014; Gude 2016). MFCs are environmentally friendly technologies as they can produce clean electricity directly from organic matter in wastewater without any need for separation, purification, and conversion of the energy products and function at mild

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operating conditions especially at ambient temperatures (Gude 2016). MFCs can produce up to 1.43 kWh/m3 from a primary sludge or 1.8 kWh/m3 from a treated effluent (Ge et al. 2015). MFCs consume only 0.024 kW or 0.076 kWh/kg COD in average (mainly for feeding and mixing in the reactor), about one order of magnitude less than activated sludge-based aerobic processes (~0.3 kW or 0.6 kWh/kg COD) (Zhang et al. 2013a, b). It means MFCs consume only about 10% of the external energy for their operation when compared with conventional activated sludge process showing great potential for energy savings as well as possible energy recovery from wastewater treatment (Gude 2016).

2.7

Constructed Wetland-Microbial Fuel System

Integrating CWs with a microbial fuel cell (MFC) for wastewater treatment and electricity generation could be an innovative approach for the improved degradation of pollutants. The integration of MFC with a CW will upgrade the CW to allow it to be used for wastewater treatment and, simultaneously, electricity generation, making the CWs more sustainable and environmentally friendly. According to a recent study, a maximum power density of 15.73 mW m
2 and maximum current density of 69.75 mA m
2 could be achieved during the treatment of synthetic wastewater containing methylene blue dye (1000 mg l
1 765) with 75% COD removal in an integrated CW-MFC system planted with an ornamental plant, Canna indica (Yadav et al. 2012). Oon et al. (2015) designed an integrated upflow constructed wetlandmicrobial fuel cell (UFCW-MFC) system planted with cattail for the simultaneous wastewater treatment and electricity generation. According to their study, the removal efficiencies of COD, NO3
, and NH4+ were 100%, 40%, and 91%, respectively, with the maximum power density of 6.12 mW m
2 and coulombic efficiency of 8.6% at electrode spacing of anode 1 (A1) and cathode (15 cm). Fang et al. (2013) studied the performance of a microbial fuel cell coupled constructed wetland system for decolorization of azo dye and bioelectricity generation. They reported that the planted CW-MFC system achieved the highest decolorization rate of about 91.24% for azo dyes and a voltage output of about 610 mV.

2.8

Nano-bioremediation

Nano-bioremediation is the new concept that integrates the use of nanoparticles and bioremediation for sustainable remediation of environmental pollutants in contaminant matrix (Cecchin et al. 2017). For instance, Le et al. (2015) conducted a study for the degradation of a solution containing Aroclor 1248 (PCB) using nZVI (1000 mg/L) and subsequently using biodegradation with bacterium, Burkholderia xenovorans. The researchers obtained 89% degradation of the congeners after application of nZVI. Subsequently, they observed a biodegradation of 90% in the

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biphenyls produced after the dechlorination of PCB by bacterial metabolism. In this study, no toxic effect toward microorganisms by the nZVI was observed. Further, Bokare et al. (2012) conducted a study into the feasibility of an integration of bioremediation process and reductive process through nanoparticles in a contaminated solution with triclosan (5 g/L). The researchers promoted a sequential degradation of the contaminant by subjecting it to an anaerobic dechlorination through the nanoparticles of Pd/Fe. Subsequently, further remediation is achieved by oxidation of the by-products through the application of the enzyme produced by Trametes versicolor (laccase producing fungi). The results showed complete dechlorination of triclosan in 20 min after application and its by-products totally oxidized by microbial enzyme. Thus, nano-bioremediation could be an excellent strategy for the remediation of contaminants in environmental matrix.

3 Challenges and Future Prospects Bioremediation has emerged as a low-cost alternative to conventional remediation technologies, which are environmentally destructive and costly and create secondary pollution and, thus, negatively affect the ecosystem. However, it may get restricted by several factors such as low or non-bioavailability of pollutants to microbes, toxicity of pollutants to microbes and remediating plants, lack of enzymes responsible for the degradation and detoxification of specific environmental pollutants, recalcitrant nature of environmental pollutants, and low biomass, toxicity of nanoparticles to microbes as in the case of nano-bioremediation and slow growth rate of remediating plants as in the case of phytoremediation. Further, molecular techniques may advance the meaning of bio- and phytoremediation by developing transgenic microbes and plants for environmental remediation, but environmental risks such as invasion of exotic plants and loss of biodiversity associated with transgenic organisms make them less feasible for environmental decontamination. Moreover, the strict US and Western countries’ regulations on the use of these organisms also restrict their filed applications. These limitations are sufficient to discredit the applicability of bioremediation technologies and together constitute a major challenge in the way of success at field scale. Moreover, several emerging and ecofriendly approaches can be suitable alternative for the conventional bioremediation technologies. Electrobioremediation, electrokinetic phytoremediation, nanobioremediation, constructed wetlands, and microbial fuel cell technology represent a highly promising and sustainable future in waste treatment and management. However, these have some serious challenges that need to be catch-up for wide applications in a more sustainable and economic manner. Further, future research efforts may provide new ways to make the bioremediation technologies more efficient for environmental remediation.

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4 Conclusion Industrial wastewater is a major source of pollution and toxicity in the environment, and bioremediation is an ecofriendly option to treat and manage such hazardous waste. To expand the scope and efficacy of bioremediation, the future research should be focused on (a) search for potential microbial degraders for environmental pollutants; (b) search for catabolic enzymes or genes for the enhanced degradation/ detoxification of environmental pollutants; (c) development of transgenic microbes and designer plants using genetic engineering for effective bio- and phytoremediation; (d) selection of suitable plants for phytoremediation; (e) search for novel rhizobacteria and endophytes for microbe-assisted phytoremediation; (f) optimization of electrical parameters such as electrical field intensity, current application mode, distance between the electrodes, stimulation period, and their effect on the mobility and bioavailability of HMs in electrokinetic phytoremediation; and (g) understanding the complex microbiology of constructed wetlands for mechanistic view of pollutant removal/wastewater treatment. However, continued efforts are required to realize the economic feasibility of bioremediation technologies including phytoremediation at the field. Acknowledgment The financial support provided by the University Grant Commission (UGC) to Mr. Gaurav Saxena is duly acknowledged. The corresponding author (Dr. Ram Naresh Bharagava) is also highly thankful to the “Science and Engineering Research Board” (SERB), Department of Science & Technology (DST), Government of India (GOI), New Delhi, India, for providing the financial support as “Major Research Project” (Grant No.: EEQ/2017/000407), which is also duly acknowledged.

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Chapter 6

Constructed Wetlands: A Clean-Green Technology for Degradation and Detoxification of Industrial Wastewaters Sardar Khan, Javed Nawab, and Muhammad Waqas

Abstract Constructed wetlands (CWs) have played a significant role in the purification and treatment of domestic, mining, agricultural, and industrial wastewater in the last few decades. CWs are designed and constructed on engineered systems to develop the natural processes involving wetland soils, flora, and their related microbial accumulations to support wastewater treatment. The CWs, therefore, present environmentally friendly, cost-effective, and favorable substitute for industrial wastewater treatment. Several techniques have been used in the removal of contaminants from CWs such as filtration, sedimentation, adsorption, volatilization, phytoaccumulation, and microbial activity. In the past, CWs have played efficient role in the removal of toxic metals, hydrocarbons, pharmaceuticals, and dyes from wastewater. However, the efficiency mainly depends on initial concentrations of contaminants, plant types, plant microbes’ interactions, climatic condition and flow rate of wastewater etc. The overall conclusion of this book chapter will contribute to the development of phyto-technology for industrial wastewater and other associated industrial problems. Keywords Constructed wetlands · Phyto-technology · Wastewaters · Contaminants · Toxicity

S. Khan (*) Department of Environmental Science, University of Peshawar, Peshawar, Pakistan J. Nawab Department of Environmental Sciences, Abdul Wali Khan University, Mardan, Pakistan Department of Environmental and Conservation Sciences, University of Swat, Mingora, Pakistan M. Waqas Department of Environmental and Conservation Sciences, University of Swat, Mingora, Pakistan © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_6

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1 Introduction A wetland is an ecological unit that rest on constant or recurrent, superficial capacity at or nearby the exterior of the substrate. The least vital features of a wetland are persistent, continued inundation or capacity at or nearby the surface and the existence of biological, physical, and chemical structures reflective of persistent, constant inundation or saturation. The most common analytical features of wetlands are hydrophytic vegetation and hydric soils. Most commonly wetlands are basically divided into two types: natural and constructed wetlands. Wetland study generally focuses on environmental engineering and ecosystem renovation (Mitsch and Gosselink 2007; Mitsch and Jorgensen 2003). In natural wetlands, the water asylums the soil, comprising marshes, swamps, sloughs, bogs and fens. Constructed wetlands are related to natural wetlands, linking physical, biological, and chemical procedures in ecosystem (Mitsch and Gosselink 2007; USEPA 2000). Constructed wetlands have been practiced to treat several types of wastewaters such as municipal and industrial wastewaters, acid mine drainage, and agricultural and urban runoff because of its sustainability and efficiency (Saxena et al. 2019; Gautam et al. 2017; Bharagava et al. 2017a, b; Saxena et al. 2016; Saxena and Bharagava 2015, 2017; Chandra et al. 2015; Mitsch et al. 2008; Scholz et al. 2007; Zhang et al. 2008; USEPA 2000; Nyquist and Greger 2009; Weber et al. 2008). Wetlands comprise of soil, water, flora, and microbe systems and play a vital role for retaining marine ecosystem biodiversity (Mitsch and Gosselink 2007).

2 Types and Characteristics of Wetlands Wetlands have diverse features. The wetland key features are determined by the soil type, water salinity in the wetland, and the animals and plants living in the wetland. In all the wetlands, the most common characteristic is that the water table is very close to the soil surface, or shallow water asylums the surface for minimum period of the year. There are two main types of wetlands: (1) natural wetlands and (2) constructed wetlands.

2.1

Natural Wetlands

Around the world, a number of different types of wetlands are present. They are generally divided into four distinct groups. Marshes The wetlands are those that are constantly flooded relatively than being inundated beneath water just during the summer or a couple of months over the year, for instance. Marshes can be saltwater or freshwater, and the quantity of water in

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marsh can be changed with different seasons. Marshes boast an excessive variety of flora that has modified particularly to live in saturated soil. The subcategories of marshes include saltwater, freshwater, inland, and coastal. Each of these has its own diverse ecosystems and can be found all around the globe. Marshes wildlife includes alligators, newts, beavers, turtles, and shrimps. Swamps The main difference between swamps and marshes is that, normally, swamps are dominated by woody plants (rather than soft-stemmed plants). Majority of the people often harvest these trees to make timber and to build their homes, which can disturb the environment significantly if too many are taken deprived of being replaced with new trees. Swamps provide home to different fishes and birds as well as smaller creatures. In Florida, swamps like Everglades are found in low-lying areas near coastal or river areas. Swamps can be divided into two main types (shrub swamps and forested swamps). Swamps are home for diversity of animals like bobcat, alligators, snakes, beaver, and large variety of birds. Bogs Bogs are categorized by spongy peat deposits and more acidic water as well as layering of sphagnum moss. Contrasting swamps and marshes, bogs are likely to get their moisture from rainfall rather than watercourses such as rivulets, streams, or excesses from rivers. These wetlands play important role for inhibiting downstream flooding since they absorb rainfall as it falls and prevents the inflammation of rivers and other watercourses. Fens Fens are peat-forming wetlands like bogs, even though they generally get their wetness from ground and surface water rather than rainfall, which mean that they are somewhat least acidic. This means that fens support a larger group of nature, from plants to birds to fish and everything in between. Like bogs, fens are valuable because they can help in mitigating the flooding of land, since they saturate up water from the ground and stop it from leaching everywhere else.

2.2

Constructed Wetlands

Humans designed engineered ecosystem which is termed as constructed wetland to remove contaminants from wastewater treatment that can mimic the habitat and hydraulic condition occurring in the swamp. The artificially constructed wetland system resembles the treatment that occurs in natural wetland by relying on a combination of naturally occurring physical, chemical, and biological process and on the aquatic plants and heterotrophic microorganisms (Olejnik and Wojciechowski 2012). For wastewater treatment, constructed wetlands have been used for nearly 60 years; the first CWs were put in operation in Israel in the 1950s. In Europe, treatment of wastewater by wetland vegetation was started in Germany. For the treatment of municipal or domestic sewage, wetlands have been used for decades. In the last several years, these systems have been constructed to treat the wastewaters

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originated from different sources for quality improvement. Various types of wastewater are treated by using CWs, i.e., agricultural wastewaters, landfill leachate, urban stormwater, and industrial wastewater including paper and pulp, domestic wastewater, acid mine drainage, food processing, petrochemical, chemical, textile, and tannery. The mechanisms of wastewater treatment in wetlands are extremely complicated, including a series of physical, chemical, and biochemical processes. The wetland efficiency to remove the pollutants from the wastewater primarily depends on the root zone relations between soil, pollutants, plant roots, and diversity of microorganisms (Olejnik and Wojciechowski 2012) (Fig. 6.1). The simple idea behind constructed wetlands is their ability to absorb, filter, and metabolize dissolved and suspended matter. In order to mimic these natural structures, engineers and scientists are very interested to work for handling wastewaters (UNEP 2004). This has encouraged the production of synthetic wetland’s construction and has prompted in order to handle with the diffuse contamination initiating from septic tanks, agriculture, and other causes. The reason for the constructions of artificial wetlands is its unique and cost-effective phyto-technology concept in order to control various types of wastewater (UNEP 2004). Constructed wetlands are commonly intended in such a way that water predominantly runs above the sediment and through hydrophytes, or the design of stagnated waterlogged bed systems in which water moves through plants for contact with plant roots or either planted or vegetated in a natural way. On the basis of hydrologic consequences in the sink, constructed wetlands are categorized into three classes: subsurface flow, free surface flow, and hybrid types (Ulsido 2014).

Constructed Wetlands

Surface flow wetlands

Horizontal flow system

Fig. 6.1 Types of constructed wetlands

Subsurface wetlands

Hydrid wetlands

Vertical flow system

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Surface Flow Constructed Wetland

Free surface flow wetlands are considered as artificial shallow ponds or shallow marshy lands occupied by aquatic plants or hydrophytes (Wetlands International 2013). They are very shallow excavations or shallow earth-banked lagoons enclosing an area of land demarcated for the purpose. The name of the free surface wetland is related to the water layers which are thin free layers that are formed at the surface in free surface wetland; for the marshy plants, the growing media are provided by soil or some other media such as gravel. The surface should be virtually flat to avoid short-circuiting with a very gentle slope toward the outlet end (Wetlands International 2013). The wastewater is distributed over the inlet end of the wetland through inlet zone, and the treated liquid is collected by a collection channel at the outlet end. The wastewater in free surface wetland flows along the surface along with settlement of solids and then come into contact with the population of bacteria on the surface of the plant stem and media. These types of wetlands can be used as substitute for the treatment systems for both smaller and larger communities. Furthermore, it can also be considered to treat point sources and nonpoint sources of contamination and particularly well appropriate for surface-mined areas (UNEP 2004). Its additional ecological advantages are the amount of nutrients; water, nitrogen, and saltwater filtering; support of threatened species; and production of food that can possibly increase the cost-effective advantages in contrast to traditional wastewater treatment plants (Ulsido 2014). In constructed wetlands, phytoremediation technique is effective but mainly depends on hydrology, site selected geology, air quality, characteristics of aquifers, conditions of soil, geochemistry, climatic conditions, distribution and types of microorganisms, and distribution and presence of vegetation and contaminants (UNEP 2004) (Fig. 6.2).

Fig. 6.2 Surface flow constructed wetlands

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Subsurface Flow Constructed Wetlands

Subsurface flow wetlands are narrow pits (1–1.5 m deep) synthetic liner or with clay in order to stop water penetration. They are packed with a media through which the liquid must flow and purify under the surface of the ground. Any type of media can be used from soil to light, expanded clay aggregate, but the most common is 5–10 mm. There is an inlet zone which is made from soil aggregates that is larger in size to ensure the effective distribution of influent liquid into the media. Another similar outlet zone in drainage pipes which pass through the liner into a level control chamber is used to collect treated liquid where a swivel pipe or simple plastic tube is used to allow the level of liquid in the wetland to be controlled (Fig. 6.3).

2.2.3

Hybrid Constructed Wetlands

To accomplish greater removal effectiveness, mainly for nitrogen, diverse types may be combined in constructed wetlands. The design mainly comprised of two phases, numerous similar vertical flow (VF) beds followed by 2 or 3 horizontal flow (HF) beds in sequence vertical subsurface flow and horizontal subsurface flow system (VSSF-HSSF system). The VSSF wetland eradicates suspended solids and organics to encourage nitrification, while in HSSF wetland, additional removal of organics, denitrification, and suspended solids take place. Additional structure is a HSSF-VSSF system. The huge HSSF bed is retained first to remove suspended solids and organics and to encourage denitrification. An alternatingly loaded small VF bed is used for suspended solids, nitrification, and extra removal of organics. To take full advantage for the total nitrogen removal though, the nitrified waste from the VF bed must be reprocessed to the sedimentation tank. In broad terms, constructed

Fig. 6.3 Subsurface flow constructed wetland

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wetlands could be joined to attain higher treatment results, but the HSSF-VSSF and VSSF-HSSF constructed wetlands are the best public hybrid systems (Vymazal et al. 2007).

2.3

Application of Constructed Wetlands

Constructed wetlands (CWs) are solid biological methods used for different types of wastewater treatment (Brix 1997). CWs have been used successfully for the management of general sewage, tannery wastewater, piggery manure, and fishpond water. Constructed wetlands are mostly used for municipal wastewater and industrial wastewater. The industrial wastewater needs to be analyzing carefully industrial effluents because the water composition is highly variable and toxic. The use of constructed wetlands for industrial wastewater has improved over the past decade (Korkusuz 2005; Kadlec et al. 2000). Constructed wetlands are nonnatural or man-made wetlands constructed for treating wastewater of industries, municipalities, and stormwater runoff. Plants, microbes, sunlight, and gravity are used to transform wastewater into reusable water by effective design of constructive wetlands. Water is treated through physical, biological, chemical mechanisms by means of filtering, absorption, and decomposition. Sewage water can be converted into reusable water by constructed wetlands and built green belts around the communities; there is no machinery, but all the processes are run naturally, and if there is limited space for constructed wetlands, then the new wetlands are integrated with old ones so this will be more efficient and inexpensive (Wastewater Gardens International). Constructed wetlands are not a new concept; it has been used from the last two decades for treating the contaminated water and wastewater (Murray-Gulde et al. 2005; Maine et al. 2006; Zhang et al. 2014). The natural wetland concept is too much old; it has been used from millions of years. Natural wetlands or volunteer wetlands are mostly associated with mining (Beining and Otte 1996, 1997). Constructed wetlands in the past have been used for different purposes such as rehabilitating areas where wetlands were formerly located to treat wastewater (Hawkins et al. 1997). In recent years, concern has been focused on the removal of toxic metals from wastewater (Horne 2000). Mostly the substrate interactions eliminate maximum toxic metals from polluted water in constructed wetlands (Walker and Hurl 2002). The stable or temporarily toxic situation in wetland soil helps to generate an atmosphere for the restriction of toxic metals in the extremely concentrated sulfite or metallic form (Gambrell 1994). Constructed wetlands are effective, cost-efficient, and environmentally sound approach for treating polluted water (Brix 1997; Stottmeister et al. 2003). Arias and Brix (2003) classify constructed wetlands on the base of dominant plant species: macrophytes floating leaf (e.g., Nymphaea alba, Potamogeton gramineus); macrophytes submersed (e.g., Potamogeton crispus, Littorella uniflora); macrophytes floating (e.g., Lemna minor, Eichhornia crassipes); and macrophytes emerged rooted (e.g., Phragmites australis, Typha latifolia).

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3 Industrial Wastewater and Different Pollution Loads Industrial wastewater can create environmental problems due to high chemical oxygen demand level (Goutam et al. 2018; Bharagava et al. 2017c; Song et al. 2000). The different types of pollution loads such as domestic, agricultural, and industrial wastes directly and indirectly in the form of toxic liquids are adding to the aquatic ecosystem (Demirezen et al. 2007). Pollutants mostly accumulate in surface water, groundwater, and plants in the form of nutrients, organic contaminants, aromatic hydrocarbons, biphenyls, pesticides, xenobiotic, metalloids, and metals such as Cu, As, Fe, Zn, Ni, Cd, Hg, Pb, Sr, Cr, Al, Se, and Ba (Aksoy et al. 2005). The most common pollution sources of Cd are plastics, metal industry, and fossil fuel combustion, and home tools are considered as harmful pollutants for animals and humans. Pb is also in toy production, vehicle emission, printing, petroleum industry, exhaust gases, and wastewater (Demirezen et al. 2007). Any form of pollution which immediate source to industries is known as industrial pollution. The linkage of pollution with industries has some kind of historical background. The issue of industrial pollution has great importance and concern recently for fighting against environmental degradation by agencies. Sudden and rapid growth of industries is a serious issue locally, nationally, and globally to be brought under control immediately. Industrial revolution brings advancement, lifestyle becomes better, technology developed, facilities added in people’s lives, and manufacturing age came, but with all these, some negative impact occur on nature and humans, that is, industrial pollution. In the past there were limited factories that cause air pollution, but with developing more industries, the flow of pollution broadens in all means either its amount, severity, or toxicity and affects all parts of the environment. Moreover, when the number of factories and working hours were limited the level of pollution was not that much high, but with the passage of time, these factories converted into large-scale industries and manufacturing units; the pollution level also increased with these advancements, and this leads to take on importance to the issue of industrial pollution. Petroleum, leather, and agro food industries produce extremely saline wastewaters due to their dynamic cycle. In the leather sector, certain sewage streams may comprise 80 g L1 of NaCl. The high amount of salts present in wastewaters may badly disturb the natural processes in the treatment of wastewater systems since hypersaline wastes are often uncontrollable (Lefebvre and Moletta 2006). The improvement of wastewater treatment in industrial technologies will be gradually reduced because of technical process. In developing countries, the industrial wastewater is much higher than those in developed countries. The release of huge quantities of toxic metals such as Cd, Cr, Cd, Cu, As, Ni, Zn, and Pb in industrial wastewater are the most dangerous among the chemical-intensive industries. Industrial pollution pollutes many sources of drinking water and releases undesirable toxins into the environment all over the globe. Industrialization has caused major environmental disasters in the past. Underneath are rare causes of industrial pollution that have resulted environmental degradation (Daniels 2001). Emissions of

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tannery industry have experienced regular problems because the design is poor and has less capacity, but for the betterment of production, they are working over the capacity. In this situation, constructed wetlands are good choices to enhance the natural presentation.

4 Removal of Toxic Elements from Industrial Wastewater in a Constructed Wetland Constructed wetlands are mostly used for treatment of domestic wastewater, but due to high pollution load from industries, the use of constructed wetlands for the treatment of industrial discharge has increased over the past 10 years (Korkusuz 2005). The removal mechanism for toxic metals is summarized in Table 6.1. Metal removal processes in wetlands: according to Lesage et al. (2007), four methods are highlighted for the effective metal removal in wetlands: (1) adsorption to organic matter and fine-textured sediments; (2) precipitation as unsolvable salts (mostly oxyhydroxides and sulfides); (c) concentration and induced variations in biogeochemical cycles by bacteria and plants; and (d) suspended solid depositions due to low flow rates. In the substrate of the wetlands, all the above processes lead to metal accumulation. Adsorption: The exchange of ions from liquid to solid phase is an essential mechanism for metal removal in wetlands. Sorption defines a group of methods which comprises chemical procedures with strong bindings and physical procedures with weak bindings; absorption is a biochemical process when a compound from the external media is entering into animals or plants and precipitation reactions take place. Metals are absorbed by ion exchange method or chemo adsorption method (Seo et al. 2008). Another useful parameter to quantify adsorption capacity of a material for an ion is the distribution coefficient Kd (Alloway 1995). Kd ¼ Equilibrium concentration of heavy metals adsorbed Equilibrium concentration of heavy metals in solution Coprecipitation and redox reactions: Some metals, e.g., Fe, Al, and Mn, can form insoluble compounds through hydrolysis and oxidation and form different oxides Table 6.1 Mechanisms of toxic metals removal. (adapted from Vymazal et al. 2007; Mbuligwe 2005) Toxic heavy metals As, Pb, Cd, Cr, Cu, Mn, Fe, Ni and Zn etc.

Biological processes Bioaccumulation/biotransformation by microbes and plants, photodegradation, photovolatilization and evapotranspiration

Chemical processes Ion exchange; precipitation and adsorption of toxic metals

Physical processes Settling of heavy metals into sediments

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(Sheoran and Sheoran 2006). The amounts and forms of Fe in solution strongly affect removal of metal. Fe (II) is soluble and characterizes an important bioavailable fraction. Under aerobic condition, it can be oxidized to Fe (III) in conjunction with Hþ ion consumption. Fe (III) can precipitate to produce oxides, hydroxides, and oxyhydroxides (Jonsson and Lovgren 2006). Fe (II) can coprecipitate with other metals such as Cd, Ni, Zn, and Cu (Matagi 1998). Iron oxides have a particularly strong affinity for cations with a similar size compared to Fe (III) and Fe (II), e.g., Cd, Zn, Ni, and Cu (Dorman et al. 2009). The other metalloids like arsenic (As) can be removed by adsorbing against amorphous iron hydroxides or by coprecipitating with iron oxyhydroxides in water column (Manning et al. 1998). Metals can also form insoluble compounds through reduction. Under chemically reducing conditions (Eh < 50 mV), sulfates can be reduced to sulfides. These can combine with various elements, i.e., As, Hg, Se, and Zn, to coprecipitate in relatively insoluble forms (Murray-Gulde et al. 2005). A constructed wetland based on a matrix with exclusively reducing conditions, however, cannot be efficient. These conditions promote huge ion discharge, mainly of Mn and Fe, into the water by reduction of the oxides and oxyhydroxides confined in the substrate (Goulet and Pick 2001). Metal Carbonates Metal carbonates can be formed from metals though; carbonates are not as much constant than sulfides and play a vital role in initial tricking of metals (Sheoran and Sheoran 2006). For the removal of Pb and Ni carbonates, precipitation is generally effective. According to Maine et al. (2006), the external wastewater structure having high carbonates, pH, and calcium concentrations preferred the metal retaining in the sediment. Metals are removed and adsorbed to carbonates in wastewater. pH The pH plays a vital role in affecting the effectiveness of metal removal in wetlands. Proton production takes place due to the conversion of ammonium into nitrites during nitrification. In promoting the nitrification process, these ions of hydrogen are normalized by the ions of bicarbonates because oxygen is released by the macrophytes. The pH decreased because the protons produced due to nitrification may not be able to neutralize by HCO3 ions (Lee and Scholz 2007). To promote adsorption and removal of oxyanions, for example, As, Sb, and Se, iron coprecipitation must occur under acidic conditions (Sheoran and Sheoran 2006). Conversely, alkaline environments are essential to encourage coprecipitation of cationic metals, such as Cd, Cu, Ni, and Zn. The constructed wetland efficiency can be reduced by metal removal as a result of high nitrification rate (Lee and Scholz 2007).

4.1

Plant’s Role in Removal of Toxic Elements from Industrial Wastewater

The option of using plants in constructed wetlands is an important concern because the plants need to survive in possible lethal effects of wastewater and against its

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inconsistency. In Europe, the usually constructed wetlands design horizontal flow system is used in which common reed (Phragmites australis) is grown (Vymazal et al. 1998). Other extensively important plants for wastewater treatment in constructed wetland are shown in Table 6.1 (Vymazal et al. 2007; Mbuligwe 2005). The main macrophyte species used in constructed wetlands in Portugal are P. australis, Cyperus spp., Juncus effusus (soft rush), Iris pseudacorus (yellow iris), Juncus spp., and Scirpus spp. as shown in Table 6.2. In Portugal, a study was conducted on plant’s use in constructed wetlands; the plant material used in pilot units was selected on the base of plants species growing in surrounding of wastewater discharge tank, i.e., Typha latifolia, Iris pseudacorus, Canna indica, and Stenotaphrum secundatum as shown in Table 6.2. Some of the above plants were transplanted in pilot units from industrial polluted site Estarreja Portugal (Korkusuz 2005). The common reed (Phragmites australis) is frequently used in CWs because of its abundance in Portugal. Macrophytes produce organic matter, uptaking pollutants, bioengineering of rhizosphere, and maintenance of habitat for microorganism. Macrophytes have the ability to accumulate one or several metals due its phenotype property (Marchand et al. 2010; Kamal et al. 2004). Some plants species can uptake heavy metals 100,000 times more than in the associated water (Mishra and Tripathi 2008). Hyperaccumulators can uptake and translocate high levels of certain metals that would be toxic to most organisms, but study on hyperaccumulation is mostly done on dryland plants (Marchand et al. 2010). Mainly macrophytes produce organic Table 6.2 English and scientific names of different plants used in constructed wetlands English name Southern cattails Cattails Bulrushes Yellow flag/yellow iris Canna Augustine grass Soft rush Asia crabgrass Sedges Para grass Bunchgrass Cordgrass Creeping burhead Duck weed Lesser duckweed Willow Common reed Tall reed Reed canary grass

Scientific name Typha domingensis Typha latifolia Scirpus spp. Iris pseudacorus Canna indica Stenotaphrum secundatum Juncus effusus spp. Digitaria bicornis Cyperus corymbosus Brachiaria mutica Vetiveria zizanioides Spartina patens Echinodorus cordifolius Lemna gibba L. Lemna minor Salix alba L. Phragmites australis Phragmites karka Phalaris arundinacea

Family Typhaceae Typhaceae Cyperaceae Iridaceae Cannaceae Poaceae Juncaceae Poaceae Cyperaceae Poaceae Poaceae Poaceae Alismataceae Lemnaceae Lemnaceae Salicaceae Poaceae Poaceae Poaceae

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matter needed to run the biogeochemical processes and ensure the availability of organic compound through exudation from roots (Jenssen et al. 1993). In acid mine drainage, the substrate and water are categorized by low pH and high metal concentrations. This leads to release of dissolved iron and protons, which, in turn, leads to the release of other metallic ions, such as Mn, Ni, Zn, Cu, and Cd. Because of the extreme conditions of acid mine drainage (Holmstrom 2000), the formation of acid mine drainage can be prevented by limiting contact between mining wastes and oxygen. One attractive and efficient solution for reducing O2 diffusion is to construct a wetland as a cover over mine waste (Stoltz and Greger 2005). Erosion and Sedimentation The macrophyte plant species (Phragmites australis) encourage suspended solids and sedimentation and also stop erosion by reducing flow rates of water by increasing the surface area per length of the hydraulic pathways through the structure (Lee and Scholz 2007). Surface flow system shows different behaviors; sometimes, it shows static behavior in which there is no flow and another is dynamic behavior in which the flow is high in static condition, and the flow is negligible and behaves like stagnant pond in which displacement effects caused by submerged plant mass decrease retention times. Under dynamic conditions, active flow-through and stem drag are increased and are more important than displacement of volume (Matagi et al. 1998). Retention times become high with increasing vegetation resulting in increased sedimentation, but it is possible after floc formation, but it may absorb metal and other types of suspended material. Flocculation can be improved by high pH, high ionic strength, high vegetation density, and high concentration of suspended material. A key point that is often overlooked is the supply of organic matter by plants. Organic matter derived from plants in wetlands constantly provides carbon sources for bacterial metabolism as well as metal sorption thus encouraging long-term operation (Marchand et al. 2010). Plant plays an important role in wetlands for wastewater treatment. Though, handling with contaminated saline wastewater importance must be given to the selection process of plants as they should be tolerable to salt and show resistant in active CW system (Brix 1997). Klomjek and Nitisoravut (2005) assessed the feasibility of CWs to remove contaminants through eight plant species from saline wastewaters. They stated that Typha angustifolia is best for nitrogen accumulation and plant growth, whereas Digitaria bicornis was best for BOD5 removal when linked to other species (Brachiaria mutica, Cyperus corymbosus, Vetiveria zizanioides, Leptochloa fusca, Spartina patens, and Echinodorus cordifolius). Emergent Plants Various types of emergent plants are used in constructed wetlands and result in different metal removals of widely used species Phragmites australis (Vymazal et al. 2007). Other used species are Typha domingensis, Typha latifolia, and Phragmites karka (Juwarker et al. 1995). In constructed wetlands, numerous emergent plants particularly Phragmites australis have been tested achieving variable metal removal rates (Vymazal et al. 2007). Phalaris arundinacea shows same capabilities like Typha domingensis, Phragmites australis, Phragmites karka, and Typha latifolia (Juwarker et al. 1995) as shown in Table 6.2. Toxic metals and suspended organic matter are removed mainly by rhizosphere immobilization

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and storage in the roots by plant species (Lesage et al. 2007). Research work has shown that heavy metals such as Cr, Zn, and Ni are efficiently accumulated in the whole plants (Cheng et al. 2002). The leading capacity of root-exuded organic acids arises in soil as anions such as oxalate, acetate, malate, citrate, fumarate, and malonate. The phytotoxicity can be decreased because the anions can chelate metallic ions to variable degrees (Ryan et al. 2001). Floating Plants Floating plants absorb metals, i.e., Eichhornia crassipes, Salvinia herzogii, and Pistia stratiotes. In contrast, floating plants store metals in their biomass and then absorb it in to subtract (Vymazal et al. 1998). Eichhornia crassipes biomass doubles in 6 days under promising situations (Mitchell 1976). Some of the plant takes up high amounts of P and N in the roots. This supports microorganism that degrades organic matter and releases oxygen into the E. crassipes that permits a significant P removal rate in a short duration after harvesting from water (Maine et al. 2006). Submerged Plants The submerged marine plants such as Hydrilla verticillata, Ceratophyllum demersum, Potamogeton spp., and Myriophyllum spicatum have been used frequently for wastewater treatment (Bunluesin et al. 2007). Besides the experiments performed, batch experiment studies have been useful, but their implementation in large-scale constructed wetlands is ambiguous, mainly due to their little winter performance, biomass and, maintenance of efficient system (Kivaisi 2001). To stabilize acid mine drainage, plantation of submerged plant species is preferred in wastewater because these plants accumulate more metals as compared to emergent macrophytes (Nyquist and Greger 2009). Submerged macrophytes are possibly not appropriate for wastewater in excessive Fe precipitation, because it constrains light distribution and photosynthesis (Nyquist and Greger 2009). The system in which the contaminants entered the plants and transpired through the plant leaves is called phyto-volatilization (Interstate 2003). Besides this, plants play a significant role in removal of heavy metals through adsorption, filtration, and cation. Plant species play a vital role in uptake and removal of heavy metals, and some plant species have the capability to accumulate more heavy metals than others, such as Lemna gibba, Salix Alba L, Lemna minor, Phragmites australis, and Typha latifolia (Zayed et al. 1998). According to Stottmeister et al. (2003), the extreme circumstances to treat wastewater through rhizosphere in wetlands can be summarized as (1) extremely condensed situation (Eh up to 90%) in the field application of the EK process using a voltage gradient of 0.8 V/cm for 60 days, but they could not reduce the electrical conductivity of the soil to the recommended value of 2.5 dS/m for crop cultivation. The removal efficiency was about 80% and 60% at the top and bottom layers of the soil, respectively. Kim et al. (2011) observed that the nitrate removal was higher than chloride concentration for 14 days of ex situ EK processing in the hexagonal electrode arrangement system

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for removal of salts in field soil. They concluded that overall removal efficiency was not sufficiently higher because of the short operation time. Most of the investigators (Cho et al. 2011; Jo et al. 2015) experienced a problem with the removal of sulfate in the contaminated soil that directly affects electrical conductivity and reduces the efficiency of the EK process. Sulfate is highly reactive with available metal ions such as calcium, aluminum, iron, or magnesium ions present in the contaminated soil. However, chloride and nitrate are easily removed from the soil. Annamalai et al. (2015) successfully achieved more than 96% of sulfate removal from a real contaminated soil collected from Tirupur, India, and concluded that the pH of the electrolyte is vital in the removal of sulfate in actual contaminated soil. The acidic pH of the electrolyte (0.01 M HCl) provoked the ionization of the sulfate ion, which can be easily removed from the soil.

2.4

Electrokinetic Removal of Organic Dye Compounds

Ricart et al. (2008) initiated dye removal from an artificial soil, and recently Annamalai et al. (2014a, b, 2016) tried to remove textile reactive dyes from a real contaminated soil. It was established that the EK process is a useful technique for elimination of these organic contaminants from a solid matrix. Several investigators (Cameselle et al. 2013; Maini et al. 2000; Maturi and Reddy 2006; Ribeiro et al. 2005; Ricart et al. 2008; Wang et al. 2007) tried to remove uncharged organic pollutants [PAH, trichlorobenzene (TCB)] from a model soil. The persistent organic molecules, which are not soluble in water, became stuck in the voids of the porous material and did not form ionized or ionic molecules. These organic molecules were removed by electro-osmosis, the transport of organic pollutant along with pore fluid water, but they cannot be removed by electromigration. The textile reactive dye components are easily adsorbed on the surface of the soil, which can affect soil quality. Generally, triazine and vinyl sulfone groups are the most important of the reactive dyes (Carmen and Daniela 2010). The reactive dyes contain at least one vinyl sulfone moiety (–SO2–CH¼CH2) or a 2-sulfatoethylsulfone moiety (–SO2– CH2–CH2–OSO2 Na), which can be hydrolyzed into the vinyl sulfone moiety. The reactive blue 19 (RB 19), copper phthalocyanine (Cu-Phy), reactive black 5 (RB 5), etc., are important industrial dyes that are used frequently. RB 19, Cu-Phy, and RB 5 dyes are based on anthraquinone, porphyrin, and naphthalene group dyes, respectively, which are more resistant to biodegradation because of the fused aromatic structures compared with azo-based reactive dyes (Pazos et al. 2008; Rajkumar et al. 2007). Pazos et al. (2008) explained the enhanced EK remediation of RB5 in kaolinite-polluted model soil. When they tried to remove RB5 from the polluted matrix, initially RB5 was removed by EK treatment, which was collected to the anolyte chamber and simultaneously oxidized by an electrochemical process. In the EK process, the electrolyte solution is the key parameter for enhancing the conductivity of the matrix and it favors the desorption of the pollutants from the surface of the soil matrix. Without a pH adjustment in the EK process, removal efficiency was

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reduced by the acidic and basic nature in electrolyte compartments that hinder the mobility of the pollutant ions. They concluded that the pH value of the soil matrix is pivotal to achieving complete remediation and that K2SO4 favored high electrical conductivity. Desorption of RB5 in the model soil matrix and decolorization were achieved in the electrolyte compartment. Ricart et al. (2008) efficiently removed both RB 5 and Cr (III) ion from a model kaolinite soil using EK remediation. The four sulfonic groups were neutralized and formed four negative charge anions at alkaline pH. The negative charge of the RB5 moved towards the anode by the electromigration process. It was found that Cr was transported towards the cathode by electromigration and electro-osmosis. The interaction among RB5 and Cr in the kaolinite sample prevented premature precipitation and allowed Cr to migrate and concentrate in the cathode chamber. Normally, reactive dyes are water soluble and therefore are functionalized with the sodium salt of sulfonate ion. In the alkaline condition, reactive dyes are ionized into weak acids that easily favor the mobility of ions by the electromigration process. Ammonium acetate is an electrolyte that reacts with metal ions, forming metal acetate. The metal acetate is easily ionized and moves towards the opposite direction. Similarly, Annamalai et al. (2016) conducted a bench-scale test to evaluate the EK process for in situ formation of electroactive species (OH˙) at the anode, which enhanced the organic degradation (68%) in a real contaminated soil. Hence, a suitable technology is needed to improve efficiency for the removal of dye components in the agricultural soil.

3 Bioremediation of Polluted Soils Many research articles have addressed the treatment of organic pollutants in soil and sediment by using autochthonous microorganisms such as Ochrobactrum intermedium (Khan et al. 2014), Saccharomyces cerevisiae (Jadhav and Govindwar 2006), Kluyveromyces marxianus (Bustard et al. 1998), Pseudomonas aeruginosa (Bhatt et al. 2005), and Sphingomona sp. strain BN6 (Keck et al. 1997). Bioremediation is an eco-friendly technique for pollutants treatment and it is considered to be an effective environmental clean-up technology as compared to the conventional methods. Moreover, it is a pollution-free/control technology for degradation of toxic pollutants converted into less harmful products or mineralization. However, the microbial activity may require an additional supply of nutrients, appropriate electron acceptors, or establishment of suitable soil pH and temperature. Schäfer et al. (1998) reported that the success of bioremediation approaches may be reduced by the low contact probability of contamination and microorganisms as a result of the heterogeneity of the soil matrix. Later, Alexander (2000) explained the highly homogeneous nature of the polluted soil arising from diffuse pollution, which easily associated with micro-colonies of microbes within the soil/solid matrix. Bosma et al. (1996) calculated that the average distance between bacterial micro-colonies in soil was in the range of 50 to 100 μm, whereas in homogeneous soil pollution it was likely to be in the sub-millimeter range. Generally, two types of bioremediation

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processes were used to remove toxic substance present in the contaminated environment. In the biological process, microbial consortia utilized the contaminants in the soil, which act as bacterial inoculum in so-called bio-augmentation. Bio-stimulation states that the enhancement of biodegradation by autochthonous bacteria in soil could be increased by the addition of nutrients. A significant number of studies have reported the degradation of polyaromatic hydrocarbon-polluted soil in laboratory experiments (Qin et al. 2015; Wick et al. 2004, 2007, 2010). Every technique has its own merits and demerits, and bioremediation also has some limitations. Before using the bioremediation process, initial assessment of the soil is needed. The bioremediation treatment process depends highly on the availability of nutrients, moisture content, pH, and permeability and temperature of the soil matrix. Geller (1991) observed that carbon (C), nitrogen (N), and phosphorus (P) are the main nutrients for microbial cell growth and its activity. The approximate ratio of 250:10:3 for C:N:P is suitable for microbial growth, but this ratio is most often not found at contaminated field sites. Sometimes higher N values at the soil site also cause microbial inhibition. The growth of microorganisms requires favorable pH condition and also requires a suitable temperature at which microbes can survive, in the range 20  C to 30  C. Some chemicals are highly resistant to the bioremediation process, for example, heavy metals, radioactive materials, and some chlorinated compounds such as polychlorinated biphenyls (PCBs). The biological process of contamination may produce secondary toxic metabolism that affects the environment. Bioremediation is a slow process and is site specific: a small-scale study should be done before implementing this technology at the pilot scale. The bioremediation process can also be combined with the mechanical treatment of the soil, which stimulates the process. Thus, the bacteria move in the entire soil matrix to encounter pollutants, but this action is not feasible for the remediation of entire, heterogeneous sites; the desired remediation cannot be achieved within economically acceptable timeframes (Harms and Wick 2006). Different research articles cover the treatment of polyaromatic hydrocarbons in contaminated soil by autochthonous microorganisms, but no studies are available on the treatment of reactive dye in contaminated soil by bioremediation. Tailored design and easy application in the field are necessary for the successful removal of dyes.

4 Biotransformation of Contaminants The biotransformation of contaminants is the alteration of environmentally persistent organic/inorganic contaminants into easily degradable substances using microorganism activity. The gain of energy resulting from microorganism growth and maintenance in organic pollutants is called the biotransformation process. The degradation reaction is mainly the result of physiological coupling of the redox process. In the redox process, reduction and oxidation reactions occur within a cell. In general, electrons are transferred from one compound (called the electron donor or oxidation) to an electron-accepting compound (called the electron acceptor or reduction). Many authors (Tiehm and Schmidt 2007; Wiedemeier 1999) have reported that,

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conversely, electron acceptors occur in relatively oxidized states and are reduced during microbial metabolism. The microbial fuel cell (MFC) is the best example of the biotransformation process, in which organic matter can be converted into energy using microorganisms (Lovley 2006; Oh et al. 2004; Pant et al. 2010). Many researchers have focused on the degradation of industrial effluents with energy production using low-cost electrodes and indigenous microorganisms or mixed cultures to develop an environmentally sustainable process (Venkata Mohan et al. 2012; Zuo et al. 2006). The CECRI research group (Karthikeyan et al. 2009, 2013; Rajeswari et al. 2016; Rengasamy and Berchmans 2012) used various industrial effluents such as soak liquor from the tannery industry with 2–4% sodium chloride and spoiled wine from the wine industry in MFCs, focusing on degradation and energy production in an eco-friendly manner. According to the Central Pollution Control Board (CPCB) of India, 3000–4000 l of soak liquor is generated per ton of skin. This type of industrial effluent should be treated in an electrochemical method, the “microbial fuel cell.” Rajeswari et al. (2016) used soak liquor as anolyte in an MFC for the first time, and Bacillus cereus and Klebsiella oxytoca were used as microorganisms which were converted into electrical energy when graphite plates were used as electrode. The chemical oxygen demand (COD) removal efficiency was about 93  5%, with a maximum power density of 44.04 mW/m2 for 168 h of MFC operation. The selection of suitable electrodes and membranes enhances the cost of the process. Further, formation of a heterogeneous biofilm on the anode surface decreased the efficiency whereby direct and indirect electron transfer determines the efficiency of this electrochemical MFC technique. It can be understood that the biotransformation process is also a slow process wherein the presence of inorganic content determines the bacterial physiology.

5 Potential Benefits of Converged Electrokinetic and Biological Processes The converged EK and biological remediation is a technique that can clean a contaminated site by the mechanism of both microbiological phenomena for degradation and the EK process for the dispersal of bacteria or transportation of the microbes, contaminant, nutrients, etc. (Chilingar et al. 1997). An overview of the electro-bioremediation process and its mechanism for successful application is discussed next.

5.1

Bacterial Mobility in the Contaminated Soil

In the bioremediation process, the microorganisms must be in contact with bioavailable contaminants. These types of soil environment are favorable for the bacteria immobilized in situ because microorganisms attached to the soil particles form

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micro-colonies that do not release single cells from the matrix into the soil (Costerton and Lappin-Scott 1989). Most researchers are interested in studying the performance of microbial transport in the porous matrix. They examined various soil types with bacterial transport and observed that inefficiency of bacterial mobility was mostly attributable to the extremely low hydraulic conductivity in the soil micropores (Li et al. 1996; Silliman et al. 2001), as well as microbial community attachment to the surface of the porous matrix (Baygents et al. 1998). Bacterial mobility can be enhanced by the addition of surfactants or chemical modification of the bacterial surface, which may enhance the efficacy of bacterial transport in terms of a centimeter scale to a meter scale in the porous media during the EK process. Some investigators reported that the negative charge of microorganisms moves towards the anode in the electrophoretic process (DeFlaun and Condee 1997; Lee and Lee 2001). Other reports suggested the dispersal of microorganisms towards the cathode with electro-osmotic water flow (Suni and Romantschuk 2004; Wick et al. 2004). Wick et al. (2004) reported that the rate of bacterial transport in electrophoretic phenomena was about 4 cm2/(V h) in porous media, which depended on the type of matrix and physicochemical parameters of the microbial cell surface. They claimed that the transport of weakly charged bacteria moved predominantly by electro-osmosis when compared to the electrophoresis process (Wick et al. 2004). The EK transport phenomena were inhibited when strongly charged and highly adhesive bacteria attached to the soil surface (Wick et al. 2002). The strong affinity between microorganisms and solid matrices is partially overcome through treating the bacteria with the non-ionic surfactant Brij 35, with up to 80% enhanced EK dispersion achieved (Wick et al. 2004). The electro-osmosis process can be used for mobilization of non-ionic charged molecules such as hydrocarbons, polyaromatic hydrocarbons, PCBs, phenol, and acetic acid contaminants in the soil matrices. Earlier, it was demonstrated that EKs alone can remove/transport non-ionic organic molecules from soil matrix to the cathode section. Generally, transport by electro-osmosis is a process in the range between 1 and 10
109 m2/(VS) for various types of soil matrices (Casagrande 1947). However, Wick et al. (2004) calculated the transport rate of the electroosmosis process with bacteria was about 0.1–0.4 cm2/(Vh). Liu et al. (1999) experimentally observed that Escherichia coli bacteria are predominantly moved in capillaries exclusively by the electro-osmosis process over a wide range of pH values and an electric field strength greater than 0.3 V/cm. It can be concluded that the mobility of bacteria or organics towards the cathode depends upon the speed of electro-osmosis.

5.2 5.2.1

Degradation of Pollutants Hydrocarbons

Polycyclic aromatic hydrocarbons (PAHs), which are the most important class of pollutants, are generated by the inadequate combustion of carbon-based fuels and are

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ubiquitously found in tar, oil, petroleum products, and coal deposits (Mueller et al. 1996). Reports are available on the removal of organic pollutants such as pentadecane, phenol, pentachlorophenol, creosote, and diesel in laboratory-, bench-, and field-scale experiments under an electric field (Lear et al. 2004; Luo et al. 2005). The organic pollutants can be classified as polar or nonpolar compounds. The polar organic compounds can be removed by an introduced direct current that can lead to the removal/migration of pollutants by the EK phenomena (Wick et al. 2004). Earlier, the non-polar organic pollutants could be removed only by an electro-osmosis process. Later, some researchers achieved the removal of nonpolar organic pollutants in the heterogeneous matrix by using different types of surfactants such as Brij 35 (Wick et al. 2004) and Triton X-100 (Lahlou et al. 2000). Wick and his group (Wick et al. 2004) studied EK transport of PAH-degradation bacteria in model aquifers and soil. They demonstrated that electro-osmosis is a valuable mechanism to transport bacteria in the subsurface, with transport efficiencies heavily depending on the retention of the bacteria by the solid phase. The weakly negative charged bacterial strains (L138, LB501TG) were highly transported by electro-osmosis and electrophoresis (0–20% only). The poor EK transport of strongly charged and highly adhesive bacterial cells in the matrix is enhanced by the addition of a non-ionic surfactant.

5.2.2

Reactive Dyes

The combined bio-EK technology was developed to accelerate the uniform transport of microorganisms in the removal of organic pollutants (polychloroethylene, trichloroethylene) that are capable of degradation of pollutants present in the soil (Hassan et al. 2016). Lear et al. (2004) explained that the impressed voltage/current does not have a direct effect on soil microorganism and also soil health will be improved by factors such as soil pH, bacterial physiology, and temperature. To the present, only one work on the influence of direct current (DC) on the bacterially mediated degradation of textile dye organic molecules and salts present in contaminated soil has been reported (Annamalai et al. 1984/DEL/2014). The electric field was used as a tool to inject the bacteria cells into the contaminated soil to enhance bacterially mediated organic degradation. The bacterial strains Brevibacterium halotolerans, Achrosomonas sp. (cellulase-, amylase-, and laccase positive), Bacillus subtilis (cellulase- and amylase positive), and Pseudomonas aeruginosa were used in this study to enhance the degradation of organic compounds in the soil. The bacteria in different soil sections were enumerated during the process shown in Fig. 8.4. The increase in the bacterial count was altered at various soil sections at different time intervals. After the seventh day of electro-bio-stimulation (EBS), the bacterial count at the cathode section reached the same level as in the anode side, which indicates the flow of bacterial cells from the anode compartment to the cathode compartment. The bacterial mobility was mediated through the electro-osmosis process (Harms and Wick 2006; Liu et al. 1999). There was a significant increase in the bacterial count at the top layer of the middle section while the bacterial growth

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Fig. 8.4 Mobility and enumeration of bacteria during EBS process: (A1, A2, A3, anodic sections; M1, M2, M3, middle sections; C1, C2, C3, cathodic sections) (personal observation)

at the middle and bottom layers of the middle section was the same as in the anode and cathode sections. Based on colony morphology, it was observed that all the inoculated bacterial types used in the consortium were uniformly distributed at all the sections. The anode and the cathode side of the soil were adjacent to the electrolyte, where diffusion of the electrolyte may dominate and pH is maintained in the acidic and alkaline range, respectively. The authors claimed that neutral pH (pH 7) favors bacterial growth and that the electro-osmosis process is the main contributor for enhancement of bacterial mobility towards cathode compartments. The possibility of cyclodextrin formation was explained by Annamalai et al. (2014a, b) (Indian Patent No:1984/DEL/2014). It can be claimed that the available protons and carbon from glucopyranose units can act as nutrients for bacteria in the soil, which reduced the COD significantly (Fig. 8.5), and possibly improved the fertility of the soil. Leitgib et al. (2008) explained that cyclodextrin as the injection solution for enhanced organic removal is more efficient when compared to the other solubility agents used for the extraction of organic pollutants. It was assumed that the in situ formation of cyclodextrin from starch may enhance organic removal. Similarly, Maturi and Reddy (2006) also explained that cyclodextrin has the potential to enhance the simultaneous removal of metal and PAHs in low-permeability soil. Annamalai et al. (1984/DEL/2014) concluded that the conductivity of the soil (Fig. 8.6) was reduced effectively at 0.28–1.5 dS/m from 15.5

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Fig. 8.5 Measurement of COD content during EBS process: (a) top layer soil section (0 to 10 cm); (b) middle layer soil section (10 to 20 cm); (c) bottom layer soil section (20 to 30 cm) (personal observation)

Fig. 8.6 Measurement of conductivity during EBS process: (a) top layer soil section (0 to 10 cm); (b) middle layer soil section (10 to 20 cm); (c) bottom layer soil section (20 to 30 cm) (personal observation)

dS/m, which agreed well with the agricultural norms (Choi et al. 2010). Furthermore, the EBS process may increase the phosphorus content in the soil treated by bacterial cells. The efficiency of plant germination in treated soil (87.5%) was better than in untreated soil (25%). Hence, this electro-bio-stimulation technique can improve the fertility of the soil.

5.2.3

Nitrate and Phosphate

Nitrate pollution has been a crucial environmental issue throughout the world since the 1970s, causing serious environmental problem such as the eutrophication of rivers, stomach cancer, cytogenetic defects, and birth defects. The international drinking water quality standards recommended nitrate in the range of 50–100 mg/l as an acceptable level. Synthetic fertilizers can improve food production in a short period of time, so many farmers apply this fertilizer for higher productivity. However, the addition of excess fertilizer saturates the soil with salts, particularly nitrates.

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The excess nitrates are hazardous to human health (Choi et al. 2009). Many research groups are concentrating on nitrate removal by such methods as biological techniques using denitrification and nitrification processes (Fierro et al. 2008; Okeke et al. 2002; Rajakumar et al. 2008; Schipper and Vojvodić-Vuković 2001), hydraulic gradients (Manokararajah and Ranjan 2005), chemical methods (Ahn et al. 2008; Su and Puls 2004), EKs (Ottosen and Rörig-Dalgård 2007; Sahli et al. 2008; Yang et al. 2008), and biofilm reactors (Park et al. 2006). Nitrate is removed from the soil mainly by EK processes: electromigration, electro-osmosis, and electrophoresis. Cairo et al. (1996) demonstrated that the nitrate moves as much as 3 m in the soil at different concentration. Kim et al. (2005) also observed dominant strains of heterotrophic Bacillus in Korean soil, where this biological process can remove nitrogen and phosphorus as well as organic matter efficiently (Choi et al. 2002). The relationship between pH and the efficiency of nitrate reduction in the electrobioremediation systems with an iron electrode was evaluated by Choi and his research group (Choi et al. 2009) in South Korean soil. They tested three types of processes used to remove nitrates from contaminated soil: the EK process, the bio-electrokinetic (bio-EK) process, and a biological process. Choi et al. (2009) identified 21 strains (Bacillus spp.) from a Jinju-vinyl house used as nitrate reducers in the EK process. Rajakumar et al. (2008) investigated various types of organic sources, such as glucose, starch, cellulose, sucrose, and acetic acid, for removal of nitrate by aerobic Pseudomonas sp. and Bacillus sp. in laboratory experiments, and they suggested that starch was the best organic source in nitrate reduction. Choi et al. (2009) postulated, however, that the supply of rich H+ ion by electro-osmosis from EKs encourages the denitrification process (Eq. 8.3). In the electro-bioremediation process, the bacteria (Bacillus spp.) converted nitrate into ammonium ions and nitrogen gas (Eqs. 8.3 and 8.4), which can be reduced by bacteria by supplying the electrons and formation of NH3 (Eq. 8.5). In another method, the high-pH OH ions attract the ammonium ions moving towards the cathode, and the ammonium ions can be converted into NH3 gas at the cathode side (Eq. 8.6). þ þ 5H2 2 NO 3 þ 2H

NO 3

þ 2H

þ

þ 4 H2

 2NHþ 4 þ 2e  NHþ 4 þ OH

!

!

! !

N2 þ 6H2 O NHþ 4

2 NH3 "

NH3 " þ H2 O

þ 3 H2 O þ H2

ðAlkaline conditionÞ

ð8:3Þ ð8:4Þ ð8:5Þ ð8:6Þ

It is understood that the biological system encourages the nitrification process which is not favorable for removal of nitrate in the contaminants, whereas the EK and electro-bioremediation processes highly encourage the denitrification process. Phosphorus is one of the essential nutrients for plant cultivation, but the dumping of manure into agricultural soil increases phosphate concentration in surface waters or groundwater and is responsible for the eutrophication of lakes and streams. In 1991, the European Union Urban Waste Water Treatment Directive insisted on the removal of phosphorus from domestic and industrial water. For the treatment of

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phosphate-containing soil/water with such techniques as physical, chemical, and biological methods, chemical methods are the most effective and best established methods to date (Boisvert et al. 1997; Fytianos et al. 1998; Yeoman et al. 1988). Choi et al. (2010), designed various systems for the removal of phosphates from agricultural soil. They used an iron anode and carbon as cathode while impressing the potential between the electrodes where phosphate ions move towards the anode because phosphate exists as H2PO4 and HPO42. The starch acts as an electrolyte for removal of phosphate, at such voltage gradients as 0.5, 1.0, and 2.0 V/cm. During the EK process, iron dissolution occurred at the anode, migrating into the soil portions. The distribution of iron is higher in anodic and cathodic section compared to the middle section, possibly because of electromigration of iron in soil by EK. The following mechanism was proposed for phosphate removal by this group. H3 PO4

!

þ H2 PO 4 þ H

ð8:7Þ

H2 PO 4

!

H PO2 þ Hþ 4

ð8:8Þ

H PO2 4 Fe

3þ

þ

H PO2 4

Fe3þ þ H2 PO 4

! !

!

PO3 4

FeHPOþ 4 FeH2 PO2þ

þ H

þ

ð8:9Þ

ðEO þ EM processesÞ

ð8:10Þ

ðEO þ EM processesÞ

ð8:11Þ

Fe3+ is formed at acidic pH (4.5) and moves from the anode by the electromigration process: the phosphate ion reacts with iron in the formation of FeHPO4+ and FeH2PO42+. The total surface charge was positive; thus, the iron phosphate complexes move to the catholyte by electro-osmosis and electromigration processes. Choi et al. (2010) concluded that iron as an anode is a feasible method for the removal of phosphate from farm soils by EKs where bacterial influence on removal of phosphates is negligible.

5.3

Limitations

Integration of EKs and the bioremediation process for cleanup of contaminated soil is a promising technology which can be used in both microbiological phenomena for degradation and EKs for the transport of subsurface contaminants, nutrients, and contaminant-degrading microbes (Chilingar et al. 1997; Harms and Wick 2006; Li et al. 2016). The bio-EK is an emerging process for treatment of contaminants present in the heterogeneous matrix. It should be mentioned here that EK has some limitations for treatment of pollution in the soil. If the contaminated soil has a higher amount of calcium and magnesium, hardness will be deposited over the electrode during the EK process (personal observation; Fig. 8.7). The bacterial mobility may be reduced by the adsorption of calcium and magnesium on the cell wall of bacteria. Once the calcium ions have precipitated over the electrode, current

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Fig. 8.7 Calcium and magnesium deposit over cathode surface during EK process

flow may be interrupted in long term experiments and the efficiency of removal may be reduced. The selection of the microorganism is an important parameter because the electro-bioremediation process is mainly associated with the bioavailability of hydrophobic/hydrophilic contaminants or the lack of nutrients. The treatment of high-saline soil contains chloride ions which are converted into chlorine gas and/or hypochlorite in the bio-EK process. The formation of hypochlorite and chlorine gas may affect bacterial mobility and sometimes inhibits bacterial growth. Hence, anolyte selection should be done with low chloride. Based on the literature, more data are available from bench-scale studies; few reports demonstrate the field setup (Table 8.3).

6 Conclusions and Future Perspectives Soil contamination is one of the most prominent fields in the treatment of polluted lands, and EK is a unique technology which can operate in low-permeability soils. EK is a promising technology for the in situ/ex situ treatment of inorganic and organic dye pollutants in agricultural soil. South Korean researchers predominantly use the EK process in the effective removal of salinity from agricultural soil. Mass transfer by electromigration is vital in the removal of metal ions and salts present in the contaminated soil where electro-osmosis is the main agent of removal of these non-ionic pollutants. The EK process can effectively reduce the salinity and conductivity of the soil, which correlates well with the agricultural norms. Hydrophobic/nonpolar organic compounds are difficult to remove by the EK process alone because these are not ionized by electrolytes or the processing solution. Therefore, two different biological and EK processes are tailored to accelerate the bioremediation process and the simultaneous removal/degradation of inorganic salts, metal ions, and polar/nonpolar organics. The integrated bio-EK process has been successful in laboratory-scale experiments whereas its field application is still in a developing stage. Thus, the bio-EK process has some research gaps for future improvement, mainly field-scale implementation. The performance of bio-EK remediation may be improved by the evaluation of ecological parameters, viz., types of pollutants, pH, applied voltage gradient, water content, selection of suitable electrodes, optimized cell design, and electrode arrangement, which can be made eco-friendly and cost-effective.

n-hexadecane contaminated soil (Wang et al. 2016)

Oil contaminated soil (Li et al. 2010)

PCP contaminated soil (Harbottle et al. 2009)

Agricultural soil (Choi et al. 2009) Model aquifers and soil (Wick et al. 2004)

Source & adopted techniques and reference Agricultural soil (Choi et al. 2013)

Bacterial collected from oil-contaminated soil The microorganism isolated from oil contaminated soil adjacent to Liaohe Oil field

Sphingomonas sp. L138, Mycobacterium frederiksbergense LB501TG Sphingobium sp. UG30

Microorganisms Heterotrophic bacteria and nitratereducing bacteria Bacillus spp.

81.5%

Pentachlorophenol/ 13% for anodic side, 75% removal was achieved in cathodic side 72.6%

PAH and petroleum hydrocarbon

Nitrate, 100%

Target contaminant/ Removal achieved (%) Nitrate, 100%

Graphite

Graphite cylindrical electrodes

Graphite

Titanium/ graphite Titaniumiridium electrodes

Electrodes Anode/ Cathode Cast iron/ graphite

Deionized water

Water (3.2)

Water (3–4)

1% starch (3) 0.05 M Trisacetate buffer (7)

Electrolyte Anolyte (pH) 1% starch

Table 8.3 Summary of the literature on contaminants removal from soil by microorganisms

Deionized water

Water (12.5)

Water (7)

Distilled water (7.5) 0.05 M Trisacetate buffer (7)

Catholyte (pH) Distilled water

45

100

36

1

7

Duration (d) 7

1.3 V/cm

1 V/cm

3.14 A/m2

2 V/cm

Voltage gradient/current density 0.5 and 1.0 V/cm, (2 V/cm higher voltage gradient poor efficiency). 2 V/cm

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Future directions will involve the development of agricultural land fertility by the simultaneous removal of inorganic and organic pollutants. EK can be used in the removal of chloride. The interference of chloride/hypochlorite on bacteria should be avoided using biochar during bio-EK. Researchers are currently looking for a natural adsorbent such as biochar with a high water capacity and self-redox behavior so it can easily desorb or remove the toxic pollutants, alter microbial communities, and reduce greenhouse gasses emissions in the agricultural field. The tailoring of biochar with bio-EKs for soil remediation is environmentally benign, cost-effective, and with promising potential for further research. Acknowledgments I gratefully thank the Academy of Science and Innovative Research (AcSIR), CSIR-Central Electrochemical Research Institute. CSIR-HRDG, New Delhi is gratefully acknowledged for the Senior Research Fellowship of Sivasankar Annamalai. The authors thank CSIR for sponsoring this project under Sustainable Environmental Technology for Chemical and Allied Industries (SETCA) – Project No: CSC 0113.

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Chapter 9

Microbial Fuel Cell (MFC): An Innovative Technology for Wastewater Treatment and Power Generation Mostafa Rahimnejad, Maryam Asghary, and Marjan Fallah

Abstract Microbial fuel cells (MFCs) have been nominated as new alternatives and novel opportunities which are able to convert biodegradable organic matters (as substrates) into green electricity with the aim of different types of active microorganisms as active biocatalysts. In terms of configurations, one-chambered MFCs (OC-MFCs), dual-chambered MFCs (DC-MFCs), tubular, H-type, upflow MFCs, and stacked ones would be introduced each for specific objectives. Basically, MFC configuration consists of a biological anode and an abiotic cathode chamber separated by a proton exchange membrane. Direct production of electricity out of substrates, enabling to be operated efficiently at an ambient temperature, and expanding the diversity of fuels used as energy requirements are some of the most praiseworthy advantages of MFCs. Due to electron and proton release resulted by oxidized substrates in anode compartment, sufficient information about electron transfer mechanisms of microorganisms is essential to reach raising amount of energy produced by an MFC system and to find out the theory about their operation. In the 1980s, scientists have figured out that adding some electron mediators causes an incredible enhancement in power output and current density of mentioned technology. By this demonstration, the mediator acts as a movable agent which transports electrons between electrode and bacteria in anode part. Moreover, the most useful applications of MFCs can be classified into four significant categories. They have the ability to be used for electricity production, generation of biological hydrogen, and wastewater treatment (WWT) plants. Besides, MFCs was used as power generator for sensors and biosensors or serve as biosensors themselves. Hence, use of MFCs in water quality improvement which is related to WWT has attracted many scientists all over the world over recent years. Consequently, by

M. Rahimnejad (*) · M. Fallah Biofuel and Renewable Energy Research Center, Faculty of Chemical Engineering, Babol Noshirvani University of Technology, Babol, Mazandaran Province, Iran e-mail: [email protected]; [email protected] M. Asghary Department of Analytical Chemistry, Faculty of Chemistry, University of Mazandaran, Babolsar, Mazandaran Province, Iran © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_9

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using these novel technologies, online monitoring of various parameters related to water quality such as biological oxygen demand, toxicity, and total organic carbon is achievable. Keywords Fuel cells · Microbial fuel cells · Wastewater treatment · Biosensors · Electricity generation

1 Introduction In these days, energy as a guaranty for the economic status of many countries all over the world is a key and vital issue to be expressed. Moreover, fossil fuels (FCs) as the main source of energy produce harmful and problematic pollution for our living environment (Rahimnejad and Najafpour 2018). As a result, finding appropriate, cost-effective, and cheap alternatives to use instead of FCs is an emergency nowadays (Logan 2004). To reduce CO2 emission and global warming, great attentions have been paid to renewable energy sources recently. Furthermore, electricity is one of the most important preferable and flexible form of energy produced by FCs (Bard and Faulkner 2001; Daud et al. 2011). The concept of how to use FCs to convert chemical energy into electricity was discovered by the German scientist Christian Friedrich Schönbein in the early nineteenth century (Grote 2010; Guo et al. 1996). Like batteries, FCs do not have the ability to save energy; they only can alter one type of energy into another type without using the materials inside the cell. Considering the fact, anode and cathode electrodes and electrolyte or membrane would be introduced as the three main parts of aforementioned cells. Based on the electrolytes used by these cells, they can be classified into many groups such as zinc-air FCs, solid oxide FCs, alkaline FCs, molten carbonate FCs, biological FCs (BFCs), and formic acid FCs (Rahimnejad and Najafpour 2018). As a matter of fact, high efficiency, high-power output, easy transportation, being compatible with environmental rules, and high-energy density are some of the most well-known advantages of FCs. In addition, there are some disadvantages for them such as high cost of catalysts, corrosive electrolytes, and the harsh conditions of operation (high temperature and pressure) (Palmer et al. 1995; Selman 1993). BFCs as one type of FCs which basically consist of two compartments (anode and cathode chambers), substrates can be oxidized by microorganisms or enzymes resulting in the releasing of protons and electrons. Then, protons and electrons go through the proton exchange membrane (PEM) and external circuit, respectively. Consequently, by the reaction of oxygen, electrons, and hydrogen ions, water molecules will be generated in cathode compartment. Microbial fuel cells (MFCs), sedimentary MFCs (SMFCs), and enzymatic MFCs are three main types of BFCs which separately can be operated for specific targets. MFCs can be introduced as a novel and cost-effective technology to convert organic matters including lignocellulosic biomass and low-strength wastewater into electricity and also would be able to beneficially combine with

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Fig. 9.1 Schematic diagram of a DC-MFC which can be connected in a series or in parallel to produce enough power for a specific applications [Rahimnejad and Najafpour 2018]

applications in WWT (Pant et al. 2010; Li et al. 2010). Both in terms of number of researches and applications of MFCs, real interest has tremendously grown over recent years. MFCs’ applications can be classified into four major issues: (1) WWT; (2) biological hydrogen production; (3) electricity generation; and (4) biosensor (Fig. 9.1).

2 Fuel Cells The development of the FCs was first invented by William Grove. He invented the first hydrogen-oxygen fuel cell in 1839 (Steele and Heinzel 2001). Typically, FC is identified as an electrochemical device which transforms the chemical energy out of a fuel directly into electricity by oxidizing fuel in anode compartment and reduces oxidant in cathode part by using noble metal catalysts (Leech et al. 2012). Every FC system consists of four main components: two electrodes (anode and cathode), where chemical reactions (anodic and cathodic reactions) take place, and electrical energy produces. The anode is the electrically negative terminal, while the cathode is the electrically positive terminal. An electrolyte medium is sandwiched between two electrodes, which separates these two electrodes and transfers charged particles from one electrode to another. A metal catalyst has been also used to speed up the chemical reactions happening at the electrodes’ surface (Fig. 9.2) (Winter and Brodd 2004). Generally, in FC system, hydrogen as fuel will break down to electrons and protons under chemical reaction. Generated electron and protons at the anode

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Fuel Cell –

e–

X

e–

+

4 OH+

4H

2 H2O

2 H2

O2 Separator

Anode

Electrolyte

Cathode

Fig. 9.2 Representation of basic mechanism of a typical FCs system. (Winter and Brodd 2004)

surface during oxidation pass through external circuit and electrolyte solution, respectively, and then recombine with oxygen in cathode chamber to produce water molecules. Generally, FCs are classified into different types according to the nature of the electrolyte solution, since each type requires particular materials and fuel to be well operated. Hence, there are various classifications for FCs including polymer electrolyte, alkaline, phosphoric acid, molten carbonate, and solid oxide.

2.1

Biological Fuel Cells

Telling the truth, BFCs are one of the most famous types of FCs that transform biochemical energy to electrical energy by using living cells (bacteria, algae) or catalysts extracted from cells (enzymes, enzyme cascades, and mitochondria) as biocatalysts. In contrast with chemical FCs, BFCs do not require any metals such as platinum (Pt) to be used as a catalyst. Also, they work in mild conditions of pressure and temperature. Table 9.1 indicates a comparison between BFCs and chemical FCs (Table 9.1). There are two major categories belonging to BFCs such as enzymatic FCs (EFCs) and MFCs [12] which are going to be discussed in following paragraphs:

2.1.1

Enzymatic Fuel Cells

The working principle of EFCs is as the same as traditional FCs, which employ enzymes as biocatalysts instead of metal catalysts (Ivanov et al. 2010). There are various types of enzymes that would be applied in EFCs’ design; however,

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Table 9.1 Comparing BFCs and FCs in terms of operating conditions (Rahimnejad and Najafpour 2018) Characteristics Catalyst pH Temperature Electrolyte Capacity Efficiency Fuel type

BFCs Enzyme, microorganism 7–9 Environmental temperature Phosphate solution Low ˃50% Carbohydrate

Chemical FCs Metals Acid solution Over 200
C Phosphoric acid High 40–60% Natural gas, hydrogen

oxidoreductases have been applied most often (Ramanavicius and Ramanaviciene 2009). Despite high selectivity of enzymes to catalyze oxidation of a wide range of fuels, the use of enzymes in EFCs is limited by their major specific problems such as their low electron transfer rate, poor enzyme stability, high cost of extraction, separation, and purification of enzymes (Asghary et al. 2016; Zhou and Dong 2011).

2.1.2

Microbial Fuel Cells

MFCs as a novel and promising technology have gained increasing attention during the past decades due to their ability to provide new opportunities for the sustainable and green energy out of biodegradable organic matters by using microorganisms as biocatalyst (Rabaey and Verstraete 2005). In MFCs, active microorganisms eliminate the requirements for the isolation of individual enzymes, and they are able to catalyze a more complete oxidation reaction of many biofuels. Also, their ability to be less susceptible in order to poisoning under neutral conditions resulting in the bioelectricity production from reductive substances has been demonstrated recently (Ramanavicius and Ramanaviciene 2009).

2.2

History of MFC

The first concept of MFC has been introduced by Potter in 1911. After that Logan and Regan figured out that bacteria such as Escherichia coli and Saccharomyces have the capability of generating electrical current by breaking down organic compounds such as acetate and glucose. After the first innovation, efforts in this area were almost stopped for about 55 years (Park and Zeikus 2002). Until the 1950s and early 1960s, no serious researches on MFCs have been done. In the 1980s it was discovered that the current density and power output could be greatly enhanced by addition of electron mediators. However, it was discovered that sometimes there is no need to use any types of artificial mediators to transfer produced electrons from bacteria to the surface of anode electrodes (Gil et al. 2003) since there are lots of microbes such as Clostridium beijerinckii and

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Actinobacillus succinogenes which have the ability to directly transfer the electrons derived from the metabolism of organic compounds to the electrodes’ surface. Aforementioned types of MFCs are known as mediator-less ones. Introducing such bacteria and construction of microorganisms which have the ability to generate electricity from degrading organic matters have been discussed tremendously (Logan and Regan 2006). In addition, the vast progress of MFC was in 1999, but the main interest in MFC has been increased over recent years.

2.3

Materials for Construction of MFC

MFCs are being made by using a variety of materials and in an ever-increasing diversity of configurations. Nowadays, various types of carbon materials such as carbon paper, carbon cloth, graphite, and carbon black are used as an anode electrode in MFCs because of their good conductivity, stability, and high surface area in a microbial inoculum mixture (Qiao et al. 2007). Providing an oxygen reduction reaction (ORR), cathode catalysts have to be coated onto the surface of cathode electrode (e.g., Pt). There are other supporting materials such as polypyrrole and polyaniline which their structures significantly affect MFCs’ performance (Bezerra et al. 2008; Lee et al. 2009). Moreover, PEM plays an effective role in MFCs, especially in DC-MFC, because of their ability to separate anode and cathode compartments while simplifying proton transfer from anode to cathode. In addition, using BPM (bipolar plate membrane) which consists of CEM (cation exchange membrane) and AEM (anion exchange membrane) has been introduced as another separator used in these novel technologies (Kim et al. 2007a, b, c; Deng et al. 2010). Consequently, metallic materials like graphite and stainless steel have been extensively used for bipolar plate membrane (Rahimnejad et al. 2015).

2.3.1

Anode

Microorganisms, substrate, mediator (it’s optional), and the anode electrode as an electron acceptor are the major components which anode part fed with them. Microorganisms as an oxidizing agent are vital to be present in anode part. Besides, substrates as an important source of nutrients and electron donor would be an effective parameter in biological processes of MFCs such as their electricity production (Pant et al. 2010; Park and Zeikus 2002; Jafary et al. 2013). Due to large surface area and supreme electric conductivity, graphite fiber brush, graphite rod, carbon cloth, carbon paper, and reticulated vitreous carbon (RVC) are some of the mostly used carbon materials in anode (Logan et al. 2006) (Table 9.2). As it’s an acknowledged fact, types of electrode and substrate used in anode compartment would be able to have an increasing effect on MFCs’ performance and efficiency which is demonstrated below.

Gammaproteobacteria and Shewanella affinis (KMM3586) Deltaproteobacterium Desulfuromonas spp. and others

Carbon paper

Graphite

Noncorroding graphite Graphite with Mn4

Cysteine

Marine sediment reached in acetate Marine sediment

Sewage sludge

Graphite with neutral red (NR)

+

Betaproteobacterium

–

Ethanol

Sewage sludge

Geobacter spp.

Carbon paper

Lactate

Escherichia coli

Escherichia coli

Saccharomyces cerevisiae G. sulfurreducens

Graphite Carbon paper

Glucose Acetate

Bacteria Geobacter spp. (Firmicutes)

Anode Carbon paper

Substrate Glucose

Table 9.2 Effect of various types of anodes on MFCs’ efficiency

SC-MFC

SC-MFC

DC-MFC

DC-MFC

DC-MFC

DC-MFC

DC-MFC

DC-MFC DC-MFC

System configuration DC-MFC

152

91

25.4

14

36

40  2

52  4.7

16 48.4  0.4

Maximum power density (mW/m2) 40.3  3.9

Reference Bettin (2006) and Jung and Regan (2007) Rahimnejad et al. (2009) Bettin (2006) and Jung and Regan (2007) Bettin (2006) and Jung and Regan (2007) Kim et al. (2007a, b, c) and Watanabe (2008) Logan et al. (2005) and Zhou et al. (2012) Rahimnejad et al. (2015) and Bond et al. (2002) Gil et al. (2003) and Tenderet al. (2002) Nevin et al. (2008) and Park and Zeikus (2003) Nevin et al. (2008) and Park and Zeikus (2003)

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Cathode

Protons which are produced in the anodic chamber transfer into the cathodic compartment passing the PEM, which completes the electrical current. The different methods to increase the amount of oxygen in cathodic solutions had been offered (Park et al. 2002; Rhoads et al. 2005). For instance, using an air pump and a proper oxidant can successfully increase the amount of oxygen and consequently increase the output power. Furthermore, using a catalyst such as Ptas an improving agent has been investigated, but the mentioned materials are not cost-effective to use in MFCs, so different biocathods have been offered to increase the amount of dissolved oxygen of solution consequently (Park et al. 2002). Several important process parameters such as microbial electron transfer, cell metabolism, internal and external resistance, and cathode oxidation greatly affect generated electron transfer from anode to cathode part and also MFC’s performance influenced by them (Rahimnejad et al. 2014).

2.3.3

Proton Exchange Membrane

Proton exchange membrane (PEM) acts as a separator between the cathode and anode box and is the most effective factor influencing MFCs’ performance (Chen et al. 2008). +Likewise, CEM same as PEM is one of the most extended used separators in MFCs. It’s worth noting that PEMs prevent transferring substrates, minerals, and oxygen from anodic compartment to cathodic compartment. Although they have such a beneficial effect on MFCs’ operation, they cause high internal resistance (Rin) which is not known as a positive element. As a result, by removing PEMs which results reduced Rin, stacked MFCs, SC-MFCs, and upflow MFCs have been proposed by recent studies to reinforce their current generation (Chen et al. 2008). Here, there are some of specific drawbacks related to PEMs as follows (Rozendal et al. 2008): (1) expensive nature; (2) cause lower cathode cell performance; (3) cause lower efficiency and performance for MFCs; and (4) increase internal resistances.

2.4

MFC and Its Design

In terms of structure, there are different types of constructions for MFCs such as SC-MFC, DC-MFC, and stacked ones, and each of the mentioned constructions would be able to be used for specific purposes. In SC-MFCs the cathode is directly exposed to air, and consequently there’s no need to provide a cathode compartment for this type of MFCs. DC-MFCs consist of an anode and a cathode compartment separated by a PEM which is well understood.

9 Microbial Fuel Cell (MFC): An Innovative Technology for Wastewater. . .

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Electron Transfer Mechanism in MFC

Rate of transferring produced electrons onto the surface of anode is the key and essential parameter, which influences MFCs’ performances and power generation. There are different types of mechanisms for electron transfer in anodic compartment including mediated electron transfer (MET) and direct electron transfer (DET). Despite the fact, there has been no accurate and complete transport protocol up to now belonging to transfer produced electrons by microorganisms to the surface of anode electrode. Microorganisms as biocatalysts and consumer of different substrates (carbon sources) are capable of releasing positive and negative ions in anodic chamber. By oxidization of substrates such as glucose as electron donors in anodic chambers, related reactions take place both in anode and cathode parts as follows (Eqs. (9.1), (9.2), (9.3) and (9.4)): (i) If glucose is used as substrate: Anodic reaction : C6 H12 O6 þ 6H2 O

! 6CO2 þ 24e þ 24Hþ

Cathodic reaction : 6O2 þ 24e þ 24Hþ ! 12H2 O

ð9:1Þ ð9:2Þ

(ii) If acetate is used as substrate: Anodic reaction : CH3 COO þ H2 O ! 2CO2 þ 2Hþ þ 8e þ



Cathodic reaction : O2 þ 4H þ 4e ! 2H2 O

ð9:3Þ ð9:4Þ

Based on reaction (glucose), 24 moles of protons and electrons are produced through the complete oxidation of 1 mole of pure glucose in an anaerobic condition. The important challenge here is how does the produced electron travel to the anode? To answer this question, various methods of transferring electrode have to be discussed.

2.5.1

Direct Electron Transfer (DET)

Interactions between microbe and electrode have been studied through different reviews (Debabov 2008; Lovley 2011). Some microorganisms can directly transfer produced electrons to anode electrode (Lovley 2011). Also, it was shown that active microorganisms could prompt electrons from organic substrate and recover electrical current. For more information, in these systems electrons can be transferred to the surface of anode electrode under anaerobic conditions and can be managed to cathode electrode typically under aerobic conditions which caused reduced oxygen (Logan 2009). Different limitation other than microbial metabolism rate is introduced previously, which seriously limits the power generation of MFCs by usage of

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mentioned method. However, necessary engineering is necessary in order to obtain envisioned applications. This will most likely not only include electron delivering systems but also require genetic engineering of the microorganisms to raise demanded reactions.

2.5.2

Mediated Electron Transfer (MET)

Some recognized microorganisms have the ability of using soluble components which physically vehicle produced electrons from an intracellular compound to anode electrode surface (Rabaey and Verstraete 2005). For example, some redox materials such as neutral red (Park et al. 1999), thionin (Lithgow et al. 1986), and methyl viologen (Roller et al. 1984) have been used as electron shuttle previously. On the other hand, there are some microorganisms producing redox mediators by themselves, which occurs in two categories: through the organic production and through oxidizable metabolite generation (Rabaey and Verstraete 2005). Also, some bacteria such as E. coli, Pseudomonas, Proteus, S. cerevisiae, and species of Bacillus are not able to transport the produced electron to the surface of anode electrode. As a result, mediators are the essential parts of MFCs. Besides, there are two types of mediators known as self-generated mediators and artificial electron mediators (Rahimnejad and Najafpour 2018). Furthermore, mediators such as pyocyanin and relevant compound produced by especial types of bacteria have been used to transport the electrons to the anode surface. It’s worth noting that sometimes the electrons themselves play the role as mediators and transfer the electrons to the anode surface.

2.6

Performance of MFC

Some specific important parameters such as rate of substrate conversion, overpotential, the performance of PEM, and Rinin MFCs tremendously affect their performance (Rabaey and Verstraete 2005). To be added here, various parameters such as the mixing and mass transfer phenomena in the reactor and the bacterial kinetics (μmax, the maximum specific growth rate of the bacteria, and Ks, the bacterial affinity constant which is related to substrate) are directly relevant to substrate conversion rate in MFCs. Moreover, MFCs’ performance can be influenced by the electrochemical characteristics of the electrode, the electrode potential, and the kinetics together with the mechanism of the electron transfer and MFCs’ current (Jang et al. 2004). Considering the fact Rin of MFCs is dependent on both the resistance of the electrolyte between the electrodes and membrane resistance. Nafion as a PEM has the lowest resistance and is the most chosen one (Deng et al. 2010).

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2.7

225

Measurement and Calculation

Bioelectricity is produced in a MFC system only when the entire reaction is exposed to the thermodynamics law (Logan et al. 2006; Zielke 2006). Power (Pcell) Generally, the performance of a MFC will be evaluated by measuring and evaluating power output and columbic efficiency. Pcell or power obtained from the cell can be calculated as (Logan et al. 2006): Pcell ¼ I  Ecell

ð9:5Þ

where: I ¼ electricity current (amps) Ecell ¼ cell potential (volt) Power Density For operational application of a MFC and a comparison between the current density of several systems, power is often normalized with projected anode surface area (Aan), since the biological reaction is taking place at the anode surface. Therefore, the power density is calculated by using the following equation (Eq. (9.6)): Power density ¼ Pcell =Aan

ð9:6Þ

Coulombic Efficiency Although power generation is the major objective in MFCs’ operation mode, extracting electrons stored in biomass and its recovery, which is called coulombic efficiency (CE), is an essential and important parameter to be discussed (Rahimnejad and Najafpour 2018). Indeed, CE is also defined as the ratio of recovered electrons as current, which transferred to the anode surface to the total coulombs in organic materials. The total coulombs transferred in the system are evaluated by integrating the current over the time. Hence, CE over a period of time (t) as a vital agent in a batch MFC mode is calculated as (Logan et al. 2006; Esmaeili et al. 2014): Z

t

CE ¼ M

I dt=F b V an ΔCOD

0

in which: M: The molecular weight of oxygen (M ¼ 32). B: The number of electrons exchanged per mole of oxygen (b ¼ 4), F: Faraday’s constant. Van: The working volume of the anode compartment. ΔCOD: The change in chemical oxygen demand (COD) over time (t).

ð9:7Þ

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Applications of MFC

As stated earlier, MFC as a new and novel technology is validated to be applied in some of the interesting sections which are classified into four superb categories.

2.8.1

Electricity Generation

As previously mentioned, providing sustainable production of energy out of biodegradable reduced organic components is one of the most valuable applications of MFCs. Considering the fact, MFCs not only function on diverse types of uncomplicated carbohydrates but also have an effective function on complex substrates presented in wastewaters (Rabaey and Verstraete 2005). Consequently, knowing about different metabolic pathways used by bacteria to optimize and develop energy production is an essential knowledge to be considered. Because of the low rate of electron transfer in MFCs, power levels in MFC systems are relatively low (Rahimnejad and Najafpour 2018), and that’s why mediators have been used as an accelerating agent for transferring electrons in anode chamber. Generating appropriate power for the small electrical devices would be the main purpose of MFCs. For instance, ten LED lamp and one digital clock had been turned on with fabricated stacked MFCs and operated for the about 2 days by Rahimnejad et al. in 2014 (Rahimnejad et al. 2015). As a matter of fact, anaerobic sewage sludge can be introduced as appropriate inoculums for MFCs due to containing super valuable bacteria communities that are electrochemically active ones (Table 9.3). Moreover, it has been investigated that at least one order of noteworthy difference in the maximum power density had been shown by MFCs using the same fuels (Kim et al. 2007a, b, c). This obviously shows that MFCs’ performance is identically influenced by their configuration style and Rin (known as overpotential) (Kim et al. 2007a, b, c). Based on numerous efforts done by researchers, power generation in MFCs can be increased by the different types of PEM and also the variety of Rin (Min et al. 2005). Consequently, optimization of electrolyte and configuration of MFCs to decrease Rin and providing a full potential of catalytic activity of microbes would be required to increase the power output (Liu et al. 2005).

Table 9.3 Various measured power density of MFCs filled by different consortia Reactor type SC-MFC SC-MFC SC-MFC DC-MFC DC-MFC DC-MFC

Fuel Glucose Glucose Glucose Acetate Glucose Acetate

Power density (mw/m2) 766 1540 480 860 5850 1030

Reference Cheng et al. (2006a, b, c) Cheng et al. (2006a, b, c) Cheng et al. (2006a, b, c) Heijne et al. (2006) Rosenbaum et al. (2006) Jong et al.(2006)

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Biosensors

The IUPAC defined electrical biosensors as a self-contained integrated device providing specific analytical information using a biological recognition element (microbes) contacting with an electrochemical transduction element (anode and cathode electrodes) (Sun et al. 2015; Dai and Choi 2013). Based on this definition, the MFC system, which uses microorganisms as the recognition element in the anodic compartment to directly and efficiently produce an electrical energy in response to exogenously added analyte, could be considered as biosensors (Sun et al. 2015). In addition, MFCs are able to act as power supplies for some small devices such as sensors and biosensors, and also they themselves have the capability of serving as biosensors. In comparison with traditional biosensors for online and rapid monitoring of environmental parameters such as biological oxygen demand (BOD), a total organic carbon (TOC) which requires an external power and a transducer to convert measured signal to electrical signal MFC-based biosensors does not need any external power supplies and transducer since the current output is electrical by itself. These biosensors have many advantages such as long-term stability, continuous monitoring, fast analyzing, being cost-effective, and being able to be a safe power source (Kumlanghan et al. 2007). For example, the MFC system can be used as an online biosensor for observing metabolic activity of the microorganisms due to its online monitoring of current and operating without a transducer to read the signal output and external power source. A MFC with duration of about 5 years to produce stable current generation has been demonstrated by Kim et al. By this demonstration, the strength of wastewater was directly proportional to the coulomb generated from MFC, which provides the opportunity to use it as a BOD biosensor (Kim et al. 2010). This is its unique property, which facilitates the construction of portable and self-powered biosensors based on MFC techniques in contrast with other types of online biosensors that often need complex nonlinear transducers (Stein et al. 2010). Several researchers have reported the use of MFC in BOD biosensor, since the coulomb and current density generated by the MFC is directly proportional to the BOD concentration and the strength of the wastewater (Kim et al. 2003). As the MFC output mainly depends on the microbial activity and transfer rate of electrons from the microorganisms to the anode surface, presence of toxic compounds which inhibited the microbial activity will decrease the MFC output compared to a situation without any toxicants. This means that the MFC is feasible to serve as a sensor for monitoring of toxicity in water too (Sun et al. 2015; Stein et al. 2010). Also, using MFC as a power supply in construction of a self-powered DNA biosensor to identify genetic defects is another application of mentioned technology which has the most novelty compared with other ones (Asghary et al. 2016).

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Biological Hydrogen Production

MFC systems can be optimized to produce biological hydrogen instead of producing electrical current. Then, the generated hydrogen could be accumulated for later applications. Generally, hydrogen production out of protons and electrons produced by microbial metabolism in MFC is not appropriate in terms of thermodynamics. But, by using a microbial electrolysis cell (MEC), the thermodynamic barrier would be solved. MEC system is a reactor for biological hydrogen production by combining MFC and electrolysis, and the anode process of an MEC is the same as that of a MFC system, but in the cathodic chamber, the electrons combine with the protons to produce hydrogen (Sun et al. 2008). Since oxygen as the electron acceptor at the cathode chamber has a higher redox potential than the microbial anode, electrons flow spontaneously through the external circuit and produce electricity in MFCs, while, in the MEC system, the reduction reaction of H+ ions to generate hydrogen at the cathode chamber has a lower redox potential than the anode and is not possible thermodynamically. Hence, electrons do not flow spontaneously through the external circuit (Liu et al. 2010). In addition, in a MEC system as bioelectrochemically assisted microbial reactor, hydrogen is evolved at the cathode chamber by eliminating oxygen and adding a small voltage to the circuit. Indeed, in a MFC-MEC coupled system, hydrogen was produced in an MEC, and the extra power was supplied by an MFC. This microbial reactors can produce 8 moles hydrogen per 1 mole glucose as substrate (Das 2009).

2.8.4

Wastewater Treatment

One of the most praiseworthy and fruitful applications of MFCs is their functional use in WWT plants which is much more superior compared with the other introduced processes. Using MFCs in WWT was started by Hoberman and Pommerin in 1991 (Habermann and Pommer 1991). Although various techniques have been offered for WWT over the last decades, high cost and time-consuming are the main drawbacks of them which have been introduced (Rahimnejad and Najafpour 2018), and most treatment methods require a high level of operational process. It should be added here that recovery of energy as electricity in WWT with the help of MFCs and lower sludge production compared with aerobic process are some of the most novel advantages of applying MFCs in water treatment (Kim et al. 2007a, b), since large amount of excess sludge disposal generated in WWT process is needed and so important. Obviously, all the energy of organic contaminant is utilized by microorganisms in an aerobic process, while in MFCs huge amount of their energy is converted to electricity, and just a small amount is dedicated to microbial growth (Park et al. 2001). Introductory efforts have been shown that MFCs’ yield is 1.5 times more than an aerobic culture (Du et al. 2007). In terms of MFCs’ ingredients, wastewater is divided into organic and inorganic wastes, and it has been figured out that by using specific microorganisms in MFCs, they would be able to remove

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sulfides from wastewater by Rabaey et al. in 2006 (Rabaey et al. 2006). It’s worth remarking that an oligotrophic MFC is able to be operated with a BOD value of about 5 mg/l (Moon et al. 2005; Phung et al. 2004), and in some cases up to 90% of COD has been removed (Wang et al. 2012), and the CE of about 80% has been shown before (Kim et al. 2005).

2.8.4.1

Water Quality Improvement by MFC

Providing clean water and adequate sanitation systems in developing countries to overcome diseases is required. Up to now, various chemical and biological methods have been used to measure the water quality and to ensure the safety of water for consumers. However, these methods have several problems including being expensive and complexity in use. But, many research have been done in the world to prove the use of the MFC technology for measuring and monitoring water quality due to its simplicity, rapid response, and its ability to function on-site and real time (Chouler and Di Lorenzo 2015; Orta et al. 2017).

2.8.4.2

Comparison of MFC with Conventional Wastewater Treatment Plant

There are two main disadvantages for the conventional aerobic treatment. The first one is the high capital investment, and the second one is the remarkable and significant operational and energy consumption cost. For instance, sewage aeration shows an energy requirement of about 0.5 kWh/m3. Besides, great amounts of surplus sludge are produced that need a suitable treatment (Aelterman et al. 2006a, b). To overcome the aforementioned drawbacks, MFCs are the best answer to several problems that traditional treatment plant faces. Ability of energy recovering out of wastewater and limiting both the energy input and the excess sludge production are the advantages of WWT by MFCs (Rabaey and Verstraete 2005). The other advantage of MFCs lies in the fact that it is their direct harvesting of electricity unlike the conventional way which is a two-step process (Aelterman et al. 2006a, b). It’s worth noting that MFC has tremendous advantages in comparison with conventional activated sludge (CAS) such as using an air-cathode MFC to consume gaseous oxygen from the atmosphere (Park and Zeikus 2003) that can significantly reduce the operation cost of an MFC-based treatment plant. Table 9.4 shows a comparison between MFC, AD (anaerobic digestion), and CAS (conventional activated sludge) (Lefebvre et al. 2011). As it’s indicated below, in terms of applied load and energy Table 9.4 Comparison between MFC, AD, and CAS for WWT WWT plants CAS AD MFC

Treatment efficiency High Moderate Moderate

Applied load Low High Low

Sludge production High Low Low

Energy balance  + 

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balance, MFC is between CAS and AD, while in terms of treatment efficiency and sludge, production is closer to AD.

2.8.4.3

Recent Advances of MFC for Wastewater Treatment

MFCs was operated by US government (Space program) as a suitable technology for both space flights and also power generation in 1966 (Bullen et al. 2006). Furthermore, recent developments (Du et al. 2007), practical implementation (Pham et al. 2009), anode performance (Li et al. 2011), cathodic limitations (Rismani-Yazdi et al. 2008), and types and substrate composition (Pant et al. 2010) in MFCs have been reviewed over the past decades. As stated earlier, MFCs were used for power generation and treatment of wastewater simultaneously (Oh and Logan 2005). Specific new improvements in MFC technology including their application as MEC, using both increased external potential at cathode and anoxic cathode, have been triggered (Oh and Logan 2005).

3 Case Studies The MFC has nominated as a promising and green progress technology for WWT and electrical energy generation simultaneously (Zou et al. 2008). For example, Lei et al. (Li et al. 2008) constructed a DC-MFC for loses of Cr6+ from actual electroplating wastewater through reduction of Cr6+ to Cr3+ in the cathodic chamber. The reactor involved plain carbon felt and graphite paper as anode and cathode, respectively. As shown in Fig. 9.3, removing of chromium ions on the surface of cathode by using proposed DC-MFC has produced maximum power density of 1600 mW/m2 and a current density of about 0.4 mA/m2. Hence, the obtained performance by MFC was higher than the other researches in WWT by MFC. In other research work, Logan et al. (Liu et al. 2004) fabricated prototype SC-MFC for local WWT which is accompanied by electrical power production (26 mW/m2). Their fabricated SC-MFCs have had the ability of removing at least 80% of the COD of the domestic wastewater. In a study carried out by our research group (Izadi and Rahimnejad 2013), a DC-MFC system was fabricated for eliminating various concentrations of S2 ions (0.1, 0.8, and 1.5 g/L) and bioelectricity production. These results revealed that during the MFC operation for 72 h, 98% of the S2 ions were approximately removed from the anodic chamber of MFC. Also, by the experimental results, the maximum value of power (48.65 mW/m2) at maximum current (231.47 mA/m2) was obtained for the DC-MFC system in the steady-state condition.

9 Microbial Fuel Cell (MFC): An Innovative Technology for Wastewater. . .

Cathode potential Power density

Power density (mW/m2)

1600 1400

650 600 550

1200

500

1000 450 800 400

600

350

400

Cathode Potential (mV)

1800

231

300

200

250

0 0.0

0.2

0.4

0.6

0.8

Current density (mA/cm2)

Fig. 9.3 Produced bioelectricity in the proposed DC-MFC by Lei et al. (Li et al. 2008)

4 Conclusion Since using FCs causes environmental pollutions, finding renewable energy sources such as MFCs as an alternative to nonrenewable energy sources has been considerably discussed over recent years. Generation of bioelectricity, second fuel (biohydrogen) production, usage in WWT plants, and also serving as a sensor or a biosensor are the most vital applications of MFCs. Besides, these analytical devices can be used for bioremediation of toxic compounds. However, MFC technology is still in research level, and tremendous efforts need to be done to make them available for commercialization.

References Aelterman P, Rabaey K, Clauwaert P, Verstraete W (2006a) Microbial fuel cells for wastewater treatment. Water Sci Technol 54:9–15 Aelterman P, Rabaey K, Pham HT, Boon N, Verstraete W (2006b) Continuous electricity generation at high voltages and currents using stacked microbial fuel cells. Environ Sci Technol 40:3388–3394 Asghary M, Raoof JB, Rahimnejad M, Ojani R (2016) A novel self-powered and sensitive labelfree DNA biosensor in microbial fuel cell. Biosens Bioelectron 82:173–176 Bard AJ, Faulkner LR (2001) Electrochemical methods. Fundamentals and applications, 2nd edn. Wiley, New York

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Chapter 10

Functional Diversity of Plant Endophytes and Their Role in Assisted Phytoremediation Angélica Leonor Guerrero-Zúñiga, Eugenia López-López, Aída Verónica Rodríguez-Tovar, and Angélica Rodríguez-Dorantes

Abstract The functional diversity term helps to understand the biological complexity through the wide range of interactions that organisms show on communities and ecosystems as they may interact. In a particular manner, organisms may have attributes or characteristics that define their role within the ecosystems. The purpose of this review is to analyze the importance of plant growth-promoting traits of endophyte bacteria that define the functional diversity of them in their relationships with plants in assisted phytoremediation techniques. Keywords Plant growth-promoting bacteria · Endophytes · Functional diversity · Phytoremediation

A. L. Guerrero-Zúñiga Gerencia de Transformación de Biomasa, Instituto Mexicano del Petróleo, Ciudad de México, Mexico E. López-López Laboratorio de Evaluación de la Salud de los Ecosistemas Acuáticos, Departamento de Zoología, Escuela Nacional de Ciencias Biológicas, Prol. Carpio y Plan de Ayala, S/N, Unidad Profesional Lázaro Cárdenas, Instituto Politécnico Nacional, Ciudad de México, Mexico A. V. Rodríguez-Tovar Laboratorio de Micología, Departamento de Microbiología, Escuela Nacional de Ciencias Biológicas, Prol. Carpio y Plan de Ayala, S/N, Unidad Profesional Lázaro Cardenas, Instituto Politécnico Nacional, Ciudad de México, Mexico A. Rodríguez-Dorantes (*) Laboratorio de Fisiología Vegetal, Departamento de Botánica, Escuela Nacional de Ciencias Biológicas, Prol. Carpio y Plan de Ayala, S/N, Unidad Profesional Lázaro Cardenas, Instituto Politécnico Nacional, Ciudad de México, Mexico © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_10

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1 Introduction In general, it is known that the community structure of plant growth-promoting bacteria is characterized by their phenotypic and genotypic characteristics which include standard plating methods, with the employing of selective media and biochemical tests to analyze the physiological profiles inside of communities. The named functional properties like colonization, phytohormone synthesis, solubilization of insoluble phosphorous, 1-aminocyclopropane-1-carboxylate (ACC) deaminase activity, and production of antibiotics and siderophores have also been used to characterize them. This review focused on the employ of these functional traits and multivariate statistical analysis to describe the functional analysis of plant growthpromoting bacteria communities associated to environmental conditions that could be related to assisted phytoremediation cases.

2 Classification of Plant Growth-Promoting Bacteria It is known that beneficial bacteria that facilitate plant growth are often named plant growth-promoting bacteria (PGPB) (Kloepper and Schroth 1978). These microorganisms must be able firstly to colonize plant surfaces, after established, survive and multiply in the surface or within plant tissues, and, finally, facilitate plant growth (Barea et al. 2005). There are some authors who classified PGPB regarding their interaction with plants as symbiotic bacteria and free-living bacteria (Khan 2005). Also as Somers et al. (2004) and Khan et al. (2009) established according to their ecological roles, they are biofertilizers (favoring the bioavailability of nutrients to the plants), phytostimulators (facilitating the plant growth by the synthesis of phytohormones), rhizoremediators (involved in the removal or degradation of inorganic and organic pollutants), and biopesticides (managing plant diseases by the production of antimicrobial metabolites.

3 Bacteria Endophyte Characteristics There are other classifications according to their plant’s localization by Khan et al. (2009): intracellular PGPB which are bacteria residing inside plant cells or localized inside specialized structures like nodules and extracellular PGPB that are living outside plant cells enhancing the plants’ growth by the production of signal compounds that directly stimulate growth, improving the disease resistance or sensing the nutrient status. These authors also mention that this group can be subdivided into three types, based on how they are in close contact with plants: those living near, but not in contact, with the plants’ surface; those colonizing plants’ surface; and those

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living inside the spaces between cells of plants’ tissues. Colonizing microbial communities are often specific to the different plant compartments or organs, including roots, stem, leaves, flowers, as well as fruits and seeds (Compant et al. 2010). Schulz and Boyle (2006) noted that endophytic bacteria are located in plant tissues beneath the epidermal cell layers, colonizing the internal tissues and establishing different interactions that include symbiosis, mutualism, or commensalism. Authors like Compant et al. (2010) and Zheng et al. (2016) mention that this kind of microorganisms establishes a harmonious relationship within the plants and does not cause negative effects on their health. Luo et al. (2011) reported that these bacteria have been isolated from a wide range of plant species, suggesting an ubiquitous existence of them. Even Dharni et al. (2014) and Ma et al. (2015b, 2016) noted that it is known that PGPB participate on plants’ growth and heavy metal phytoremediation in polluted soils; He et al. (2013), Chen et al. (2014), Babu et al.(2015), and Ma et al. (2015a) resumed that there is little knowledge about plant endophytic bacteria interactions and their potential role in phytoremediation. Phetcharat and Duangpaeng (2012) named also these bacteria as plant growth-promoting endophytic (PGPE) bacteria and suggested that their colonization and plants’ propagation enhance soil fertility and stimulate the host plant development by providing growth regulators. It also has been demonstrated by Ma et al. (2011) that endophytic bacteria may help host plants to adapt under unfavorable environmental conditions and increase the phytoremediation efficiency, promoting the plants’ growth, alleviating the metal stress, reducing metal phytotoxicity, and finally altering the metals’ bioavailability and translocation of them inside the plants. Finally, to ensure that the isolated bacteria are really a plants’ endophyte, Visioli et al. (2014) and Maropola et al. (2015) noted that all the methods reviewed that are used to isolate and characterize them include the surface sterilization of the host tissues and isolation of the endophytic bacteria using appropriate growth media and the bacterial detection and identification by molecular methods (e.g., direct amplification of bacterial DNA from colonized plant tissues).

4 Functional Diversity Definitions in Ecological Context Petchey and Gaston (2006) noted that Zak et al. (1994) and Stevens et al. (2003) reviewed the use of the term “functional diversity” and commented that it has grown exponentially over the last decade and has been managed in studies of marine, freshwater, and terrestrial ecosystems; actually, this spanned a wide range of taxa including bacteria. Even now there isn’t a standard knowledge about how to define functional diversity, how to measure it, and how to assess its performance. Petchey and Gaston (2006) established that in general, it is known that functional diversity generally involves the understanding of communities and ecosystems based on what organisms do rather than on their evolutionary history and noted that this concept

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about “what organisms do” could be interpreted as the phenotype of the organism, like phenotypic trait, and then some authors equate the functional diversity to phenotypic diversity. In this context, Tilman (2001) proposed a more specific definition: “functional diversity is the value and range of those species and organismal traits that influence that ecosystem functioning.” Thus, a consequence of this definition is the extent of functional differences among the species in a community. Petchey and Gaston (2002a) considered functional traits as components of an organism’s phenotype that influence ecosystem-level processes and established that the measure of functional diversity requires, ideally, the following steps: first, appropriate functional information (traits) about organisms to be selected and included in the measure, and as consequence the irrelevant information needs to be excluded. The simple answer to which traits to use in functional classifications is all traits that are important for the function of interest and no traits that are functionally uninformative. Second, the traits must be weighted according to their relative functional importance; a common measure of functional diversity is the number of functional groups represented by the species in a community (Naeem and Li 1997; Hooper 1998; Hector et al. 1999; Rastetter et al. 1999; Walker et al. 1999; Fonseca and Ganade 2001; Tilman 2001; Tilman et al. 2001; Petchey and Gaston 2002b; Roscher et al. 2004). Third, trait diversity must statistically measure and have desirable mathematical characteristics (Mason et al. 2003; Botta-Dukát 2005; Ricotta 2005). Finally, the measurement of the functional diversity would be able to explain and predict the variation inside an ecosystem-level process. Regarding the statistical methods applied to determine the functional diversity, Petchey and Gaston (2002a) mention that multivariate methods can be associated with functional grouping and functional dendrograms (FD) could be used because they are similar to phenetics, grouping organisms based on observed physical similarities employing primarily multivariate methods proposed by Sokal and Sneath (1973). These authors also noted that it is very important that even this analysis is similar to phenetic analysis, it possesses important differences; phenetics focuses on morphological traits, whereas functional traits are used for functional groupings; thus, different weightings of functional traits will produce different functional dendrograms, none of which are, a priori, correct or incorrect. And the emphasis on FD must be remarked regarding the use of the total branch length of a functional dendrogram to measure functional diversity. Leishman and Westoby (1992), Chapin et al. (1996), Díaz and Cabido (1997), and Fonseca and Ganade (2001) mention that traits must be linked to the function(s) of interest, and Euclidean distance and the unweighted pair group method with arithmetic mean (UPGMA) which produced the distance matrix and functional dendrogram, respectively, are commonly employed. Pace (1997) noted that the greatest research challenge is to integrate these ecological analyses to microbial ecology and environmental microbiology, because the genetic and metabolic diversity of microorganisms is vast. Kirk et al. (2004) mentioned that Trevors (1998) and Ovreas (2000) reported that the methods of studying soil microbial diversity, particularly regarding species diversity, involve the species richness, the total number of species present, species evenness, and the distribution of species. Hughes et al. (2001) categorized these

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methods in biochemical-based techniques and molecular-based techniques and mention that typically, diversity studies consider the relative diversities of communities across a gradient of stress, disturbance, or other biotic or abiotic difference. Atlas and Bartha (1993) noted that the information gathered by diversity studies are finally reduced to discrete numerical measurements defined as diversity indices.

5 Plant Growth-Promoting-Bacteria: Functional Attributes Actually, PGPB’s traits are those that include knowing functional activities: the ability to produce or change the concentration of plant hormones (Mahmoud et al. 1984; Frankenberger and Arshad 1995; Glick 1995; Garcia de Salamon et al. 2001; Ahmad et al. 2008); N2fixation (Zaidi 1999; Wani et al. 2007a); antagonism against phytopathogenic microorganisms by production of siderophores, b-1,3-glucanase, chitinases, antibiotics, and cyanide (Renwick et al. 1991; Shanahan et al. 1992; Flaishman et al. 1996; Khan et al. 2002; Wani et al. 2007a, b); and solubilization of inorganic phosphates and toxic metals (Wani et al. 2007b, c). Becerra-Castro et al. (2012) and Cabello-Conejo et al. (2014) established that the screening methods employ inert growth substrates that allow a rapid selection of interesting strains and after they could be tested in more complex systems, where bacterial mechanisms should be freely induced. Ullah et al. (2015) mention that it is important to note that not only works about PGPB’s functional diversity precise mode of action; ecophysiology and promising potential bioinoculants for maintaining soil fertility and the sustainability of crops in diverse agroecosystems are also important to determine if these microorganisms possess the ability to affect heavy metal mobility and availability to the plant through the release of chelating agents, acidification, phosphate solubilization, and redox changes as Yan-de et al. (2007) and Gadd (2004, 2005, 2010) established. There have been a large number of studies (Lebeau et al. 2008; Kidd et al. 2009; Glick 2010, 2014; Becerra-Castro et al. 2013; Sessitsch et al. 2013; Muehe et al. 2015) indicating that plant-associated microorganisms are indeed essential players in metal phytoextraction or phytomining, enhancing the plants’ growth and health by the increase of nutrient uptake and improving their resistance to pathogens and stress (Göhre and Paszkowski 2006; Lebeau et al. 2008; Lugtenberg and Kamilova 2009). It is known that most of phosphate-solubilizing bacteria and siderophore producers, bacteria with ACC deaminase activity and phytohormone producers, improve plants’ growth and transform heavy metals into soluble and bioavailable forms, favoring that plants take up contaminants (Ullah et al. 2015). Thus, these kinds of bacteria can assist in the phytoremediation of heavy metals, either directly or indirectly: directly involving the solubilization and removal of them from solid matrices, such as soil, dumps, sediments, and other industrial and municipal wastes, giving more bioavailability and final accumulation by plants, and indirectly, by the improvement of plants’ growth to prevent the effect of phytopathogens, facilitating the accumulation of heavy metals (Gadd 2004; Yan-de et al. 2007; Glick 2010).

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6 Analyzing the Functional Diversity of Plant GrowthPromoting Bacteria: Study Cases Griffiths et al. (2000, 2001) and Wertz et al. (2006) noted that several studies indicate that the functional redundancy of microbial communities, in general, is high and that microbial diversity may be substantially eroded without affecting ecosystem functions. The importance of biodiversity for ecosystem functioning has been much debated, and some contrasting results have been reported. The final and general agreement showed that “it is not biodiversity per se that is important, but rather the prevalence of individual species or functional groups” (Cederlund et al. 2008). Shukla et al. (2004), Boruvka et al. (2005), Guiffre et al. (2006), and Sharma et al. (2011a) reported that multivariate analysis including principal component analysis (PCA) and cluster analysis could be a useful tool to select effective PGPB and has been employed in crop yield prediction, agricultural soils, and under contrasting management systems, to identify the origin of potentially toxic elements in ecosystems and to develop operationally important quality ecosystem indicators for longterm sustainability. The authors noted that multivariate analysis takes into consideration the whole data set instead of individual variables, thereby taking into account several factors simultaneously; particularly, PCA is an effective data reduction analysis that helps to explain most of the variances in a multivariate data, reducing the number of variables into a few uncorrelated components. Subsequently, the employ of cluster analysis allows the confirmation of PCA data, grouping the individual inoculants and variables, allowing the understanding and characterization of the nature of each cluster, and, finally, giving a profile that relates to the occurrence of specific clusters to auxiliary data not used in the original analysis. There are some particular studies regarding this kind of PGPB functional diversity analysis; Naik et al. (2008) studied the genetic and functional diversity of phosphate-solubilizing fluorescent pseudomonads associated with rhizospheric soils of rice and banana by an array of in vitro assays and gene amplification technique. Since strains of fluorescent pseudomonad bacteria have also been reported for biodegradation of agricultural pollutants (Bano and Musarrat 2003; Naik and Sakthrivel 2006), as well as for weed control in agricultural fields (Kremer et al. 1990; Charudattan 1991), these bacteria have been considered as an important bioinoculants due to their innate potential to produce plant growth-promoting hormones and enzymes (O’Sullivan and O’Hara 1992; Cattelan et al. 1999; Glick et al. 1995; Ramamoorthy et al. 2001; Sunishkumar et al. 2005). Bacteria belonging to the genera Mesorhizobium, Rhizobium, Klebsiella, Acinetobacter, Enterobacter, Erwinia, Achromobacter, Micrococcus, Aerobacter, and Bacillus have been reported as phosphate solubilizers, and strains belonging to pseudomonads are also noted as efficient phosphate solubilizers (Villegas and Fortin 2001). Considering the multiple applications of phosphate-solubilizing fluorescent pseudomonads, it is essential to study their diversity, which is useful in the design of strategies to the employ of native strains as bioinoculants.

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Sharma et al. (2011a) analyzed the functional and taxonomic diversity of rhizobacteria in Saccharum munja in abandoned mine, analyzing the variations in plants’ growth-promoting traits of rhizobacteria in different seasons and the significance of rhizobacteria in the establishment of this plant species at the mine site. Knowing that the functional diversity among the rhizobacteria may vary with the specific requirements of the host plants during their different growth phases (vegetative and reproductive) under different seasons, these authors concluded that the rhizosphere of S. munja contains functionally diverse bacteria dominated by Bacillus spp. and Paenibacillus spp. Abundance of the most efficient PGPB during the reproductive phase of the plant indicated the significance of analyzing seasonal variation among the rhizobacteria and concluded that plant growth-promoting traits analyzed in the isolated rhizobacteria and their potential to enhance seed germination and seedling growth of S. munja in mine spoil suggest their significance in the natural colonization of this plant species at abandoned mine site. Another particular example is the Sharma et al. (2011b) study, where the authors evaluated the efficiency of selected strains of plant growth-promoting Pseudomonas that enhanced the productivity of soybean-wheat cropping system in central India, by the analysis of soil enzyme activities, nutrient acquisition, soil nutrient status, and productivity of soybean and wheat under field conditions in vertisols of central India, employing suitable multivariate statistical tools (PCA and cluster analysis). Finally, in this review, a particular case must be noted, considering that the term functional diversity helps to understand the biological complexity through the wide range of interactions that organisms show on communities and ecosystems as they may interact. In addition, organisms may have attributes or characteristics that define their role within the ecosystems. Ortega-Acosta (2015) analyzed the functional diversity of isolated endophyte bacteria from Lemna gibba L. plants obtained from three Xochimilco channel zones in the rainy and dry seasons. According to OrtegaAcosta et al. (2015), Xochimilco’s lake is located in the southern part of Mexico City basin which is a lacustrine zone that comprises a network of channels that, along with the chinampas, confirms a unique ecosystem which has served as a source of aquatic resources, and the use of aquatic macrophytes to the removal of pollutants in a natural way is recommended. There are studies related to the rhizosphere of aquatic plants that are focused on the direct functional analysis and investigated plantmicrobe interactions at the full biological hierarchy. The author selected three distinct areas in the Lake of Xochimilco, interconnected by a system of water channels according to their land use and environmental conditions, namely, (A) Urban zone (“U”, where domestic waters are discharged into the water channels), (B) Tourist zone (“T”, a located zone of markets and boat rides on trajineras in water channels), and (C) Chinampera zone (“CH”, an agricultural zone adjacent to the water channels). Plants of L. gibba were collected from each water channel zone in two seasons (dry season (May, (MA)) and rainy season (August, (AG))), in each selected channel, with the establishment of three sites along them, taken in each site samples of L. gibba plants. In each zone and period of study, environmental factors were recorded in situ: air and water temperature, dissolved oxygen, turbidity, conductivity, pH, transparency, and depth. In addition, water samples were taken

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to assess their quality analyzing the content of total nitrogen, nitrites, nitrates, ammonium nitrogen, sulfates, orthophosphate, and total phosphorus. Determinations of color, total suspended solids, alkalinity, hardness, chlorides, and biochemical oxygen demand were also assessed. To analyze the functional diversity of the isolated endophytes from each zone, these were characterized according to their plant growth-promoting attributes: ACC deaminase activity, indole acetic acid (IAA), siderophore production, and tricalcium phosphate solubilization. These functional attributes were analyzed applying a multivariate analysis; a distance matrix was built using the conventional standard distance coefficient and a phenogram resolved using the UPGMA (unweighted pair group method with arithmetic mean) method. After it, a correlation coefficient of Pearson was obtained. Finally, a PCA was performed using Pearson’s correlation with the environmental variables and the functional attributes of phytobacteria from different seasons and zones of study. Water quality results indicate that the Tourist zone showed most of the parameters analyzed with high values in both periods. Therefore, the Tourist zone was the most affected area, compared to the Chinampera and Urban zone since they do not have high nutrient values and other indicators of eutrophic water, such as mineralization, alkalinity, high concentrations of carbonates, and organic matter decomposition. In the dry season, 17 phytobacteria were isolated from L. gibba plants from the three study areas. Of all the isolated phytobacteria (Table 10.1), only two of the Urban zone tested are positive for siderophore production; two of the Urban zone and five in the Chinampera zone showed tricalcium phosphate solubilization; the activity of the ACC deaminase activity was presented by four isolated zones from plants of the urban zone, two of the Tourist zone, and three in the Chinampera zone. Finally, four of the urban zone, one of the Tourist zone, and five in the Chinampera zone were classified as high producers of IAA. In the rainy season, 14 phytobacteria were obtained (Table 10.1), of which only 2 phytobacteria isolated from plants of the Tourist zone, 1 isolated phytobacterium from the Urban zone, and 3 phytobacteria from the Chinampera zone were positive for phosphate solubilization. Only one isolated from plants collected in the Chinampera zone presented siderophore production and activity of the ACC deaminase, four phytobacteria isolated from plants of the Tourist zone, one of the Urban zone, and three in the Chinampera zone. The 14 phytobacteria isolated showed a high production of IAA in each area and season. A wide range of genus-isolated phytobacteria was obtained, Bacillus sp., Achromobacter spp., Enterobacter sp., and Pseudomonas spp. for the dry season, being more abundant in this season Bacillus spp. In the rainy season, the most representative and abundant isolated genus were Serratia spp., Stenotrophomonas spp., and Pseudomonas spp. Figure 10.1shows the phenogram with the associated groups according to the functional attributes determined in the isolated endophyte bacteria, where two groups forming at first, group I made only by Enterobacter spp. CH-MA-2 with the highest number of attributes, and the rest of the rhizobacteria comprise the group II and group IIa that is composed by the isolated endophytes from the L. gibba plants collected from the three zones in the dry season and group IIb that is also subdivided in two other groups group IIb1 that includes all the

Dry Season Bacillus pumilus U-MA-4 Bacillus stratosphericus U-MA-5 Bacillus subtilis U-MA-1 Bacillus stratosphericus U-MA-6 Achromobacter spp. CH-MA-17 Pseudomonas spp. CH-MA-19 Rahnella aquatilis CH-MA-23 Enterobacter spp. CH-MA-24 Achromobacter spp. CH-MA-13 Enterococcus faecium CH-MA-16 Bacillus pumilus CH-MA-9 Paenibacillus spp. TU-MA-9 Bacillus simplex TU-MA-10 0.2

0.2 0.1 0.2
0.2 0.1


































A A

C

C

C

C

C

C

C

B

B

0.53

0.55







0.4










0.4

0.65 0.48

0.53




+

A

0.83

0.2

+

A

ACC deaminase activityb (mm)

L

L

L

L

L

H

H

H

H

L H

H

H

IAA producerc

2

2

2

2

2

2

2

2

2

3 2

3

4

Functional traits present

(continued)

57.89

10.5

5.30

Traits percentage (%)

Identified endophyte bacteria

Phosphate solubilizationa (mm)

Table 10.1 Functional diversity of endophyte bacteria isolated from Lemna gibba plants collected from two seasons

Siderophore production

Functional Diversity of Plant Endophytes and Their Role in Assisted. . .

Zone

10 245

Identified endophyte bacteria Bacillus spp. U-MA-2 Bacillus spp. U-MA-3 Enterobacter spp. TU-MA-7 Deinococcus spp. CH-MA-15 Rainy season Serratia spp. CH-AG-21 Serratia spp. TU-AG-5 Serratia spp. TU-AG-8 Stenotrophomonas spp. U-AG-17 Serratia spp. CH-AG-20 Serratia spp. CH-AG-22 Stenotrophomonas spp. TU-AG-1 Enterobacter spp. TU-AG-6

Table 10.1 (continued) Phosphate solubilizationa (mm)





0.2 2.1 0.95 1.25 0.6 0.25



Siderophore production







+












Zone A A B

C

C B B A

C C B

B

0.49

0.5 0.51 0.53

0.59 0.48 0.52 0.48




ACC deaminase activityb (mm)




H

H H H

H H H H

L

IAA producerc L H H

2

3 3 2

4 3 3 3

1

Functional traits present 1 1 1

14.28

7.16 35.71

Traits percentage (%) 26.31

246 A. L. Guerrero-Zúñiga et al.

a























B

B

B B B

C

Phosphate solubilization halo diameter, bcolony diameter, cIAA quantified

Pseudomonas spp. TU-AG-3 Stenotrophomonas spp. TU-AG-9 Serratia spp. TU-AG-10 Serratia spp. TU-AG-11 Exiguobacterium spp. TU-AG-12 Raoultella spp. CH-AG19













H

H H H

H

H

1

1 1 1

1

1

42.85

10 Functional Diversity of Plant Endophytes and Their Role in Assisted. . . 247

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Fig. 10.1 Phenogram of the isolated bacteria endophytes from Lemna gibba plants collected in dry ( ) and rainy ( ) seasons; the groups were established according to the phosphate solubilization, ACC deaminase activity, and IAA production functional traits (r ¼ 0.83)

phytobacteria isolated from plants collected in the rainy season and group IIb2 formed by the rest of the endophytes isolated from plants collected in dry season. Figure 10.2 that shows the PCA comprising two principal components (PC1 37.90% and PC2 24.89%) accounted for 62.79% of total variance. In this figure, there was an association between the values of high alkalinity, total nitrogen, turbidity, total suspended solids, ammonium nitrogen, and water hardness for Tourist zone in both seasons; for the rainy season, Chinampera zone was associated to biochemical oxygen demand, chlorides, conductivity, and water color. The Urban zone was associated with higher values of sulfates, nitrates, temperature, transparency, and depth of water and with those isolated endophytes that presented particular plant growth-promoting traits like siderophore production and phosphate solubilization. For the dry season, the PCA associated the Chinampera and Urban zones with the higher values of some water characteristics: nitrites, dissolved oxygen, and total phosphorus and orthophosphate contents. Also in this season, these zones showed an interrelationship with the plant growth-promoting traits of the isolated phytobacteria that possess ACC deaminase activity and higher production of IAA. It is important to note that Chinampera zone has the majority number of the isolated endophytes; thus in this study the ecological characteristics of each site determine the diversity of the isolated phytobacteria, particularly related to IAA production, showing a considerable number of isolates with higher production in both seasons, independently of the site where L. gibba plants were collected. Kebede and Kebedee

10

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Fig. 10.2 PCA of the study zones related to the seasons of collected Lemna gibba plants: (a) PC of the three study zones for dry and rainy seasons and (b) PC of water quality characteristics and functional traits of the isolated endophytes from Lemna gibba plants for both seasons

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(2012) mention that the PCA is a multivariate statistical method that applies to reduce the dimensionality of a data group and through the evaluation of the study done by Ortega-Acosta (2015), the association between the physicochemical characteristics of the water quality of the Xochimilco water channels tested and the functional attributes of the isolated endophytes from L. gibba plants from both seasons; it suggests the functional participation of the phytobacteria in those water channels according to the reported conditions, with the next final consideration for each zone, as follows: Tourist zone was the most stable site, because it presented the similar values of environmental and physicochemical conditions in both seasons, followed by the Urban zone that showed considerable variations with some physicochemical parameters like sulfates, nitrates, total phosphorus and orthophosphate contents, and pH. It is important to note that in this zone there was the minor number of isolated endophytes with the only two phytobacteria that presented the siderophore production trait. At least, Chinampera zone was considered the most irregular zone, where all the environmental conditions tested presented contrasting values; this zone also possesses the highest number of isolated endophytes with the highest IAA production and ACC deaminase activity. The knowledge of the functional participation of the endophytes isolated in each zone would help to consider this microorganism as an alternative to be employed as an assisted strategy of phytoremediation of these water channels in the Xochimilco Lake.

7 Conclusions Finally, with the particular study cases noted, the analysis of the functional diversity of plant growth-promoting bacteria in terms of phenetic clustering and principal component analysis showed differences among isolates, according to the zones under study, and also the environmental variables in each zone determined the diversity of the isolated organisms. Acknowledgments The authors are grateful to the Research Project, SIP-20131494 and SIP-20141314 of the Secretaría de Investigación y Posgrado del Instituto Politécnico Nacional, Comisión de Operaciones y Fomento de Actividades Académicas (COFAA-IPN), EDI (Estímulo al Desempeño de Investigadores-I.P.N.), and Sistema Nacional de Investigadores (SNI-CONACyT), and the fellowships for its support.

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Ullah A, Heng S, Farooq M, Munis H, Fahad S, Yang X (2015) Phytoremediation of heavy metals assisted by plant growth promoting (PGP) bacteria: a review. Environ Exp Bot 117:28–40 Villegas J, Fortin JA (2001) Phosphorus solubilization and pH changes as a result of the interactions between soil bacteria and arbuscular mycorrhizal fungi on a medium containing NO3as nitrogen source. Can J Bot 80:571–576 Visioli G, D'Egidio S, Vamerali T, Mattarozzi M, Sanangelantoni AM (2014) Culturable endophytic bacteria enhance Ni translocation in the hyperaccumulator Noccaea caerulescens. Chemosphere 117:538–544 Walker B, Kinzig A, Langridge J (1999) Plant attribute diversity, resilience, and ecosystem function: the nature and significance of dominant and minor species. Ecosystems 2:95–113 Wani PA, Khan MS, Zaidi A (2007a) Synergistic effects of the inoculation with nitrogen fixing and phosphate-solubilizing rhizobacteria on the performance of field grown chickpea. J Plant Nutr Soil Sci 170:283–287 Wani PA, Khan MS, Zaidi A (2007b) Chromium reduction, plant growth promoting potentials and metal solubilization by Bacillus sp. isolated from alluvial soil. Curr Microbio l54:237–243 Wani PA, Khan MS, Zaidi A (2007c) Co inoculation of nitrogen fixing and phosphate solubilizing bacteria to promote growth, yield and nutrient uptake in chickpea. Acta Agron Hung 55:315–323 Wertz S, Degrange V, Jl P, Poly F, Commeaux C, Freitag T, Guillaumaud N, Le Roux J (2006) Maintenance of soil functioning following erosion of microbial diversity. Environ Microbiol 8:2162–2169 Yan-de J, Zhen-li H, Xiao-e Y (2007) Role of soil rhizobacteria in phytoremediation of heavy metal contaminated soils. J Zhejiang Univ Sci B8:192–207 Zaidi A (1999) Synergistic interactions of nitrogen fixing microorganisms with phosphate mobilizing microorganisms. PhD thesis, Aligarh Muslim University, Aligarh Zak JC, Willig MR, Moorhead DL, Wildman HG (1994) Functional diversity of microbial communities. Soil Biol Biochem 26:1101–1108 Zheng YK, Qiao XG, Miao CP, Liu K, Chen YW, Xu LH, Zhao LX (2016) Diversity, distribution and biotechnological potential of endophytic fungi. Ann Microbiol 66:529–542

Chapter 11

Toxic Metals in Industrial Wastewaters and Phytoremediation Using Aquatic Macrophytes for Environmental Pollution Control: An Eco-Remedial Approach Mansi Rastogi and Meenakshi Nandal

Abstract Toxic pollutants contaminate water by discharging wastewater generated through municipal, industrial, and landfill site waste, etc. It is emerging as a worldwide problem as it enormously affects human, fauna, and flora health of receiving water. During last few decades, the exponential population growth, productivity variation and consumption rates, and resources exploitation along with rapid industrial and technical development are seen as major contributors that accompany water pollution. Wastewater treatment has been a problem for mankind since the discovery of additional environmental problems caused by wastes discharge into surface waters was done. Though control and prevention technologies are being applied to most of these industrial and municipal sources and there is availability of a wide range of wastewater treatment technologies for restoring and maintaining the biological, chemical, and physical quality of wastewaters, still there is a staggering amount of these agents released into the environment. Another major proven threat to water is heavy metal toxicity with several associated health risks. Although they do not play any big biological role, their trace present in certain form can harm the human body and its proper functioning. This chapter discusses wastewater characteristics, toxic metals added to water, the role of plants in constructed wetlands in removal of various pollutants to remediate the wastewaters from various sources, and constraints and future of constructed wetland as a cleanup technique in wastewater remediation. Keywords Wastewater · Pollution · Treatment · Constructed wetlands · Remediation

M. Rastogi (*) · M. Nandal Department of Environmental Sciences, Maharshi Dayanand University, Rohtak, Haryana, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_11

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1 Introduction Toxic pollutants contaminate water by discharging wastewater generated through municipal, industrial, and landfill site waste, etc. It is emerging as a worldwide problem as it enormously affects human, fauna, and flora health of receiving water. During last few decades, the exponential population growth, productivity variation and consumption rates, and resource exploitation along with rapid industrial and technical development are seen as major contributors that accompany water pollution. Though control and prevention technologies are being applied to most of these industrial and municipal sources and there is availability of a wide range of wastewater treatment technologies for restoring and maintaining the biological, chemical, and physical quality of wastewaters, still there is a staggering amount of these agents released into the environment (Mohan and Pittman Jr 2007). Another major proven threat to water is heavy metal toxicity with several associated health risks. The heavy metals, such as cadmium, copper, lead, chromium, zinc, and nickel, are the main environmental pollutants (United States Environmental Protection Agency 1997). Although they do not play any big biological role, their trace present in certain form can harm the human body and its proper functioning. The conventional technologies that reduce and remove wastewater contaminants have limited approach due to indulgence of high capital, operational, and maintenance cost and often lead generation of secondary wastes that hinder treatment process (Aksu 2002). An intricate technology for wastewater treatment is constructed wetland phytoremediation which is one such highly promising technology with competitive performance, costeffectiveness, and being environment-friendly that remediates polluted site being aided by green plants to render them harmless in controlled environment (Raskin et al. 1994) and can be applied both in soil and water for organic and inorganic pollutants (Salt et al. 1998). This chapter discusses wastewater characteristics, toxic metals added to water, and the role of plants in constructed wetlands in removal of various pollutants to remediate the wastewaters from various sources, constraints, and future of constructed wetland as a cleanup technique in wastewater remediation.

2 Wastewater: Natural and Anthropogenic Sources and Characteristics Wastewater generated from agricultural, industrial, municipal, and other activities like extensive use of insecticides, fertilizers, and pesticides result in polluting soil and our environment and gradually make the water resources unproductive (Hakeem 2015) as shown in Fig. 11.1. Among the different sources of pollution, industries are the major polluters as these discharge a highly toxic wastewater containing a variety of organic and inorganic pollutants which may cause serious toxic effects in living beings upon exposure (Saxena et al. 2016, 2019; Goutam et al. 2018; Gautam et al. 2017; Bharagava et al. 2017a; Saxena and Bharagava 2015). The waste materials are

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Fig. 11.1 Sources of wastewater generation

disposed in three main forms (solid, liquid, and gaseous), but the major disposal state is liquid as both the solid and gaseous waste can be eventually altered through hydrological cycle and transported through water course (Armah et al. 2010). The major sources may include the pollutants introduced in the wastewater collection system by rainwater runoff from domestic and commercial sources. The pollutants are divided into two main groups: (1) potentially toxic elements (PTEs) including cadmium (Cd), copper (Cu), chromium (Cr III and Cr VI), zinc (Zn), mercury (Hg), lead (Pb), and nickel (Ni) and (2) organic pollutants including PAHs, PCBs, DEHP, LAS, NPE, dioxins (PCDD), and furans (PCDF). Most of the contaminants, out of the detected 6000 organic compounds in raw water sources, are highly persistent, while others are easily biodegradable (Table 11.1).

2.1

Natural Sources

Wastewater generation in a region is governed to a certain level by the natural factors. The natural factors affecting include rate of precipitation, weathering, and soil erosion. The phenomenon of runoff being seasonal depends on the climatic conditions. The sources mainly include water from floods (stormwater), runoff (rainwater running through cracks present in the ground and into gutters), water from swimming pools, water from car garages, and cleaning centers. The quality of wastewater differs from site to site and season to season due to variation in chemical

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Table 11.1 Major role played by macrophytes in constructed wetlands Wetland plant part Aerial plant tissues Aquatic plant tissues Roots and rhizomes

Role Less growth of phytoplanktons due to light attenuation with reduced wind velocity and risk of settled solid resuspension with good nutrient storage Filtering of larger debris with low current velocity and high rate of sedimentation, less risk of settled solid resuspension, increased nutrient uptake and aerobic degradation with increased photosynthetic oxygen excretion Less soil erosion due to sediment surface stabilization, increased nutrient uptake and organic degradation due to oxygen release and nitrification with release of antibiotics

composition which highly depended on topography, climate, and mineralogical composition of the of the bedrock. The most important natural influences on surface water are geology, hydrology, and climate, since these affect the quantity and quality of available water and their influence is generally greatest when the available water quantities are low (Singh et al. 2004). Temperature is the main factor affecting almost all physicochemical equilibrium and biological reaction and also increases water temperature which will enhance dissolution, solubility, degradation, and evaporation.

2.2

Anthropogenic Sources of Water Pollution

The rapid urbanization, industrialization, intensive agriculture, and growing demand for energy have adversely affected the physiochemical parameters of surface water (Avtar et al. 2011). The anthropogenic sources contributing to water pollution involve, organic and inorganic wastewater discharge from industrial plants and municipal sources (point as well as nonpoint sources of pollution) as presented in Fig. 11.1. (a) Industrial Wastes The rapid industrial development and global growth have led us to recognize the interrelationship between environment and pollution. About two million tons of waste materials constitute the world’s water part on daily basis. In developing countries, 90% of sewage waste and 70% of industrial wastes are becoming the part of water bodies. Almost every industry has a role in water contamination the major contributors being manufacturing, mining, power generating, food processing, and construction industries (McKinney et al. 2013). While almost all industrial activities produce waste that causes some pollution, rather few industries produce pollution in bulk. Industrial pollution caused by generation of wastewater effluent is one of biggest problems being faced by surface water resources. Industries release pollutants like nitrates, nitrites, cations (K+, Ag2+, Na+, Mg2+, Ca2+) and anions (Cl
, CO3
, HCO3
, Cl
), and toxic metalloids such as arsenic, chromium, copper,

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lead, iron, nickel, mercury, cadmium, zinc, and cobalt (Wang and Yang 2015; Alvarez et al. 2016) that are harmful for living organism. Certain waste-contributing manufacturing industries are textile printing, paint, dying, pharmaceutical, leather tanning, plastics, petroleum, paper, and pulp (Raja and Venkatesan 2010; Hakeem 2015). These inorganic and organic polluted effluents may find their way to surface water bodies such as lagoon, Table 11.1. Major roles of macrophytes in constructed wetland Wetland plant part Role Aerial plant tissues Less growth of phytoplanktons due to light attenuation with reduced wind velocity and risk of settled solid resuspension with good nutrient storage Aquatic plant tissues Filtering of larger debris with low current velocity and high rate of sedimentation, less risk of settled solid resuspension, increased nutrient uptake and aerobic degradation with increased photosynthetic oxygen excretion Roots and rhizomes Less soil erosion due to sediment surface stabilization, increased nutrient uptake and organic degradation due to oxygen release and nitrification with release of antibiotics M. Rastogi and M. Nandal rivers, lakes, and ponds through canals and become vulnerable to such pollution (Bhuiyan et al. 2011). One more major contributor is heavy metals being the most toxic pollutant of groundwater and industrial wastewater. Heavy metals are persistent and recyclable bio-geochemically and have biological risks (Alves et al. 2012) and so are of particular concern for surface water quality. (b) Domestic Wastes Domestic wastes comprise of wastes from the houses that become ultimately a part of water resources such as rivers, streams, ponds, lakes, and seas and pollute them severely. They can be categorized further on the basis of nature: 1. Inorganic waste – Nitrate and phosphates being the major component of soaps and detergents being released in freshwater causing damage to humans, animals, and plants. 2. Organic waste – Papers, plastic, polythene bags, rotten fruits and vegetables, and recalcitrant by-products are major causes of organic water pollution. 3. Liquid waste – Dirty and percolating effluents from houses and toxic metalcontaining liquids that leach down and deteriorate the water quality severely (Hudak 2012). In addition, domestic sewage comprises of large amount of suspended solids (such as partially disintegrated particles), colloidal solids, large floating particles of rags, plastic, etc. (Awadallah and Yousry 2012). (c) Agricultural Wastes These wastes hold a major share in wastewater generation by discharging pollutants and sediments into the surface water by periodic storm water. The use of this wastewater contaminates the crops and causes transmission of diseases to humans. The agricultural drainage water contains pesticides, fertilizers, and effluent from manufacturing industry activities and runoff, in addition to sewage effluents and the presence of decaying plant and animal residues that supply the surface water bodies with huge quantities of pollutants.

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Pesticides and insecticides: The pesticides, insecticides, and herbicides percolate in and become a part of soil and water. Pesticides like aldrin/dieldrin, hexachlorobenzene, DDT, chlordane, and benzopyrene produce persistent and recalcitrant dioxin and chlorinated hydrocarbon that tend to persist and accumulate in water bodies and damage beneficial microorganisms. Dioxins generally are unwanted by-products from organochlorine compounds, such as paper production (chlorine bleaching of cellulose) and chlorine alkali electrolysis. (d) Pharmaceutical and Radioactive Wastes The pharmaceutical wastes such as chemicals, medicine waste, plastic wastes, and glass vials expired, contaminated, or spilled in water sources and radionuclidecontaminated solid, liquid and gaseous waste in larger quantity are a major source of wastewater. (e) Other Biological Wastes This waste includes living organisms majorly the microorganisms generated from hospital waste, e.g., patient excreta. It constitutes of various bacteria, fungi, protozoa, and viruses. Septic tank leakage: Septic systems that are damaged or improperly constructed, designed, sited, and maintained can cause contamination with detergents, bacteria, viruses, oils, nitrates, and organic and synthetic chemicals like trichloroethane or methylene chloride.

3 Toxic Metals in Wastewater Environment “Heavy metals” or trace elements are basically defined as the elements with specific gravity >5 g cm_3 (density five times greater to water) in standard state (Holleman and Wiberg 1985). These comprise of a heterogeneous group of elements with varying chemical and biological properties (Mukesh et al. 2008). Their wide application (batteries, chemical compounds, dyes, cosmetic products, alloys, and pharmaceutical) and toxicity effects on plants and human beings increase the pollution risk. Some common heavy metals present in wastewater are arsenic, cadmium, copper, chromium, zinc, nickel, and lead that enter the environment by natural and anthropogenic activities like soil erosion, earth’s crust weathering, mining, urban runoff, sewage discharge, industrial effluents, etc. (Morais et al. 2012). Heavy metals can be classified in four major groups on basis of their requirement and toxicity: (1) essential heavy metals, Cu, Zn, CO, Cr, Mn, and Fe essential for good health and growth in small quantities; (2) nonessential metals, Ba, Al, Li, and Zr; (3) less toxic, Sn and Al; and (4) highly toxic, Hg and Cd even in low quantities. Major heavy metals have been described:

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Fig. 11.2 Lead toxicity in humans

3.1

Lead

Lead (Pb) is a soft, malleable, bluish-white metal tarnished to dull gray upon exposure to air usually found in the +2 oxidation state. Lead is a toxic metal mainly emitted from industrial sources, leaded aviation gasoline, and plumbing materials causing health effects like behavioral problems, learning disabilities, seizures, and death. Lead is widely used in the industries manufacturing batteries, metal products (solder and pipes), ammunition, and X-ray shielding devices. The most common source of lead exposure in humans is ingestion of food, drinking water, soil/dust, or lead-based paint. It damages the nervous connections (especially in young children) and can cause nephropathy and colic-like abdominal pains causing blood and brain disorders in adults or children ultimately causing death, miscarriages in women, and fertility reduction in males (Golub 2005). Chronic exposure to lead can result in dyslexia, mental retardation, muscular weakness, birth defects, paralysis psychosis, autism, allergies, weight loss, hyperactivity, brain damage, and kidney damage and may even cause death (Martin and Griswold 2009). Figure 11.2 shows lead toxicity in human body.

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Fig. 11.3 Cadmium toxicity in human

3.2

Cadmium

Cadmium (Cd) is a silvery-white soft metal with high melting (320.9) and boiling point (765 C), respectively. This heavy metal is rapidly oxidized to cadmium oxide, and so is of great concern being highly toxic to animals and humans. Its natural occurrence is in association with zinc, copper, or lead in sulfide form (Cameron 1992) and is found in argillaceous and shale deposits such as green rocks (CdS) or otavite (CdCO3). The natural sources of Cadmium are volcanoes and anthropogenic being industrial sector, steel plating, nickel-cadmium batteries (Fassett 1980) polyvinyl chloride plastic (PVC) manufacture, motor oils solders, textile manufacturing, electroplating, fungicides, enamels, rubber, sewage, and phosphate fertilizers (Bagshaw et al. 1986) causing severe environmental problems. It has carcinogenic, neurotoxic, and mutagenic effects (shown in Fig. 11.3) as is highly soluble in water, and so human body entry is easy via food chain. Cadmium can cause both acute and chronic intoxications (Chakraborty et al. 2013). It is highly toxic to kidney, causing renal dysfunction, develops osteoporosis (skeletal damage), stomach irritation that results in vomiting and diarrhea. A disease “Itai-itai” (it hurts-it hurts disease) was caused due to cadmium poisoning in Japan in 1950. The rice fields received this water and absorbed cadmium that accumulated in the bodies of people causing coughing, anemia, and kidney failure leading to death.

3.3

Mercury

Mercury (Hg)/quicksilver or hydrargyrum is the only metal in liquid form at standard temperature and pressure with boiling temperature 356.73  C and freezing 38.83  C. It is used in thermometers, sphygmomanometers, barometers, float valves, manometers, electrical switches, and scientific apparatus, though element’s toxicity concerns have led to phasing out of thermometers and sphygmomanometers in clinical environments with alcohol-filled, digital/thermistor-based instruments. It is a potent

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Fig. 11.4 Keratosis caused by arsenic

neurotoxin that affects the brain, liver, CNS, and kidneys and leads to developmental disorders in children. Minamata disease, sometimes referred to as Chisso-Minamata disease, is a neurological disorder caused by severe mercury poisoning, with symptoms including ataxia, hands and feet numbness, muscle weakness, and damage to vision, hearing, and speech leading to paralysis, coma, and finally death.

3.4

Arsenic

Arsenic is naturally present in groundwater that has threatened its use as major source of potable water (Bhutta et al. 2002; Harvey et al. 2002; Smedley and Kinniburgh 2002; USEPA 2003; Mroczek 2005; Mohan et al. 2007). The major sources are anthropogenic, biological, and geochemical reactions that mobilize arsenic to groundwaters. Arsenic can exist in _3, 0, +3, and +5 oxidation states (Smedley and Kinniburgh 2002). It is a hard acid and generally forms complex with oxides and nitrogen, while the environmental forms comprise of arsenious acids, arsenate, arsenic acids, etc. with major form as arsenite (AsO33_)/arsenic (III) and arsenate (AsO43_)/arsenic (V) in water. It is highly toxic and is considered global health problem as it affects millions of people. Arsenic contamination of groundwater and soil is a resultant of mining, milling, smelting of copper, lead ores, coal, and pesticides (Adriano 2001; Ng et al. 2003). The only method to identify arsenic in soils and other media in routine is its extraction by hot acid. Arsenic in water can cause keratosis (shown in Fig. 11.4) and cancer in the skin, kidney, lungs, and bladder when ingested by people at concentrations greater than 50 μg/liter.

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Fig. 11.5 Schematic representation of chromium toxicity

3.5

Chromium

Chromium (Cr) is a hard, steel gray-colored and bright transition metal. It derives its name from Greek word “chroma” meaning color as its compounds exist in different color. This metal is extracted as chromite ore from the mines. It exist in various oxidative states like +1, +2, +3, +4, +5, +6, +2, +3, and +6 being most common. It is majorly used in stainless steel alloys, chrome plating, and metal ceramics. The hexavalent form Cr(VI) is carcinogenic (Mei et al. 2002) and of greatest concern (Fig. 11.5). The anthropogenic sources are greatest contributors, but the toxicity of Cr is of less concern due to poor absorption and translocation of this metal by plants. The major diseases caused are dermatitis, bronchial carcinomas, skin eczema, nasal septum perforation, asthmatic reactions, gastroenteritis, and renal problems.

3.6

Copper

Copper (Cu) is a soft, malleable, and ductile metal that occurs in directly usable metallic form in nature with a very high thermal and electrical conductivity. It is an essential trace minerals; its widespread occurrence in food, water pipe, medicines, and birth control pills increases copper toxicity. It is not poisonous generally, but some salts are poisonous sulfate or the blue vitriol (NilaTutia) and the subacetate or

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verdigris (Zangal). Copper is a powerful enzyme inhibitor, has therapeutic and antiseptic properties but when its salts are retained in the kidney and liver indirectly affects the nervous, reproductive, adrenal, connective tissue etc. can cause mouth, salivation, vomiting, burning pain in stomach, nausea, Hair loss, anemia, autism, anorexia, anxiety, attention deficit disorder, candida overgrowth, arthritis, asthma, depression, male infertility, prostatitis, fibromyalgia, migraine headaches, PMS, chronic infections, etc. SeleniumSelenium (Se) is a rare chalcogen possessing metalloid characteristics of

both metal and nonmetal. It is obtained as mining by-product of metals like copper, iron, and lead ores. The compounds of selenium when inhaled cause respiratory membrane irritation, bronchial inflammation, pulmonary edema, and pneumonia and produce irritation in mucous membrane, vomiting and nausea, cardiovascular effects, nose bleeding, coughing, and ophthalmic irritation.

3.7

Nickel

Nickel (Ni) is a silvery-white, hard and ductile transition metal, which usually occurs in combined form with sulfur (millerite), with iron (pentlandite), and with arsenic (nickeline). It is a ferromagnetic element and is considered very toxic metal. The major sources of Ni are the effluents generated from zinc base casting, silver refineries, storage battery, and electroplating industries. Nickel is also released to the atmosphere by volcanoes, combustion of fuel oil, windblown dust, municipal incineration, nickel refining, and steel production. It adversely affects depending on the exposure route being inhalation, oral or dermal causing systemic, developmental, immunologic, carcinogenic, neurologic, reproductive effects. Dermatitis (Ni itch) is caused by frequent exposure to Ni, such as in coins and jewelry. Higher concentration of Ni causes bone, lungs and nose cancer, dermatitis, acute poisoning causes headache, chest pain, dizziness, cyanosis, nausea and vomiting, tightness of the chest, dry cough and shortness of breath (Duvnjak and Al-Asheh 1997).

3.8

Silver

Silver (Ag) occurs as elemental silver and monovalent silver ion +l, 2, and +3 oxidation state naturally. Silver is used in the manufacture of formaldehyde, acetaldehyde, and higher aldehydes by the catalytic dehydrogenation of the corresponding primary alcohol and photography industry. When silver-containing dust is inhaled, it is most toxic causing “argyria.” Silver dust precipitates in the tissues and cannot be eliminated from the body and can cause skin and nasal septum ulcers.

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Zinc

Zinc (Zn) is a lustrous bluish-white metal brittle and crystalline at ordinary temperatures, but it becomes ductile and malleable when heated between 110  C and 150  C. The major sources are industrial activities, such as mining, steel processing, coal and waste combustion (Raut et al. 2012), galvanization, smelting, paint, fertilizers, pesticides, batteries, fossil fuel combustion, etc. (Holdren et al. 1991). Zinc is generally considered nontoxic, but overdose toxicity causes symptoms like nausea, lethargy, vomiting, epigastric pain, and fatigue.

3.10

Aluminum

Aluminum (Al) is the third most abundant metallic element and constituting about 8% of the Earth’s crust present in only one oxidation state (+3) in the environment. It occurs naturally as silicates, oxides, hydroxides, organic matter complexes, and complexes with other elements like sodium and fluoride in the environment. Toxicity of aluminum is highly influenced by pH of water and organic matter and increases with a decrease in pH (Jeffrey et al. 1997). The main entry modes in humans are through inhalation, ingestion, and dermal contact with the sources being drinking water, beverages, and aluminum-containing drugs and food. A higher dose causes nausea, skin rashes, mouth ulcers, diarrhea, skin ulcers, vomiting, and arthritic pain in humans as Mg2+ and Fe3+ are replaced by Al3+ and osmoregulatory failure in aquatic animals by damaging the plasma and hemolymph ions which causes many disturbances associated with intercellular communication, cellular growth, secretory functions, and secondary hyperparathyroidism, leading to other diseases such as aluminum-induced adynamic bone disease and aluminum-induced osteomalacia.

3.11

Iron

Iron (Fe) being the second most abundant metal on the earth’s crust is a very essential element for all living organisms (Valko et al. 2005). It is a vital component of algae, enzymes, and oxygen-transporting proteins (hemoglobin and myoglobin) and exists in ferrous (Fe2+) and ferric (Fe3+) oxidation states. The major anthropogenic sources are iron and steel industry, iron mining, fertilizer, and herbicide. A higher dose affects gastrointestinal tract and biological fluids with free ions penetrating into cells of the heart, liver, and brain.

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Fig. 11.6 A wetland ecosystem

4 Wetland Ecosystems: Introduction and Basic Characteristics The ability of wetland ecosystems to act as a source or sink for conditioning of nutrients and carbon has led their widespread use in the environment. These possess bunch of resources that are highly important for plant and animal life. They exist either as natural or constructed wetland form and further characterized by several factors like presence of water, presence of vegetation, and nature of soil (Cheng et al. 2002). These are best reserved for finishing of prior partially treated industrial or domestic waste with specific pollutants removal (pesticides, nitrogen, phosphorus, selenium, copper, lead, organic compounds, viruses, or protozoan cysts) from waste material. The plants present in wetland facilitate the major mechanisms for removal of contaminants; process called phytoremediation has been developed as a costeffective and environment-friendly method for remediation of contaminated water.

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Wetland Systems

Natural wetlands are the areas of transition between terrestrial and aquatic systems (Fig. 11.6). Previously, depending on plants present, wetlands have been called swamps, bogs or ecotones, sloughs, and marshes (Fig. 11.6). Wetlands can be defined as: An area where the water table is at or near the surface/land that is covered by shallow water supported predominantly by hydrophytes (water-tolerant plants) and consisting of undrained hydric soils wet enough for time period to produce anaerobic conditions limiting the variety of plants growing on them.

Wetland support vegetation and have higher biological activity, act as groundwater recharge and discharge, and stabilize and protect the shores of seas, riverbanks, and lakes from erosive tides, floods, storms, winds, and waves.

4.2

Constructed Wetlands

These are the wetlands that are constructed or engineered to treat wastewater and improve water quality (Bharagava et al. 2017c). Their wastewater treatment capability and wide range of applications have attracted researches all over the world to explore this technology to treat domestic and industrial wastewater, agricultural flows, landfill leachates, etc. In a constructed wetland, the three main compartments, i.e., substrate (sediment), hydrology, and vegetation, play an important role in the remediation of toxic metals or other pollutants. The substrate harbors consortia of detritus microbes, which are responsible for toxicity reduction and transformation, which may cause the biomineralization. The physical processes in constructed wetlands through which these metals are removed from contaminated waters include settling and sedimentation following the adsorption on particulate matter. To some extent, mats of floating plants serve as sediment traps, while rooted and floating plants regulate the flow rate of water by providing the resistance. In sediments, the metals are adsorbed onto the particles by either cation exchange or chemisorptions. There are two major types of constructed wastewater wetlands shown in Fig. 11.7: 1. Free water surface wetland (FWS) resembles natural wetlands and encompasses shallow water flowing over plant media with water depths varying all over wetland. 2. Vegetated submerged bed wetland (VSB) constitutes of coarse substrate media like gravel through which water travels. They remove contaminants by different means but with common mechanisms and basic processes. Figure 11.7 below depicts the types of constructed wetlands. For treatment in wastewater wetlands, there are three basic mechanisms: Biological, physical, and chemical. In the traditional physical method of wastewater treatment, plants in the wetland help

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Fig. 11.7 Types of constructed wetland

trap sediment with low flow that allows settlement of particles (DeBusk 1999) under influence of gravity and relative densities of suspended material. In the chemical process, various mechanisms like photooxidation, volatilization, and sorption are involved. In biological methods, plants uptake contaminants directly into the root structure, called phytodegredation, while when they secrete substances to aids biological degradation, it is called rhizodegradation. The transpiration of contaminants through the plant leaves is called phytovolatilization (Interstate 2003). Also the microorganisms in soil of wetland store nutrients, but organic pollutant removal is crucially driven by the metabolic functions.

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5 Intricate Mechanisms for Metal Removal in Wastewater by Wetland Technology The metals copper, selenium, and zinc are required for proper growth and functioning of plant and animal in small amounts. In higher concentrations some metals turn toxic and cause health hazards, i.e., lead, cadmium, and mercury, generally found in industrial. The pollutants are removed by physical, chemical, and biological processes in a wetland. We have to understand the basic processes of the removal and potential applications and knowledge of benefits along with limitations of these treatment systems to design constructed wetlands.

5.1

Physical Processes

In a constructed wetland, various processes like plant uptake, filtration, chemical transformation, and adsorption can take place for removal of heavy metals from water (Walker and Hurl 2002); however, the primary process considered is sedimentation with suspended particles. As heavy metals come in contact with wetland, a number of removal processes may occur (Zoppou 2001). During sedimentation, the surface water velocity turns very low through the wetland due to plants and rate of suspended particles affecting the sedimentation rate being proportional to the settling velocity of particle and water residence time (DeBusk 1999). Another process, flocculation, which is a physical process, is influenced by pH, ionic strength, suspended particles, and microorganism concentration (Droppo et al. 1997; Matagi et al. 1998; Sheoran and Sheoran 2006). Sedimentation is generally considered as a reversible process that causes accumulation of particles and association of contaminants on the wetland soil/sediment surface and releases them with the environmental change (DeBusk 1999). The turbulence should be controlled as the sediment can resuspend and reenter the waters by wind-driven turbulence and bioturbation (disturbance by animals and humans).

5.2

Biological Processes

This is the most important method for contaminant removal in wetlands as plant uptake is considered the most widely recognized biological process. The plants can absorb pollutants directly in water or supply oxygen to microorganisms around the rhizosphere within the wetland.

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Metal Mobility in the Rhizosphere

Rhizosphere plays a major in the wetland treatment system. In sediments, wherein reducing condition the metal pollutants (Fe and Mn) are generally associated with carbonates and sulfides, the oxidation process aids in remobilizing these pollutants resolving the metals in wetland sediment and accumulating in surface water (Lacerda et al. 1993; Weis and Weis 2004). The combination of acute microbial activity with oxygen in wetland sediments creates anaerobic and aerobic zones, so that both reduction and oxidation reactions take place together in soil and plant interface (Sobolewski 1999). Also the microbial symbionts such as mycorrhizae present in wetland can affect the accumulation of metals (Weis and Weis 2004) by increasing the surface area absorption by root hairs and providing an interface between soil and root hairs.

5.2.2

Metal Uptake by Plants

The uptake and accumulation of metals by plants is element and plant specific. They are generally accumulated initially in root tissues and then in shoots. There are various factors that influence metal accumulation like temperature, metal concentrations, pH, and nutrient levels in the wetland surrounding. Zn can be accumulated in leaves, while lead accumulates in roots and shoots (Weis and Weis 2004). The heavy metals accumulated can be essential micronutrients such as Zn, Mn, Ni, and Cu and nonessential toxic heavy metals, such as Cd, Pb, As, and Hg (Kabata-Pendias and Pendias 2001; Papoyan et al. 2007). Besides, pH, temperature, and other factors, such as sediment organic matter content, microbial biomass, grain size, ions, and nutrients, may also affect the metal uptake process by wetland plants (Dong et al. 2007; Reboreda and Caçador 2007).

5.3

Chemical Processes

There chemical processes that cause the heavy metal removal in wetlands are described below.

5.3.1

Sorption

It is considered the most important and efficient process in wetland soils/sediments, for chemical removal of contaminants by two processes: (1) short-term retention and (2) long-term immobilization. In this process ions are transferred from the solution phase (water) to the solid phase (soil/sediment) by adsorption and precipitation reactions (Sheoran and Sheoran 2006). The capacity of soils/sediments (cation

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exchange capacity) to retain these cations in wastewater, such as ammonium ion (NH4 +), Cu2+, Cd2+, Pb2+, Zn2+, and Ni2+, increases with substrates like clay colloids and organic matter (Locke et al. 1997). The metal ion adsorption depends upon the medium which indirectly affects the metal solubility and binding capacity onto medium surface. This process is suitable for Cu that predominantly exists as an organic compound and carbonates in leachates of road dust, Mn, Zn, and Cd that exist in the form of ions and carbonate.

5.3.2

Precipitation and Coprecipitation

These two mechanisms are majorly used for removal of heavy metals in wetlands by sedimentation-precipitation-adsorption phenomena (Yao and Gao 2007). For example, Pb, As, Cu, Zn, and Cd can be coprecipitated with pyrite releasing insoluble sulfides available to biota (Morse 1994). Acidic conditions are needed for removal of cationic metals such as As, Sb, and Se; Fe facilitates adsorption and alkaline conditions for cationic metals like Cu, Cd, Zn, and Ni for coprecipitation.

5.3.3

Oxidation and Hydrolysis of Metals

The sediments sometimes act as a source of pollutants when there is a decrease in pH or redox potential and toxic metals may be released from anoxic sediments causing drying/aeration of the wetland sediment with exposure to air (van den Berg et al. 1999; Wilson and Chang 2000; Hartley and Dickinson 2010). The degradation rate of organic matter increases with high oxidization rate of metal sulfides due to increase of redox potential in sediment, accelerating the rate of heavy metals adsorption. For example, in case of a stable Cd compound, an increase of redox potential in sediment causes a decrease in its quantity from 65% to 30% will form a more labile mobile (Zoumis et al. 2001; Kelderman and Osman 2007; Peng et al. 2009). Acidithiobacillus ferrooxidans can catalyze the oxidation of ferrous to ferric iron under acidic conditions. The reaction can be expressed as: 4Fe2+ + O2 + 4H + ! 4Fe3+ + 2H2O. There will be a decrease in sediment pH due H+ ions release into pore water sediment and affect the iron hydroxide solubility and cause a secondary release of heavy metals (shown below in reaction) (Kusel 2003; Hartley and Dickinson 2010). Some of this release material will be re-adsorbed onto the mobile binding compounds: 4FeS2 + 15O2 + 14H2O 4Fe(OH)3 + 8SO42
+ (Fe(OH)3 + 3H + Fe3+ + 3H2O) 16H+.

5.3.4

Metal Carbonates and Sulfides

The heavy metals may be removed in the form of carbonates with their hydroxides as the suspended particles in wetland adsorb heavy metal carbonates and hydroxides both forming their components. Most of the metal cations may combine with CO32–

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and S2– to form compounds of slightly soluble carbonates and sulfides, and the carbonates being less stable than sulfides may be transformed to more stable forms (ITRC 2003; Sheoran andSheoran 2006). In case of Cu and Mn carbonate, accumulation in a natural wetland sediment Cu forms very insoluble compounds with sulfion; cupric (CuS) and cuprous (Cu2S) sulfides may be ambushed by complexation with plant litter or organic matters, while Mn may be precipitated as metal carbonates and sulfides (Dulaing et al. 2006).

6 Phytoremediation Technologies Phytoremediation is an economically viable integrated technology in which green plants and their associated rhizospheric microorganisms cause detoxification, degradation, and removal of chemical pollutants from the contaminated site (Saxena et al. 2019; Saxena and Bharagava 2017; Bharagava et al. 2017b; Chandra et al. 2015). The term phytoremediation is formed from the Greek word “phyto” meaning “plant” and the Latin word “remedium” meaning “to heal again.” The various plantbased technologies for metal decontamination include stabilization, extraction, volatilization, and rhizofiltration. The efficiency of this treatment process is governed by various soil and plant factors such physical and chemical properties of soil, plant, and microbial exudates, plant’s uptake ability, detoxification, metal bioavailability, sequestration, accumulation, and translocation of metal quantity. The selection of plants for phytoremediation is a difficult task, and generally native plants are preferred as they offer less competition among them under certain environmental conditions, and they should be fast growing, have high biomass, and have the ability for heavy metal hyperaccumulation, high salt concentration tolerance, and efficient metal translocation to aerial parts (Vangronsveld et al. 2009; Sharma and Johri

Fig. 11.8 Process of phytodegradation

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2003). This technology has cumulated increased scientific and commercial interest as it is environmentally compatible and less expensive contaminated site remediation method in comparison to other engineering-based methods such as excavation, soil washing, or soil incineration (Clayton 2007). Below are described the various phytoremediation technologies.

6.1

Phytodegradation

Phytodegradation or phyto-transformation is a process of reduction of contaminants through metabolic processes occurring in the plant or effect of compounds called enzymes released by the plants. In this the complex organic pollutants are disintegrated into simpler molecules and are incorporated into the plant tissues to aid in faster plant growth (Fig. 11.8). The enzymes present in plants catalyze and accelerate the chemical reactions, with some involved in conversion of ammunition wastes, few in degradation chlorinated solvents such as trichloroethylene (TCE), and others causing herbicide degradation.

6.2

Phytoextraction

This process relies on the plants that accumulate metals and transport them into the harvestable aboveground shoots (Salt et al. 1998; Vassil et al. 1998). The plant material finds subsequent use as nonfood (e.g., wood, cardboard) or ash, followed by landfill disposal or recycled in case of valuable metals called phytomining (Chaney et al. 1997). The species mainly used for phytoextraction are Indian mustard and sunflower as they grow faster with high biomass, high salt tolerance, and accumulation of metals (Blaylock and Huang 2000; Salt et al. 1995). It is considered a low impact technology and beneficial for environment. Furthermore, there is reduction in leaching and erosion as plants cover the soil. The amount of contaminants can be reduced in soil with successive cropping and harvesting (Vandenhove et al. 2001).

6.3

Phytostabilization

Phytostabilization, also imputed as in-place inactivation, is basically used to remediate soil, sediment, and sludge (United States Environmental Protection Agency 2000). In this method plant roots reduce the mobility of contaminant and its bioavailability in the soil. It focuses mainly to (1) decrease the water percolation through the soil matrix that otherwise may form hazardous leachate, (2) act as a barrier to avoid direct contact with the contaminated soil, and (3) check soil erosion and the toxic metal distribution to other nearby areas (Raskin and Ensley 2000). The

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various processes that take place are sorption, complex formation, precipitation, and metal valence reduction and treat sites contaminated with lead (Pb), copper (Cu), arsenic (As), chromium (Cr), cadmium (Cd), and zinc (Zn) and mining-affected and superfund sites.

6.4

Rhizofiltration

Rhizofiltration also called phyto-stimulation is a method for remediation of wastewater, extracted groundwater, and surface water with low contaminant concentrations (Ensley 2000) by use of plants roots (terrestrial and aquatic) for absorption, concentration, and precipitation of contaminants from polluted aqueous sources. It may be used to retain for Pb, Ni, Cd, Cu, Zn, and Cr within the roots. It is advantageous in its ability to undergo either in situ or ex situ process by utilizing both terrestrial and aquatic plants, and the contaminants do not need to be translocated to the shoots. Due to the fibrous and much longer root system, the terrestrial plants are preferred as they also increase root area (Raskin and Ensley 2000).

6.5

Phytovolatilization

This process involves the transformation and transpiration of contaminants from soil into the atmosphere by the use of plants to take into volatile forms. It is primarily used for mercuric mercury. It could be transformed into less toxic form or in reverse may be disadvantageous by being redeposited back into lakes or oceans, and reproduction of methylmercury occurs by anaerobic bacteria. Members of the Brassica genus and some microorganisms are particularly good volatilizers of Se (Terry et al, 1992). Among the aquatic species, rice, rabbit-foot grass, Azolla, and pickleweed are the best Se volatilizers (Hansen et al. 1998; Lin et al. 2000; PilonSmits et al. 1999; Zayad et al. 2000).

7 Future Perspective and Challenges Wetlands being the most important biodiversity areas in the world vary widely because differences in locality and region, vegetation, topography, hydrology, and human involvement are experiencing a rapid decline. We can meet many future challenges like food and water security, climate change resilience, human health, and disaster risk reduction that can be checked by conservation and sustainable use of wetland. Constructed wetlands are considered an efficient technology for treating wastewater and removing organic matter, nitrogen, suspended solids, phosphorus,

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pathogens, and metals. The application of this environment-friendly technology to treat commercial wastewater could signify a step toward “green technology.” There is settlement of pathogens in slow-moving water of constructed wetlands, and high percentages of fecal coliform and other pathogens are removed. Proper plant species should be selected for wastewater wetlands to achieve relatively high percentages of metal removal due to soil adsorption and precipitation mechanisms. An understanding of contaminated removal helps in taking proper decisions for implementation of wetlands. Research has found that constructed wetlands being a less complicated technology make a good secondary treatment method for domestic wastewater and offer aesthetic pleasing environment. There should be continuous performance monitoring of existing constructed wetlands to optimize design and minimize construction cost. The support of local governments and international organizations involved in water and wastewater sector is important for building local capacity and scaling up application of this technology. Acknowledgment This book chapter would not have been possible without support from "Springer Singapore" published under Springer Nature Singapore Pte Ltd. We are especially indebted to Dr. Ram Naresh Bharagava and Mr. Gaurav Saxena for providing us with this oppurtunity. We wish to place on record the valuable supervision rendered by the them in reviewing and editing the chapter.

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Chapter 12

Microalgae: An Eco-friendly Tool for the Treatment of Wastewaters for Environmental Safety Jae-Hoon Hwang, Anwar Sadmani, Seung-Jin Lee, Keug-Tae Kim, and Woo Hyoung Lee

Abstract Algae-based wastewater treatment can provide renewable biomass generation for sustainable bioenergy production while treating wastewater as a growth medium for algae cultivation. In addition, algae are excellent at sorbing and/or degrading inorganic materials (e.g., heavy metals) and emerging contaminants (e.g., endocrine-disrupting chemicals (EDCs)), indicating that utilizing an algaebased treatment process is one of emerging strategies for advanced wastewater treatment as an eco-friendly way. Economic advantages and environmental safety associated with algae-based wastewater treatment also constitute a driving force for its utilization in biofuel feedstock generation or fertilizer production. This chapter discusses the principles and rationale for algae-based wastewater treatment coupled with biodegradation of wastewater and renewable energy production. Several biomass technologies for energy production are proposed, which improve the economic feasibility of algal biofuel production. The integration of membrane bioreactors with algae cultivation is also addressed. A new method with separated trophic conditions, enhanced algal nitrification process (EANP), is introduced for practical applications. It seems that pretreatment of raw wastewater and separated culture condition is required to overcome the challenges of scale-up and enhance nitrification rates. Furthermore, synergistic coupling of the microalgae production via advanced wastewater treatment is highlighted in the context of sustainability benefits. J.-H. Hwang · A. Sadmani · W. H. Lee (*) Department of Civil, Environmental and Construction Engineering, University of Central Florida, Orlando, FL, USA e-mail: [email protected]; [email protected] S.-J. Lee Department of Computer Science, Engineering, and Physics & Department of Geography, Planning, and Environment, University of Michigan-Flint, Flint, MI, USA e-mail: sjleeum@umflint.edu K.-T. Kim Department of Environmental & Energy Engineering, University of Suwon, Hwaseong-si, Gyeonggi-do, Republic of Korea (South Korea) © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_12

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Keyword Biodegradation · Biofuels · Biosorption · Enhanced algal nitrification process (EANP) · Environmental safety · Life cycle assessment (LCA) · Microalgae · Wastewater

1 Introduction With greater understanding of the impact of wastewater on the environment and more, a sophisticated advanced technology, an eco-friendly tool for sustainable wastewater treatment is highly demanded. Industrial wastewater is a major source of pollution as it carries a variety of highly toxic organic and inorganic chemicals which may cause serious toxicity in living beings upon exposure (Goutam et al. 2018; Bharagava et al. 2017a, b). Biological treatment processes are relatively natural processes in which beneficial microbiological agents (e.g., bacteria) or plants (i.e., phytoremediation) are utilized to treat contaminated water (Saxena et al. 2016, 2018; Gautam et al. 2017; Bharagava et al. 2017c; Saxena and Bharagava 2015, 2017; Chandra et al. 2015). Wastewater treatment using green algae is an innovative technology with several benefits, which can mitigate greenhouse gas emissions and produce clean and safe water with less energy consumption (Birol 2007). Algal biodegradation is an eco-friendly, cost-effective, highly efficient approach compared to traditional physicochemical methods (e.g., chemical coagulation, filtration, ion exchange, and activated carbon adsorption) which are expensive as well as unfriendly towards the environment (Javaid et al. 2016). Microalgae are capable of rapid growth in nutrient-rich wastewater under light with improved removal efficiency compared to existing biological nutrient removal processes (e.g., 90% nitrogen [N] and phosphorus [P] removal within 2 days for microalgae compared to 90% of N and 80% of P removal in A2O processes) (Jia and Yuan 2016). They can also sorb inorganic and organic nutrients such as EDCs (Zhou et al. 2014) and heavy metals (Mallick 2002) from wastewater, representing an environmental safety tool for advanced wastewater treatment. Several algal systems of different configuration such as stabilization ponds, hyper-concentrated cultures, and immobilized cell systems use wastewater for algae cultivation; however, they may generate secondary pollutants such as hydrocarbons (HC) and nitrogen oxides (NOx) and require relatively large area of operations. Algae-integrated wastewater treatment processes emphasize on bioenergy production such as biodiesel (Batan et al. 2010), alcohol (e.g., butanol and ethanol) (Choi et al. 2011) and biohydrogen production (Hwang et al. 2014), and capture/sequester CO2 which is one of the primary greenhouse gases (GHG). A study by Kim et al. (2013b) showed 60% of biodiesel production cost saving by downstream processes for algae cultivation. It is suggested that reducing nutrient cost for algae cultivation by using wastewater ensures the economic feasibility of microalgal biofuel production. However, despite inherent potential as a biofuel resource, there are still many challenges to be resolved which have impeded the development of algal biofuel technology at commercial scale. For

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example, although removal of nutrients by microalgae has been studied for decades, scale-up of algae-based wastewater treatment is still technically challenging. Currently, it seems that high rate algal ponds (HRAPs), waste stabilization ponds (WSPs), and algal turf scrubbers (ATSs) are available technologies for scale-up. However, long water retention time requiring large land area is a major drawback of these technologies. This chapter provides information on state-of-the-art technologies and environmentally safe tools for wastewater treatment and subsequent biofuel production from microalgae. Knowledge gaps are also identified, and future directions of eco-friendly tools using algae for sustainable wastewater treatment are addressed.

2 History of Microalgae for Wastewater Treatment The history of the commercial use of microalgae for wastewater treatment started 75 years ago in the early stage of wastewater treatment processes (Abdel-Raouf et al. 2012). A low-cost and environment-friendly algal treatment is an effective method to treat or remediate the industrial wastewater with high N and P concentrations, and some of algal species can assimilate heavy metal pollutants and toxic organic chemicals to some extent (De-Bashan and Bashan 2010) (Table 12.1).

2.1 2.1.1

Biodegradation of Nutrients Removal of Nitrogen and Phosphorus

Eukaryotic algae have shown their ability to remove N and P, about 90% of inorganic N and 80% of inorganic P, from wastewater. Many parameters (e.g., biotic factor, pH, temperature, and light) affect algal nutrient removal, and thus nutrient removal efficiencies are different depending on wastewater characteristics. In addition, the initial biomass concentration is critical for designing the algae-based wastewater treatment.

2.1.2

Reduction of Biochemical and Chemical Oxygen Demands

Microalgal species are used to remove organic matters (i.e., biochemical and chemical oxygen demands [BOD and COD]) in wastewater (An et al. 2003; Kong et al. 2010). Improved BOD and COD removal efficiency by algae, which is found commonly, is associated with food/microorganism (F/M) ratio, typically optimum F/M ratio range was 0.05–0.1 (Choi and Lee 2012); however, F/M ratio over 0.2 decreases BOD and COD removal less than 80%. It is suggested that F/M ratio can be a key parameter for algal biodegradation of organic matter in wastewater.

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Table 12.1 Summary for algal treatment on bioremediation of nutrient in different wastewater Wastewater Secondary effluent Secondary effluent

Algal species Chlorella vulgaris

Piggery wastewater Swine wastewater

C. zofingiensis

Municipal wastewater Municipal wastewater

Chlorella sp.

Phormidium sp., C. vulgaris Chlamydomonas reinhardtii, and Spilarctia rubescens

C. sorokiniana

C. vulgaris

Artificial wastewater

C. vulgaris and Chlamydomonas mexicana

Raw wastewater

C. reinhardtii

Scenedesmus obliquus

Chlorella pyrenoidosa

C. vulgaris

Industry effluent

Anabaena subcylindrica

Nutrient treatment removal efficiency 99% N and P removal, reaching 20%). The photobioreactor study of Spirulina on the removal of nutrients from tannery effluent, reported the maximum COD of 91%, nitrates of 97%, and phosphates of 100% with 3% loading rate. These outcomes are due to higher volumetric loading rates, where the microalgae were exposed to higher ammonia levels and organic matter when compared to diluted effluent (Abeliovich 1983). In order to function as a growth medium, the microalgae growth required a reduction in organic and toxic load from tannery effluent through aerobic or anaerobic treatment. It was evident from the large-scale investigation on the high-rate algal pond (HRAP) integrated with the anaerobic waste stabilization pond. The study unveiled a maximum organic load (COD) reduction of about 84%, settleable solids of 99%, and the removal of ammonia, phosphate, and sulfides of about 95%, 99%, and 99%,

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respectively. Furthermore, the addition of recirculated carbonate-rich effluent to HRAP for pH adjustment reported an increase in biomass productivity of Spirulina from 8 to 11 g m 2 day 1 (Dunn 1997). Scenedesmus sp. used for the treatment of wet blue tannery effluent reduced the heavy metals pollution load (Cr-81.2, 96%; Cu-73.2, 98%; Pb-75, 98%; and Zn-65, 98%) and nutrients (NO3>44.3% and PO4 >95%) (Ajayan et al. 2015). In general, metal ions and other nutrients enter via biological metabolism across the cell membrane into the cell. This is known as active uptake mechanism, whereby these nutrients are assimilated by microalgae and metal ions are accumulated. FTIR analysis of Scenedesmus sp. biomass confirmed the involvement of hydroxyl amino, carboxylic, and carbonyl groups for heavy metal adsorption on the cell wall.

3.6.2

Carpet Mill Wastewater

Combining industrial wastewater with municipal wastewater enhances the potential of microalgae growth to produce valuable biomass. A study was conducted in the United States, specifically Dalton (the carpet capital of the world) in Georgia which produces about 100 to 115 million liters of wastewater per day. A consortium of 15 native species isolated from the carpet mill wastewater used to treat the wastewater consisting of 85–95% carpet mill effluent and 10–15% of municipal sewage could remove more than 96% of nutrients (phosphate, 96.6–99.1%, and nitrate, 99.7–99.8%). Further, 9.2 to 17.8 tons ha year 1 of biomass production and 6.82% of lipid content were obtained in this study. It was further reported that 63.9% of algal oil obtained from the survey could be utilized to produce biodiesel (Chinnasamy et al. 2010).

3.6.3

Acid Mine Industry

Wastewater from the mining industry possesses a threat to the environment especially in countries where this industry generates the primary income. Acid mine drainage (AMD) refers to the wastewater produced by this economic activity, and a pilot-scale experiment with a 1 cubic meter treatment cell was used for the treatment. A substrate consisting of woodchips, soil, and powdered goat manure was used to trap the cyanobacterial–microbial consortium which forms a microbial mat; this occurs mainly due to the interaction of bacteria present and filamentous blue-green algae, mostly Oscillatoria sp. The maximum removal percentages of metals, 59–83% of Co, 79–97% of Cu, 95% of Fe, 28–45% of Mn, 22–62% of Ni, 88% of Pb, and 84–86% of Zn, were achieved in the treatment system (Sheoran and Bhandari 2005).

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Limitations

Phycoremediation of wastewater treatment has its limitations despite its simplicity, effectiveness, and eco-friendly nature. These are explained in more detail below: • Raw sewage has been studied broadly in the fields of microalgae culturing and wastewater treatment. It is considered to be a suitable nutrient medium because of its N/P ratios availability and low levels of toxicity (Li et al. 2011). Even though agricultural wastewaters (such as manure-based wastewaters) are regarded as a potential nutrient source for microalgae, the presence of pesticides in agricultural wastewaters may be detrimental to a few microalgae species. Cyanobacteria species are appropriate for pesticide bioremediation (usually at trace concentrations), but again the toxic cyanobacteria create a problem for the environment (Caceres et al. 2008). • In the case of effluent containing high organic and nutrient load (e.g., wastewater from livestock, soybean processing, breweries, etc.), dilution is required prior to microalgal treatment in order to prevent unwanted inhibitory effects of ammonium levels (Abeliovich and Azov 1976). At high nutrient concentrations, the formation of chlorophyll a is excessive in microalgae, and it could have the effect of reducing the amount of nutrients removed due to the limited amount of light (4100 lux) (Aslan and Kapdan 2006). • Though microalgae have demonstrated the ability to remove heavy metals (absolute concentration) from industrial effluents and simulated industrial effluents, a high level of heavy metals can be harmful to algae growth. Also, some metal ions (e.g., Fe3+, Cu2+, and Zn2+) may precipitate phosphorus and reduce its availability to microalgae (Zhen-Feng et al. 2011). • Industrial effluents which are characterized by low nutrients and high toxic levels require either added external nutrients or combined toxic industrial effluents with municipal wastewater to support the microalgae growth (Chinnasamy et al. 2010). • Microalgae can be cultivated either by batch mode or chemostat (continuous) mode in which the continuous mode provides maximum efficiency compared to that of batch mode since the nutrients are supplied continuously to the treatment system. There are a twofold rise in biomass productivity, greater N and P removal (> 99%), and higher bioenergy production (5.3–6.1 MJ m 3 day 1) from 1.33 days culturing of Scenedesmus sp. in municipal wastewater effluent (McGinn et al. 2012). In spite of its advantages over the batch mode, this technique is yet to be implemented in the industrial wastewater system, and most studies have been conducted in batch or semi-batch mode (Kotteswari et al. 2012; Chinnasamy et al. 2010; Solovchenko et al. 2014; Mata et al. 2012). • Phycoremediation requires comparatively high downstream processing costs. Algae harvesting from wastewater after treatment remains a major problem to the relevant industries, due to the size of algal cells and their diluted cultures (most commonly 200–600 mg/l 1) (Uduman et al. 2010). Also, the cost of algae

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recovery has been estimated to be high (approximately 20–30% of the total cost of biomass production) (Grima et al. 2003). • It is hard to maintain a monoculture in the treatment system and microalgae cultivated in a well-organized open pond system with no or less contamination is only having the market value. • The growth rate of heterotrophic algae is comparatively smaller or slower than that of autotrophic or photosynthetic algal culture. Consequently, the process requires much improvement in order for the maximum efficiency to be obtained. However, these limitations can be overcome by adopting automation in microalgae-based treatment in the case of fluctuations in organic and inorganic load. This can be achieved by timely regulation of the light intensity and dilution rate in the treatment system (Solovchenko et al. 2014)

4 Potential Application of Algal Biomass from Wastewaters The algal technologies have made substantial progress in the past few decades. Initially, microalgae aroused much interest in the research community as a renewable source of biofuels owing to their high productivity and significant lipid accumulation potential in a short period of time. The dual role of microalgae, i.e., phycoremediation (for both organic and inorganic pollutants) coupled with energy production, has been well established. Also, the recent advances in eco-friendly harvesting techniques with the use of low-cost green coagulants, electrochemical harvesting, etc. have made the recovery of microalgae biomass very energy-efficient and economical (Ravindran et al. 2016). Although its biomass is used in modern applications such as adsorbent, microbial fuel cell (MFC) with substrate, food additives, fertilizers, bioactive agents in pharmaceutical formulations and oil or energy reservoirs, etc., their application is still under research and development for as-yet unknown products.

4.1

Microalgae Biomass: As Adsorbents

Adsorption is a physical phenomenon, widely applied for the removal of toxic heavy metals and dyestuffs from industrial and municipal wastewater streams. It is a rapid, reversible, economical, and environmentally friendly technology which can employ low-cost materials as an adsorbent. “Biosorption” is a process in which living or dead plants and microorganisms (algae, bacteria, and fungi) are employed for the removal of heavy metals and dyes from an aqueous environment. In microalgae, biosorption occurs through physical sorption or interactions between a component of interest (metal ions/ionic dyes) and functional group present over the cell wall biopolymers of living and dead biomass. The major functional groups which play a significant role as binding sites include carboxylate (–COO ), amide (–NH2), thiols (–SH), phosphate

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Table 13.4 Microalgae biomass as an adsorbent for heavy metals and dyestuffs

S. No 1

Microalgae Ulva lactuca sp.

Heavy metal/ dyestuffs Pb(II)

Adsorption capacity (mg/g of algae) 105.26

2

Chlorella vulgaris

Cd(II)

85.3

3

Spirogyra condensata and Rhizoclonium hieroglyphicum Chlorella minutissima (immobilized)

Cr(III)

14 and 11.81, respectively

Cr(VI)

57.33

5

Spirogyra sp.

Cu(II)

34.94

6

Spirulina platensis

482.2

7

Spirulina platensis

8

Spirogyra rhizopus

Reactive Red 120 (RR-120) textile dye Food dyes acid blue 9 and FD&C red no. 40 Acid Blue 290 and Acid Blue 324

4

1653.0 and 400.3, respectively 1356.6 and 367.0, respectively

Removal efficiency 93 % of 175.6 mg/L Pb(II) – >90% of 100 mg/L Cr(III) > 80% of 100 mg/L Cr(VI) 91% of 40 mg/L Cu(II) 94.4–99.0%

– –

References Bulgariu et al. (2010) Aksu (2001) Onyancha et al. (2008) Singh et al. (2012) Bishnoi et al. (2004) Cardoso et al. (2012) Dotto et al. (2012) Özer et al. (2006)

(PO43 ), and hydroxide (–OH) (Davis et al. 2003). The main advantage of using biomass (nonliving or dead) as the adsorbent is, firstly, its energy-/metabolismindependent adsorption process and, secondly, its ability to proceed rapidly through the combination of the following mechanisms, i.e., complexation, ion exchange, and physical adsorption (e.g., electrostatic). However, the removal efficiency percentage of heavy metal and dyes varies according to the biomass source, initial pH, temperature, metal ion, and dye concentration of the aqueous solution being treated. Examples of microalgae (biomass) contributing to the treatment of heavy metal- and dye-contaminated wastewater are depicted in Table 13.4.

4.2

Sustainable Power Generation Through Fuel Cells: Concurrent Wastewater Treatment

The ability of microalgae to produce oxygen and a prominent level of readily degradable proteins (32%) and carbohydrates (51%) content makes it suitable to generate electricity in fuel cells. Microbial fuel cells (MFCs) are being developed

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and represent yet another promising and challenging technology for the sustainable production of energy. MFCs are bioelectrochemical systems (BESs) which convert biodegradable organic waste including wastewaters and lignocellulosic biomass into electricity through microbial metabolic activity. They are considered mainly for using wastewater as a substrate to accomplish its treatment (mineralization of organic carbon into CO2) and to meet increasing energy demand by generating electricity simultaneously. This surge offsets the costs of wastewater treatment plants’ operations (Lu et al. 2009). Different substrates were explored as a feed which includes various kinds of artificial and real wastewaters and biomass from algae and plants (Pant et al. 2010).

4.2.1

Design Principle and Operation of MFC

An MFC functions like an electrochemical cell in that it consists of anodic and cathodic chambers separated by an ion-selective membrane. In the anodic chamber, microorganisms anaerobically oxidize biomass or organic matter to produce electrons and protons. Electrons then flow through an external circuit to the cathode (flow of current), while protons go through the membrane to the cathode. Electrons and protons are then consumed by an electron acceptor (mainly oxygen) in the cathode to produce water. A variety of electron acceptors includes metal ions like Fe (III), Cr (VI), and Mn (V), inorganic electron acceptors like sulfates, and microorganisms such as denitrifying bacteria and microalgae (Wang et al. 2008; Gouveia et al. 2014). Different combinations of microorganisms used in MFC as the substrate, electron donor, and acceptor are as follows: (i) bacteria as an electron donor in the anode compartment (Xing et al. 2008); (ii) microalgae in both cathode and anode compartments (Gouveia et al. 2014); (iii) bacteria (electron producer) + microalgae biomass (as substrate) in the anode compartment (Strik et al. 2008); and (iv) bacteria (electron producer) in anode and microalgae in the cathode chamber for oxygen production (electron acceptor) (Gouveia et al. 2014). Based on the MFC design, there is the two-chambered (Fig. 13.6a) and singlechambered (Fig. 13.6b) system where the former is the traditional (two compartments) method. The two-chambered system has disadvantages, such as large volume, low power output, and costs for aeration (Lu et al. 2009). Simpler and more efficient MFCs were developed by neglecting the cathode compartment. That is, the single-chambered MFCs consist of a cathode electrode on the membrane directly exposed to the air.

4.2.2

Current and Power Outputs from MFCs with Different Substrates

The generation of current in MFCs is directly related to the ability of the bacteria to oxidize a given substrate and electrons transport to the anode electrode. The current density and the efficiencies in removing pollutants also differ according to the

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Fig. 13.6a Schematic of two-chambered microbial fuel cells.

Fig. 13.6b Schematic of single-chambered microbial fuel cell

experimental conditions, specifically the type of substrate used, inoculum concentration, and MFC set up conditions, etc. Table 13.5 summarizes the power output (current density) from different substrates and source of microorganisms used in MFC studies. For MFC in wastewater treatment, the build-up of biofilm on a large surface area required on the anode and innovations are necessary to build low-cost electrodes that resist fouling.

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Table 13.5 Microbial fuel cells (MFCs) with different substrates and the maximum current produced

S. No 1.

2

3

4

5

6.

Substrate Starch processing wastewater Acetate

Microalgae, Chlorella vulgaris (dry biomass) Macroalgae, Ulva lactuca (dry biomass) Scenedesmus sp. (dry biomass) Scenedesmus sp. (dry biomass)

7

Chlorella vulgaris (dry biomass)

8

Dunaliella tertiolecta (dry biomass)

Max power density (mW/m2) 239.4

Inoculum Source or Oxidizer Bacterial inoculum Starch processing wastewater Bacterial consortium

Electron receptor Air cathode

Chlorella vulgaris

62.7

Primary clarifier overflow of wastewater plant

Air cathode

980

Primary clarifier overflow of wastewater plant Activated sludge as an inoculum

Air cathode

760

Air supply at cathode

1780

Chlorella vulgaris in BG 11 medium Ferricyanide in phosphate buffer

1926  21.4

Activated sludge as an inoculum

Microbial consortia from a municipal sewage sludge digester Microbial consortia from a municipal sewage sludge digester

Ferricyanide in phosphate buffer

Reference Lu et al. (2009) Gouveia et al. (2014) VelasquezOrta et al. (2009) VelasquezOrta et al. (2009) Rashid et al. (2013) Cui et al. (2014)

15.0  0.1

Lakaniemi et al. (2012)

5.3  2.6

Lakaniemi et al. (2012)

Although the power yields from MFCs are currently relatively small, improvements in technology such as the use of new acidophilic microbes or “superbugs” are imminent. These include enhanced electron transport rate and coexistence of methanogenic anaerobic digestion with MFC and will improve the power density output from these systems.

4.3

Other Commercial Applications

Microalgae are considered as existing cell factories that produce several bioactive compounds, for instance, proteins, lipids, carbohydrates, carotenoids, vitamins, antioxidants for health, food and feed additives, cosmetics, fertilizers, and energy.

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Nowadays, applications of microalgae biomass are evident in various commercial sectors like the food industry, aquaculture feed, pharmaceuticals (including skin care), agro-industry, and oil and energy sector.

4.3.1

Microalgae and Food Additives/Nutrient Supplement

Due to the presence of sources rich in crucial nutrients including vitamins A, C, B1, B2, and B6 and niacin, microalgae are considered to constitute a primary source of food. Widely commercialized microalgae used mainly as nutritional supplements for humans and as animal feed additives are Chlorella, Haematococcus, Dunaliella, and Spirulina. The most important microalgae grown for its photosynthetic pigment and beta-carotene (orange dye as a food colorant and as a vitamin C supplement) are Dunaliella. New findings suggest that beta-carotene acts as an anticarcinogenic agent and is also effective in controlling cholesterol and reducing the risk of heart disease. These properties make it a more valuable product, and it is expected to increase people’s demand for it. Arthrospira sp. is another proteinaceous microalga, commercialized for its excellent nutritional value and other possible healthpromoting properties (Spolaore et al. 2006). Chlorella minutissima is a marine algae with a high proportion of polyunsaturated fatty acids (PUFA) and is useful in preventing or treating several diseases (Priyadarshani and Rath 2012).

4.3.2

Microalgae in Cosmetics

Species of Ascophyllum, Alaria, Spirulina, Arthrospira, Nannochloropsis, Chlorella, and Dunaliella are species specifically utilized for cosmetics production as thickening agents, water-binding agents, and antioxidants. Extracts from microalgae are found in skin care products as antiaging cream, sun protection creams, refreshing agents, emollients, anti-irritant in peelers, etc. (Spolaore et al. 2006).

4.3.3

Microalgae as Fertilizers

Microalgae serving as biofertilizers and soil conditioners in agriculture are widely applied in India and Southeast Asia. Cyanobacteria play a vital role in maintaining and building up of soil fertility, functioning as a natural biofertilizer through its ability to fix atmospheric nitrogen. A variety of free-living cyanobacteria have now been identified as active components to benefit crop plants by producing various growth-promoting substances such as vitamin B12 and auxins like indole-3-acetic acid, indole-3-propionic acid, or 3-methylindole. Furthermore blue-green algae such as Nostoc, Anabaena, Tolypothrix, and Aulosira can be employed as inoculants for paddy crops to increase yield and improve the soil physicochemical properties (Priyadarshani and Rath 2012).

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Microalgae as Aquaculture Feed

The primary applications of microalgae in aquaculture are linked to nutrition as sole component/as a food additive to necessary nutrients. Some algae species commonly used for aquaculture are Chlorella, Tetraselmis, Isochrysis, Pavlova, Phaeodactylum, Chaetoceros, Nannochloropsis, Skeletonema, and Thalassiosira. Natural pigments obtained from Dunaliella salina, Haematococcus pluvialis, and Spirulina are used for coloring the flesh, in order to induce the biological activities of prawns, salmonid fish, ornamental fish, etc. A protein-rich diet is supplemented and tested with shrimp through microalgae species Hypnea cervicornis and Cryptonemia crenulata (Spolaore et al. 2006).

4.3.5

Microalgae as Energy Reserve

High lipid and carbohydrate content of microalgae biomass make it suitable for energy extraction. Moreover, the quantity of oil production from microalgae biomass is comparable to the yield of the best oilseed crops. For instance, biodiesel yield from microalgae (containing 30% oil by it) is 58,700 l ha 1 compared with 1190 l ha 1 from rapeseed, 1892 l ha 1 from Jatropha, and 2590 l ha 1 from Karanj (Singh and Gu 2010). Botryococcus braunii, Isochrysis sp., Nannochloropsis sp., Schizochytrium sp., Nitzschia sp., Chlorella sp., and Neochloris oleoabundans are a few examples of microalgae with higher oil content. There are two crucial ways by which microalgal biomass is converted into energy, and these are thermochemical and biochemical conversion processes. In thermochemical conversion, thermal decomposition of biomass into fuel products (such as syngas, bio-oil, charcoal, and electricity) is achieved by various methods, viz., combustion, gasification, and liquefaction. The process of biochemical conversion of biomass into energy fuels (methane, hydrogen, and ethanol) includes anaerobic digestion, alcoholic fermentation, and photobiological hydrogen production. Selecting the conversion process depends on the following factors: type and quantity of biomass feedstock; form of energy; end product required; and economic considerations (Brennan and Owende 2010).

5 Possibilities and Future Needs of Algal Technology in Wastewater Treatment The positive impact of phycoremediation based on the above literature studies indicates that it is an efficient strategy for WWT processes. However, these processes have so far been carried out in pilot-scale tests and are yet to be applied to WWT plants. The continuous flow of water in WWT systems and the demand to treat a million liters per day (MLD) have led to the focus on improving retention

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period to replace conventional systems. Therefore, the application of phycoremediation must have commercialization benefits if it is to be more than superficially successful. Future growth of phycoremediation heavily depends on the generated biomass application since the current technologies do not compensate the environment and economic advantages. We reviewed the advantages and disadvantages of phycoremediation in various wastewater treatment systems. Especially, municipal WWT streams indicate the presence of emerging contaminants and should be studied in more detail so that technology is not transferred from one mode to another. Several reports have proved that raceway ponds are better in comparison with photobioreactors in terms of effective remediation. This is despite the fact that the PBR has produced more biomass, fewer contaminants, and less evaporative losses. Proper design considerations and footprint reduction are the drawbacks to operating PBR in a commercial context. In the meantime, the predicted water insufficiency and climate change have led to make amendments in several environmental policies and encouraged industries to recycle water. Apart from the nutrient removal, there are always constraints especially in industries due to socioeconomic factors. Indicators such as the return on investment, fossil energy returns, and net present value are key issues that stakeholders need to take into account. Due to the limitations noted above, it is difficult to identify the suitable biomass opportunities from WWT systems by algal cultivation. Manninen et al. (2016) showed that biogas from algae did not cover the associated electricity cost. Yet the consumption during cultivation and harvesting can be excluded provided the nutrient recovery efficiency is met. Efroymson et al. (2016) suggested that assessment of socioeconomic indicators has outweighed the negatives of algal-based biofuel systems due to the production of adequate energy security. Regarding land footprint, constructed wetland proved to be better because the area required is less compared to the high-rate algal pond. Overall, better design aspects and isolation of algae from the local environment will help this technology to overcome contrary claims.

6 Conclusions The importance of phycoremediation for energy generation specifically from wastewaters is swiftly improving and becoming much more productive. It is also incorporating hybrid approaches that exist in current technologies to nullify the costs incurred on the energy required. Interestingly, engineering of microalgae to improve lipid production and CO2 fixation has led to developing tailored microbes susceptible to the environment. This is an expanding area of research in WWT operations. The data that are now accessible in scientific forums readily offers sufficient technical knowledge that the general public is better able to understand. Furthermore, ongoing research facilitating long-term field applications warrants attention being paid to phycoremediation. The immediate effect of advancing scientific knowledge and mainly microbial tailoring is likely to lead to decisive

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shifts in environmental policies and their subsequent outcomes and implications. Nevertheless, sustainable practices will continue to be crucial for a fruitful and safe natural environment.

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Chapter 14

Pulp and Paper Mill Wastewater: Ecotoxicological Effects and Bioremediation Approaches for Environmental Safety Izharul Haq and Abhay Raj

Abstract Pulp and paper industry is one of the important industrial sectors in India, which consume huge amount of water in the papermaking process. The final wastewater is often characterized by high color, BOD (biochemical oxygen demand), COD (chemical oxygen demand), AOX (adsorbable organic halides), SS (suspended solids), TDS (total dissolved solids), phenolics, heavy metals, and plant components like lignin, tannin, resin acids, and extractives. Finally, these compounds are reached to aquatic and terrestrial ecosystem and causing serious environmental pollution. The generated wastewaters are treated by conventional biological treatment like activated sludge process (ASP) after primary treatment. Biological treatment of paper mill effluent significantly removes BOD, COD, SS, and also COD, but it is insufficient in removal of lignin and chlorophenols due to its low biodegradability and toxicity. During last few decades, several physical and chemical methods have been developed with the aim to use as pre- and posttreatment method. However, application of this technology at large scale is costly. Therefore, bioremediation which involve the use of pollutant-specific microorganism for wastewater treatment has been considered as cost-effective and eco-friendly treatment method. Thus, this chapter provides the updated information on paper processing and wastewater generation and their characteristics and toxicity. Processes based on physicochemical and biological methods for the treatment of pulp and paper mill wastewater have been also discussed. Keywords Pulp and paper mills · Environmental pollution · Toxicity · Wastewater treatment · Bioremediation I. Haq Environmental Microbiology Laboratory, Environmental Toxicology Group, CSIR-Indian Institute of Toxicology Research (CSIR-IITR), Lucknow, Uttar Pradesh, India Department of Civil Engineering, Indian Institute of Technology Guwahati, Guwahati, Assam, India A. Raj (*) Environmental Microbiology Laboratory, Environmental Toxicology Group, CSIR-Indian Institute of Toxicology Research (CSIR-IITR), Lucknow, Uttar Pradesh, India e-mail: [email protected] © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_14

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1 Introduction Environmental pollution is a serious problem that people’s facing worldwide and its protection is one of the challenges required for the survival of living organisms. Our natural environment is being heavily contaminated by industrial wastewaters that contain potentially toxic organic and inorganic heavy metals causing serious environmental threats (Saxena and Bharagava 2015; Kumari et al. 2016; Saxena et al. 2016; Haq et al. 2016a, b; Bharagava et al. 2017a, b, c; Gautam et al. 2017; Goutam et al. 2018; Kishor et al. 2018). The pulp and paper industry is one of the imperative industrial sectors of the country, but unfortunately it discharges huge volume of high-strength wastewater containing many toxic pollutants that may cause adverse impact on environment. To cope with these challenges, in the last decade, various projects have been implemented in Indian paper industry to upgrade the technology for the improvement of energy and water consumption and wastewater generation. Pulp and paper industry is one of the most polluter industries in the environment (Baruah et al. 1996; Murugesan et al. 2000; Singh et al. 2016), and effluents released from the industry enter into the environment without proper treatment (Srivastava et al. 1994). The Indian pulp and paper industry is highly water intensive, consuming 100–250 m3 freshwater/ton paper and generating a corresponding 75–225 m3 wastewater/ton paper (Tewari et al. 2009). As per the Ministry of Environment and Forest (MoEF), Government of India, the pulp and paper industry is categorized in the “Red Category” list of 17 industries having a high pollution load and owing its toxic effect on flora and fauna. Therefore, it is obligatory for paper industry to conform the prescribed pollutant discharge limits set by Central Pollution Control Board (CPCB) (Tewari et al. 2009). The wastewater generated from paper industry during pulping, bleaching, and washing processes is often characterized by their high color, lignin, BOD, COD, TDS, SS, and potentially toxic chlorophenols, organic acids, and phosphorus and sulfur compounds along with metals (Pokhrel and Viraraghavan 2004; Haq et al. 2016a, b; Kumar et al. 2017a, b, c). The adverse effect of pulp and paper mill effluent pollutants is well studied. The higher molecular weight compounds may be nontoxic to living organism because they cannot penetrate easily inside the cellular membrane, but after degradation of these compounds, they may be converted into lower molecular weight compounds, which could be active and toxic to living organisms (Pokhrel and Viraraghavan 2004). Partially treated effluent also induces carcinogenic, mutagenic, clastogenic, and endocrine-disrupting effects in exposed organisms (Savant et al. 2006; Ali and Sreekrishnan 2001). Various studies have reported detrimental effects of pulp and paper mill effluent on animals living in water bodies receiving the effluent. The effects are in the form of respiratory stress, liver damage, and genotoxicity (Vass et al. 1996; Johnsen et al. 1998; Schnell et al. 2000). The wastewater generated from pulp and paper industry causes various health implications such as headache, vomiting, nausea, diarrhea, and eye irritation in industries workers and children (Mandal and Bandana 1996). In addition, this wastewater also affects invertebrate communities, zooplanktons,

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zoobenthos, phytoplankton, and benthic algal ecosystem (Johnsen et al. 1998). The environmental impact of pulp and paper mills is of particular concern, in addition of pollution; there is a huge water scarceness and poor quality of water in many parts of India. Thus, in order to improve the declining quality of water and availability of freshwater for consumption and to minimize high level of environmental pollution from improperly treated effluent, there is an urgent need for efficient water management in pulp and paper mills. The paper mill effluents are often treated by industries using secondary treatment processes such as activated sludge process (ASP) or aerated lagoon. Recently, several physical, chemical, and biological methods as well as combination of different methods in series that have been developed for the treatment of pulp and paper mill wastewater showed successful application for the treatment methods. Over the years, the physical (adsorption, microfiltration, photoionization, etc.) and chemical (sedimentation, coagulation, oxidation, ozonation, etc.) methods have been shown to improve the quality of treated effluent (Kamali and Khodaparast 2015). Further, improvements have been observed by including a combination of the biological process followed by physicochemical methods for the removal of organic pollutants from paper mill effluents (Gonzalez et al. 2010; Pavon-Silva et al. 2009). However, these systems require upgradation of technology in industry for the improvement of existing wastewater treatment facilities. Hence, any approach which proposes to improve the existing treatment facilities with a better efficacy will be more preferable and acceptable to industry. Further, majority of the work on the treatment of paper mill effluent by whole culture, crude enzyme preparation, and purified ligninolytic enzymes had been carried out using fungi (Raghukumar et al. 2008). However, fungi are not stable under actual treatment due to extreme environmental (Hataka 1994) and physiological stress conditions (Crawford and Muralidhara 2004). For this reason, bacterial ligninolytic systems have been employed for the bioremediation of effluent. The recent study suggested that the use of bacterial system for the removal of toxic pollutant from paper mill effluent has been recognized as a cost-effective and eco-friendly method (Singh et al. 2011; Raj et al. 2007; Thakur 2004; Gupta et al. 2001).

2 Pulp and Paper Industry Globally paper industry is one of the top most industries in the context on the socioeconomic development. Including in India, this industry plays a key role in the overall economic growth. Indian paper industry is one of the world’s fastest growing industries. Paper industry utilizes plant material, various chemicals, and large amount of freshwater during production of paper. In India about 759 paper mills are functional, out of which there are 99 in Uttar Pradesh, with an installed capacity of 12.7 million tons (Mt) producing 10.11 Mt. of paper and paperboards which is 2.52% of the total world production of 402 Mt. per annum (Mtpa). Apart from 10.11 Mt. productions, 1.04 Mt. is imported annually. Therefore, the present

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consumption of paper and paperboard including news print is at 11.15 Mt. There are nearly about 26 large integrated paper industries using wood and bamboo that contribute 31% of the production which is 3.19 Mt. and 150 mills using agroresidues like bagasse, wheat and rice straw, etc., producing 2.2 Mt. which is 22% of the total production (Kulkarni 2013). Due to limited forest resources, other raw materials including agro-residues like rice straw, wheat straw, sarkanda, bagasse, and jute rags as well as waste paper were identified and are now extensively used in the paper mill for the production of paper products. The major raw material used in Indian pulp and paper industry are hardwood and bamboo, agro-residues, and recycled fiber/waste paper.

2.1

Hardwood and Bamboo

In present during the papermaking process, consumption of raw material (wood) is 9 Mtpa. Major portion (75%) of the wood demand is supplied from farm/social forestry sources. The consumption of wood materials by the year 2025 has been projected to increase up to 12 Mtpa. Woody plant materials are mainly made up of approximately 50–55% cellulose, 25–30% lignin, and 20–25% hemicellulose. Eucalyptus, casuarina, subabul, poplar, and bamboo are the plant material used for pulp production.

2.2

Agro-Residues

The use of nonwoody plant is currently widespread in developing countries due to the lack of forest resources and advanced processing technologies. China is the largest user of nonwoody plants as raw materials for paper production followed by India and Thailand. Nonwoody plant materials included are crops (hemp, kenaf, flax, jute), agricultural residues (wheat, corn or rice straw, bagasse, sisal), and wild plants (grasses, bamboo, and seaweed). Bagasse and wheat straw are the two major agroresidues used by the paper industry in India. The agro-residues contain the following contents such as lower lignin content, higher silica, and ash. In low-grade pulp, mainly bagasse and straw fiber are used, whereas kenaf, hemp, flax, and cotton are used in higher-grade pulp.

2.3

Recycled Fiber/Waste Paper

The recycled fiber/waste paper is best suited for end products, such as newsprint, duplex board, kraft paper, etc., and it comprises both pre-consumer and postconsumer. The pre-consumer wastes are the shavings and trimmings from paper

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machine such as printers, rejects, etc. The post-consumer waste is usually old wastepaper collected from consumers. The requirements of raw material are sourced both indigenously and through imports.

2.4

Chemicals

Paper manufacturing from wood is a chemical-intensive process. Large amount of various chemicals used during papermaking process, viz., optical brightening agent, pigments, retention agent, caustic soda, sizing agent, wet strength additive, dry strength additive, mineral fillers, coating binders, pulping chemicals, and bleaching chemicals, are summarized in Table 14.1. In kraft and sulfite pulping, the cooked wood chips are dipped in a chemical solution to dissolve the lignin that binds the fibers together. Caustic soda (NaOH) solution was used for processing of recovered fibers. Pulp and paper mills also use combinations of chlorine and oxygen-based chemicals to bleach or brighten the pulp. At the time of papermaking process, various coatings, fillers, and also many other additives are added to the pulp to assist the manufacturing and encounter the functional requirements of various types of paper.

2.5

Water

In papermaking process, water plays an important role because it carries fibers through each step of paper manufacturing process and chemical treatment and separates spent pulping chemicals and the complex combination of organic matter from the pulp. Water is the basic process medium of pulp and paper manufacturing; it carries the fibers through each manufacturing step and chemical treatment and separates spent pulping chemicals and the complex mixture of organic residues from the pulp. Huge amount of water is used during the paper manufacturing process.

3 Steps in Pulp and Paper Production The pulp and paper production process involved mainly four steps: wood handling and debarking, pulping, bleaching, and papermaking. The flow diagram for paper manufacturing process employing wood and/or agro-residues as raw material is shown in Fig. 14.1.

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Table 14.1 Major chemicals used in paper manufacturing process Common name AKD Alabaster or annaline Alginic acid

Chemical name Alkyl ketene dimer Calcium sulfate (anhydrate)

Chemical formula CaSO4

Notes Sizing Paper loading material

Sodium alginate (“Naalginate”) Chlorine dioxide

Na(C6H8O6N) ClO2

Chlorine Ethylenediaminetetraacetic acid

Cl2 C10H16N2O8

Formamidine sulfuric acid Natural polymer Hydrogen peroxide

CH4N2SO2 H2O2

Pulp bleaching Used for chelation (removal of transition metals from pulp) Used in deinking Post-deinking bleaching Dry strength additive In pulp bleaching

Hypochlorous acid

HOCl

In pulp bleaching

Calcium oxide

CaO

Limestone

Calcium carbonate

CaCO3

Magnesium bisulfite Oxygen Ozone Sodium bisulfite Sodium dithionite Sodium hypochlorite Sodium peroxide Sodium sulfide Sodium thiosulfate Sulfur

Magnesium bisulfate

Mg(HSO3)2

Alkaline pulping process, chemical recovery, bleaching To make precipitated CaCO3, is used as filler and in coating Used in sulfite pulping

Oxygen Ozone Sodium bisulfite

O2 O3 NaHSO3

In pulp bleaching In pulp bleaching Used in sulfite pulping

Sodium hydrosulfite

Na2S2O4

Bleaching

Sodium hypochlorite

Na2S2O4

Bleaching

Sodium peroxide

Na2O2

Bleaching

Sodium sulfide

Na2S

Sodium thiosulfate

Na2S2O3

Active chemical in kraft/sulfate cooking liquor Bleaching

Sulfur

S

Chlorine dioxide Chlorine EDTA Enzyme FSA Guar gum Hydrogen peroxide Hypochlorous acid Lime

Wikipedia (https://en.wikipedia.org/wiki/Paper-chemicals)

Coating and surface treatment Pulp bleaching

To make HSO3 for bisulfite pulping

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Fig. 14.1 Flow diagram of paper manufacturing process in paper industry

3.1

Debarking

The wood materials for the paper production started with conversion of wood logs into chips suitable for pulping in a series of steps which may include debarking, sawing, chipping, and screening. Logs are debarked hydraulically with highpressure water jets, or mechanically by rubbing logs against each other or with metal chippers. The plant bark is a useless material for paper industry as it contains little fiber and high content of extractives; it is dark in color and often carries large quantities of grit. The chips for pulping are then passed over a series of screens to separate chips on the basis of length or thickness. The suitable chip sizes are available to the pulping process (Ljungberg and Brannvall 2011).

3.2

Pulping

Pulping of wood chips is an important step in paper manufacturing. Pulping is performed to obtain cellulose fibers from raw plant materials including hardwood, agro-residues, and recycled fiber/waste paper. Pulping can be performed by chemical, mechanical, and chemo-mechanical processes. Chemical process such as “kraft

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process” is one of the most dominating pulping technologies worldwide. Kraft process involves digestion of wood chips in an alkaline solution of sodium hydroxide (NaOH) and sodium sulfide (Na2S) for removal of hemicelluloses and lignin components. Sulfite process is another common popular pulping process which is based on acidic or neutral cooking with salts of sulfites (SO32
) or bisulfates (HSO4
) and bases like calcium (Ca2+), magnesium (Mg2+), ammonium (NH4+), sodium (Na+), etc. (Hultman 1997). Heat and pressure is used for the pulping process. The reaction continues to a certain predefined degree of delignification, and the products receive a good strength. About 40–50% of fiber yields from chemical pulping, and the remaining is considered as by-products or burned in recovery boiler (Hultman 1997).

3.3

Bleaching

The residual lignin (5%), which remained in pulp after pulping process, is removed by multistage bleaching process to give the paper a specific brightness (Ljungberg and Brannvall 2011). Chlorine (Cl2), chlorine dioxide (ClO2), hydrogen peroxide (H2O2), caustic, oxygen, ozone, hypochlorite, and sodium bisulfite are the major used bleaching agents.

3.4

Chlorine Bleaching

In papermaking process bleaching is an important process in which chlorine bleaching with elemental chlorine is the most common technology, which generates adsorbable organic halides (AOX). In this process residual lignin is removed between the range of 5% and 10%. In this process, for whitening the pulp, many steps are performed with the help of chlorine dioxide or hypochlorite.

3.5

Elemental Chlorine-Free (ECF) Bleaching

In large-scale paper mills, the brightness of pulp is achieved by ECF bleaching where it uses oxygen delignification (ODL), followed by ClO2 and other chemical agents. The sequence of the chemicals includes chlorine dioxide, caustic soda, oxygen, and hydrogen peroxide. The ECF bleaching is based on the utilization of ClO2, in which only small portions of AOX are formed.

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Total Chlorine-Free (TCF) Bleaching

TCF bleaching includes combination of ODL with ozone/peroxide which leads to no formation of AOXs. There are some agents, such as pH, temperature, and reaction time, which decide the bleaching stage. Before the progression to the next stage, the pulp is washed with caustic to remove bleaching chemicals and dissolved lignin. At the end of bleaching process, the pulp is pumped through a series of screens and cleaners to remove contaminants. It is then concentrated and conveyed to storage. The bleaching process can be enhanced by the use of enzymes, and a “chelating” agent (ethylenediaminetetraacetic acid, EDTA) is added to bind the metal ions contained in the pulp and prevent them from decomposing the hydrogen peroxide.

3.7

Hydrogen Peroxide Brightening

The bleaching of pulp with hydrogen peroxide is very effective in removing of high content of lignin. Hydrogen peroxide alters the chemical structure of lignin by oxidizing and remains with the pulp. Though hydrogen peroxide is environmentally benign, it is expensive.

3.8

Paper Production

After the bleaching, the pulp is then converted to paper products such as writing paper, books, and grocery bags. Sulfite pulp, which is primarily cellulose, is converted into specialty paper, rayon, photographic film, plastic, adhesive, and even ice cream and cake mixes. The fibers in pulp from recycled paper are usually shorter, less flexible, and less water permeable and can therefore not be used for high-quality paper products. Recycled pulp is therefore mainly used for the production of soft paper products like tissue paper, toilet paper, paper towel, and napkins. To produce market pulp, the pulp consistency is adjusted to 4–10% before it is ready for the pulp machine to make pulp sheet using plastic mesh or travelling metal screen. The pulp sheet is passed through a series of rotating rolls that squeeze out water and air until the fiber consistency is 40–45%. The sheet is then floated through a multistory sequence of hot air dryers until the consistency is 90–95%. Further, the obtained pulp sheet is cut into small portion and stacked into bales. Finally, the pulp bales are compressed, wrapped, and packed into bunches for storage and delivery (Ljungberg and Brannvall 2011).

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4 Pollutants Discharged in Wastewater During Paper Production Large amount of various chemicals are used in paper production process including sodium hydroxide, sodium carbonates, sodium sulfide, bisulfites, elemental chlorine or chlorine dioxide, calcium oxide, hydrochloric acid, etc. resulting in the generation of huge quantities of wastewater loaded with organic and inorganic salts and toxic pollutants, which are let out after incomplete treatment. The studies of Fazeli et al. (1998) showed an increased level of heavy metals accumulated in plant tissues of paddy crop irrigated with paper mill effluent. Copper is also one of the main contributors among other heavy metals of acute toxicity in the kraft mill effluent (Reyes et al. 2009). There are also various compounds such as chlorinated organic compounds, phenolics, and lignin which are often not removed adequately by conventional approaches due to their highly toxic nature and low biodegradability and pose a threat to aquatic receiving environments (Ali and Sreekrishnan 2001). During the pulping, bleaching, and chemical recovery process, lignin is produced in wastewater. During paper manufacturing process, it undergoes a variety of reactions including aryl–alkyl cleavages, strong modification of side chains, and various ill-defined condensation reactions fragmenting the polymer into small water/alkalisoluble compounds (Chakar and Ragauskas 2004). There are some other phenolic compounds such as dichlorophenol, trichlorophenol, dichloroguicol, tetrachloroguicol, and pentachlorophenol which are formed during the papermaking process and observed as more toxic compounds (Ragunathan and Swaminathan 2004). The study of Savant et al. (2006) reported that the paper mill effluent is contaminated by about 500 different chlorinated organic compounds. In the aquatic system, the dark brown color of paper mill effluent blocks the photosynthesis and decreases the dissolved oxygen (DO) level, which leads to adverse effects on flora and fauna and poses toxicity to aquatic ecosystem (Latorre et al. 2007; Sponza 2003). The accumulation of toxic pollutants and metals in effluent contaminated soil causes adverse effect on growing plant system (Kumari et al. 2014; Raj et al. 2014a, b; Haq et al. 2016b). The characteristics of some pulp and paper mill effluent are given in Table 14.2. Major pollutants found in wastewater or effluents are briefly discussed below:

4.1

Organic Pollutants and Suspended Solids

In pulp and paper wastewater, there are many components such as cellulose, hemicellulose, fugitive fibers, starch, and organic acids which are mainly responsible for organic pollution. This results in a COD discharge of pulp in the range of 25–125 kg/ton1‒. In the aquatic ecosystem, high concentration of BOD and COD value is responsible for depletion of dissolve oxygen. There are several toxic components like fatty acids, resin, and toxic metalloids which are present in pulp mill wastewater and absorbed by the organic solids. Further, it causes long-term

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Table 14.2 Characteristics of some large- and small-scale paper mill effluent and minimum national standards for pulp and paper mill wastewater discharge

Parameter pH Color Lignin BOD COD T.S. T.D.S. T.S.S. Sulfate Total phenol Total nitrogen Phosphate Nitrate Chloride Potassium Sodium CO2 Heavy metals Cd Cr Cu Fe Mn Ni Pb Zn

Large-scale paper mill (Chandra and Singh 2012) 6.70 Dark brownish 149.80 351.00 1057.00 1423.00 1115.00 273.00 234.00 72.96

Small-scale paper mill (Yadav and Chandra 2015) 8.50 Dark brownish black 413.00 6033.00 15766.00 1570.00 1274.00 296.00 405.49 44.64

Permissible limit (CPCB 2000) 5.5–9.0 – – 30 250 – 2100 50 – 1

Permissible limit (ISI 1974) 5.5–9.0 – – 100 350 – 1000 – – 1

39.15

152.93

–

–

683.69 32.38 362.00 24.45 285.00 56.20

500.00 35.30 351.00 26.57 28.98 257.00

– – 1000 – – –

– – 600 – – –

0.135 0.058 0.216 0.182 0.04 0.122 ND 0.062

ND 0.056 0.074 2.004 0.166 0.016 ND 0.066

– – – – – – – –

– – 3 – – 2 0.1 5

CPCB Central Pollution Control Board; ISI Indian Statistical Institute

effects over an extensive area and it consequences of bioaccumulation and transportation through the food chain.

4.2

Organochlorine Compounds

There are a large number of organochlorine compounds, such as chlorinated derivatives of phenols, acids, and dibenzo-p-dioxins/furans, and other neutral compounds

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that have emerged during paper production processes, which are a major concern for the environment. Bleached effluents may contain chloroform and carbon tetrachloride, which are classified as carcinogens. The hypochlorite stage is the major producer of chloroform. In pulp and paper mill wastewater, many compounds such as phenols, chlorinated benzenes, epoxy stearic acid, and dichloromethane are found, which act as carcinogens.

4.3

Chlorophenolics

The chlorophenolics are toxic, persistent, and bioaccumulative which emerged from pulp bleaching process. These compounds are formed due to the use of large amount of chlorine compounds. Chlorophenolics, such as chloroguaiacols, chlorophenols, and chlorocatechols, and some chloroaliphatics are extremely toxic and recalcitrant compounds. These compounds accumulate in sediments or at different levels of the trophic chain causing detrimental effects toward all life forms.

4.4

Dioxins (Polychlorinated Dibenzodioxins) and Furans (Polychlorinated Dibenzofurans)

It is well known that dioxins and furans are persistent in nature and reported as potent carcinogenic agent. Chemical structures of furans are similar with dioxins but of less magnitude. These two toxic compounds are found in sludge after the wastewater treatment and cause serious concern. Chemicals, such as chloroform, chloroacetones, aldehydes, and acetic acid, are formed during bleaching process, but in lower concentrations than chlorophenolics. Generally these compounds are nonpersistent and non-bioaccumulative, but some of these are moderately toxic, mutagenic, and suspected carcinogens.

5 Environmental Contamination and Toxicity Toxic effect of paper mill effluent to fish has been well documented since long back (Johnston et al. 1997; Mishra et al. 2011). Water bodies receiving paper mill effluent were observed with reduced oxygen levels, eutrophication, and deposition of sludges and accompanying microbial growth. Highly toxic chlorinated compounds identified in paper mill effluent during 1980 included phenolics, organic acids (fatty acids and resin acids), and highly toxic dioxins and furans (Ali and Sreekrishnan 2001). These compounds are bioaccumulative and mutagenic in exposed flora and fauna in respective environment (Ali and Sreekrishnan 2001; Hewitt et al. 2006; Dey et al.

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2013). Effluent emerged from elemental chlorine bleaching process showed reduction in effluent toxicity and also symptomatic changes but not absence of chronic impacts (Munkittrick and Sandstrom 1997). However, recent studies showed that effluent emerged from elemental chlorine-free (ECF) or total chlorine-free (TCF) processing still exhibited acute lethality in their discharges (Sepulveda et al. 2003). These suggest that there are many other compounds present in paper mill effluent which are potential to cause toxicity in aquatic organisms. Paper industry having secondary effluent treatment process had a huge impact on effluent quality. Disappearance of acute effect of paper mill effluent due to process changes and improvement has been observed (Munkittrick and Sandstrom 1997).

6 Wastewater Treatment Process The Indian pulp and paper industries use conventional primary and secondary treatment process for the treatment of wastewaters.

6.1

Primary Treatment

Generally, in the primary treatment process, reduction of suspended solids (fibers, bark particle, filter, coating materials, etc.) has been achieved by the use of screens and settling tanks. Primary treatment devices are used to remove the large suspended or floating solids, heavy inorganic solids such as sand and gravel, as well as metal or glass. These solids consist of pieces of wood, cloth, paper, plastics, garbage, etc. Primary treatment of effluents consists of two main processes, preliminary treatment and primary sedimentation. The devices of this treatment system reduce the velocity and dispense the flow of wastewater. In primary treatment, the velocity of flow is reduced to 1–2 feet/min‒1 to maintain a quiescent condition so that the material denser than water will settle out and material less dense than water will float to the surface. The effluent from primary treatment therefore contains mainly colloidal and dissolved organic and inorganic solids. This additional removal of organics can be accomplished by secondary treatment.

6.2

Secondary Treatment

In the secondary treatment process, various types of bioreactors are used for the mitigation or removal of wastewater contaminants. The most common bioreactors are activated sludge process (ASP), aerated lagoon (AL), membrane bioreactors (MBBR), and biofilm process for degradation of organic pollutants by suspended growth biomass. At present most of the paper industry effluent treatment process is

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Table 14.3 Characteristics of some large-scale pulp and paper mill effluents after treatment

Parameter pH TDS (mg L
1) Color (CU) BOD (mg L
1) COD (mg L
1) Lignin (mg L
1) Phenolics (mg L
1)

Star Paper Mill, Saharanpur, UP (Raj et al. 2014a) 8.2 850

Century Paper Mill, Lalkuan, Nainital, Uttarakhand (Yadav and Chandra 2015) 8.5 977

Nagaon Paper Mill, Assam (Das et al. 2013) 7.6 1460

CPCB standards (2001) 5.5–9.0 2100

2242

2538

NS

NS

385

7250

270

30

792

16,550

438

250

436

800

NS

NS

42

386

71.6

1.0

NS not specified

based on activated sludge process. In activated sludge process (ASP), aeration, and sedimentation, two steps are used for the treatment of wastewater. During the first step, wastewater is treated with a high concentration of microorganism and a powerful aeration, and the retention time can vary between a couple of hours and up to a day, while in the second step, water and sludge are separated in a sedimentation basin and parts of the sludge are pumped back to the aeration basin. For the significant reduction in organic compounds/materials, recirculation of sludge facilitates high concentration of microorganism. In comparison to aerated lagoons, activated sludge system is more susceptible and cannot stand fast load changes. However, the degree of efficiency can be controlled, and very high BOD reductions are seen (Thompson et al. 2001; Persson 2011). Although ASP is the most efficient effluent treatment method, it is insufficient to degrade various components of pulp and paper mill effluents. The characteristics of the effluents released after these treatments by some large-scale pulp and paper mills do not conform the discharge standards set by the CPCB in case of BOD, COD, and phenolic. The characteristic of treated effluent of various industries is given in Table 14.3. Therefore, more advanced alternative biological wastewater treatment strategy is required to meet discharge limits set by regulatory agency.

7 Existing Treatment Technologies Efforts are being made toward minimization of effluent toxicity by various internal process changes and management measures through the available technology (Wu et al. 2005; Yang et al. 2008; Wang et al. 2011). Many physical, chemical, and biological methods are available for the paper industry effluent treatment/

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detoxification. However, complete removal of pollutants has not been achieved (Orrego et al. 2010).

7.1

Physicochemical Methods

Physicochemical processes based on sedimentation and floatation, coagulation and precipitation, filtration, reverse osmosis, adsorption, ozonation, and other advanced oxidation have been well reported for the effective removal of pollutants as well as variety of suspended and floating matters from the paper industry wastewaters. Application of sedimentation and floatation processes, used for separation of solids from effluent, has been evaluated in paper industry effluent as an effective decolorization process (Pokhrel and Viraraghavan 2004; Thompson et al. 2001). However, practical viability of such systems at large scale is not feasible due to its high cost and generation of large amount of sludge which pose disposal problem. Coagulation and precipitation methods using metal salts to generate larger flocs from small particles have been also used for the treatment of paper industry effluent. Aluminum chloride as coagulant and modified natural polymer (starch-g-PAM-gPDMC) as a flocculant for treatment of wastewaters from primary sedimentation tank have been studied by Wang et al. (2011), and result suggested that at the optimal condition (coagulant dosage of 871 mg L
1, flocculants dosage of 22.3 mg L
1, and pH 8.35), turbidity and lignin removals and water recovery efficiency were 95.7%, 83.4%, and 72.7%, respectively. Membrane technologies have also been applied for pulp and paper mill wastewater treatment. Membrane treatment technology in pulp and paper industry is performed to optimize loop closure and therefore helps to reduce freshwater intake as well as wastewater treatment. The other purposes of membrane treatment processes are to improved product quality because of minimizing pollution of wastewaters reuse of treated effluent and recovery of valuable substances and reduction of environmental pollution because of improved effluent quality. Ultrafiltration (UF) and reverse osmosis are some examples of membrane filtration technology for color removal of effluent. Usually, they have been applied for the removal of high molecular weight dissolved organic components from paper mills effluent. The filtration techniques are now being used for the removal of color and for further reuse of process wastewaters from paper mill industry (Joensson et al. 1996; Katkar and Sasidharan 2000; Laitinen et al. 2001). Membrane techniques usually represent a large capital investment. This process requires pretreatment in order to prevent problems of fouling. The technique of adsorption has been found to be an efficient and cost-effective method to remove pigments, dyes, and other colorants of industrial effluents and to control the BOD and COD. Activated carbons, inorganic oxides, and natural adsorbents (such as clays and clay minerals, cellulose materials, chitin, and chitosan) have been extensively used as adsorbents by many workers to treat wastewaters (Al-Asheh et al. 2003; Naseem and Ve Tahir 2001). It is an advantageous technique

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which use small amount of adsorbent with large surface area (Marquez and Costa 1996). Various types of adsorbents derive from different sources can be used to adsorb specific adsorbate or target compounds. The best adsorbent for the treatment of specific target compound can be selected on the basis of higher adsorption capacity of the adsorbents for target solute to be removed. Ozone gas is a strong oxidizing agent, and it reacts with inorganic and organic compounds directly or indirectly by the formation of hydroxyl radicals. It mainly oxidizes electron-rich molecules containing carbon–carbon double bonds and aromatic alcohols. The uses of ozone gas for the effluent treatment comprise splits of long-chain compounds and making them biodegradable. The combination of ozonation and aerobic bio-treatment is demonstrated to be an effective method for destroying lipophilic extractives and hence increases the biodegradability of pulp and pulp mill wastewater before returning them to the bio-treatment unit (Kamenev et al. 2008). Advanced oxidation processes (AOPs) are among the promising technological approaches that have received interest for the treatment of pulp and paper bleach effluents. AOPs are employed to oxidize complex organic compounds of wastewater that are difficult to degrade biologically into simpler end products. The mechanism of AOPs is based on the generation of very reactive nonselective transient oxidizing species such as the hydroxyl radicals (OH•), which were identified as the dominant oxidizing species (Al-Rasheed 2005). The most common AOPs are H2O2, UV/O3, O3, Fenton’s reagent UV/H2O2, and photochemical processes (Metcalf and Eddy 2003).

8 Biological Approaches (Bioremediation) for the Treatment of Pulp and Paper Mill Wastewater Biological treatment (or bioremediation) is an eco-friendly technique that uses microbes/plants or their enzymes for the treatment and detoxification of industrial effluents containing organic and inorganic pollutants (Chandra et al. 2015; Bharagava et al. 2017a, b; Saxena and Bharagava 2017; Haq et al. 2017, 2018; Haq and Raj 2018; Bharagava et al. 2018; Saxena et al. 2019). Biological treatment method (bioremediation) of paper mill effluents employed several microorganisms such as bacteria, fungi, algae, and many enzymes, as a single-step treatment or in combination (biphasic) with other physical and/or chemical methods (Singhal and Thakur 2009). In comparison to the physicochemical process, biological methods for the treatment of wastewater are considered to be of cost-effective, environment friendly, and appropriate for BOD and COD reduction from the wastewater.

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Treatment with Fungi

Paper industry wastewater is mainly adequately treated by fungi (Yang et al. 2011) because they produce extracellular ligninolytic enzymes (laccase, MnP, and LiP) and are more effective in comparison to bacteria (Singhal and Thakur 2009). Among all the fungal species, white rot fungi (Phanerochaete chrysosporium and Trametes pubescens) have reported for the degradation of lignin or phenolic compounds by ligninolytic enzymes (Zhang et al. 2012; Gonzalez et al. 2010; Chandra and Singh 2012; Freitas et al. 2009). In addition, kraft pulp and paper mill wastewater bioremediation was achieved by immobilized fungal consortium, which consists of two basidiomycetes and deuteromycetous fungi such as Merulius aureus with an unidentified genus and Fusarium sambucinum, respectively. In this study, after 4 days of incubation, reduction of color, lignin, and COD was 78.6%, 79.0%, and 89.4%, respectively (Malaviya and Rathore 2007). Freitas et al. (2009) have reported that Pleurotus sajor-caju and Rhizopus oryzae are used for the reduction of color and COD of wastewater discharged from the secondary treatment of a bleached kraft at their relative absorbance. In this study, it was observed that after 10 days of incubation period, reduction of color was 25–46% at 250 nm and 72–74% at 465 nm and COD was 74–81%. In laboratory-scale study at different optimized conditions (3% inoculum, pH 6, shaking at 160 rpm, 30  C, and 60–72 h), pulping effluent was treated with the help of Aspergillus niger. Further, it was observed that reduction of color, COD, and turbidity was 43%, 60%, and 77%, respectively, while methyl tertiary butylether (MTBE) was about 97% (Liu et al. 2011). Various researchers reviewed that treatment of pulp and paper mill wastewater by using various fungal species (Garg and Modi 1999; Garg and Tripathi 2011; Tripathi et al. 2007). However, the main constraint in using a fungal degrading system is the requirement to maintain growth and/or enzymes (ligninases) activity at the prevailing low pH (4–5). However, at low pH, the solubility of high molecular weight fragments that are derived from lignin is reduced. Furthermore, the natural pH of pulp and paper mill effluent is alkaline (pH ¼ 8–9). Therefore, any requirement to reduce the pH to the acidic range prior to fungal augmentation would be uneconomical (Raj et al. 2007). There are several facts mentioned in this section, and bacterial treatment systems that have an optimum pH (7–9) may play a key role in biodegradation of pulp and paper mill wastewater, without any previous requirement of pH modification. Due to this reason, studies on the bacterial decolorization/ degradation were more suitable for pulp and paper wastewater treatment, and on the basis of available literature, production of bacterial ligninolytic enzymes has been more demanding in recent years (Rahman et al. 2013; Raj et al. 2014a).

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Treatment with Bacteria

Currently, research works suggest that the use of ligninolytic bacterial system is a promising approach for the degradation/decolorization of paper industry wastewater. The large number of bacterial spp. has been employed for the degradation of paper mill effluent, and some of them are now being used commercially for this purpose. Flask-scale studies with the bacterial spp. (Bacillus subtilis, Micrococcus luteus, Pseudomonas aeruginosa, Bacillus cereus, Pseudomonas putida, Acinetobacter calcoaceticus, Ancylobacter, Methylobacterium, Citrobater, Enterobacter) have been well reported to reduce color (50–97%), BOD (80–96%), COD (80–97%), lignin (50–97%), and AOX in paper mill effluent (Tyagi et al. 2014; Ramsay and Nguyen 2002; Tiku et al. 2010; Raj et al. 2007; Abd-ElRahim and Zaki 2005; Keharia and Madamwar 2003; Tiku et al. 2010). Numerous studies have been reported that bacteria can degrade monomeric lignin substructure models, while only few strains are reported to degrade lignin derivatives obtained from different pulping processes (Hao et al. 2000; Chandra et al. 2011; Chandra and Bharagava 2013). Lignin degradation by bacterial genera, such as Bacillus, Alcaligenes, Arthrobacter, Nocardia, Pseudomonas, and Streptomyces, has been well reported (Chandra et al. 2007). Aerobically or anaerobically, lignin is apparently not biodegraded by neither rapid nor extensive bacterial system. Lignin degradation has not been most extensively studied in Streptomyces viridosporous and Streptomyces, but a mixed population of bacteria and protozoa derived from lake-bottom sediment near the effluent kraft paper mill was shown to degrade lignin (Hossain and Ismail 2015). In a study the use of commercial kraft lignin as sole source of carbon by Streptomycetes badius and S. viridosporous was characterized by the use of FTIR and SEC (Abd-ElRahim and Zaki 2005; Chandra et al. 2011). Resin acids are also found in paper effluent and toxic to aquatic animals, which are reported to be degraded by many bacterial spp. (Bacillus sp., E. coli, Flavobacterium sp., Pseudomonas, Alcaligenes eutrophus, Arthrobacter, Sphingomonas, Zooglea, Comamonas, Mortierella isabella, Chaetomium cochliolidae, Corticum sasakii, and Fomes annosus) (Tiku et al. 2010; Khansorthong and Hunsom 2009; Raj et al. 2014a; Ramsay and Nguyen 2002). The studies of Tiku et al. (2010) and Aftab et al. (2011) using Burkholderia cepacia were interesting as they showed that organism was able to hydrolyze triglycerides to free fatty acids and liberated unsaturated fatty acids were then degraded to some extent and saturated fatty acids were not degraded. Nearly 30% of the stearyl esters, 25% of the dehydroabietic, and 45% of the abietic and isopimaric resin acids were degraded during 11 days by increasing the concentration of free sterols. The degree of unsaturation seemed to be of greater importance for the degradation of fatty acids. The induction of ligninolytic enzyme (laccase, lignin peroxidases, and manganese peroxidases) during bioremediation of rayon grade pulp and paper mill effluent by a bacterial consortium of Pseudochrobactrum glaciale, Providencia rettgeri, and Pantoea sp. has been proven by Chandra and Singh (2012). The consortium of these bacteria effectively reduced lignin and chlorophenol (90%) within 216 h of treatment. The recent study by Raj et al.

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(2014a) also confirmed induction of laccase enzyme during bioremediation of paper mill effluent by Paenibacillus sp. Bacterial treatment systems are particularly attractive, since in addition to color and lignin, they also reduce the BOD and COD of the effluent (Rodrigues et al. 2008).

9 Conclusion Finally, it is concluded based on available literature on each and every step associated with paper production and wastewater treatment, and we came to an end that environmental pollution by pulp and paper mill is continued due to lack of efficient wastewater treatment method. Therefore, paper industries need a cost-effective and environmental friendly treatment technique that treat pulp and paper wastewater holistically. In addition, the available literature has also shown on the biodegradation of lignin and chlorophenol by using physicochemical and biological (bacteria, fungi, and algae) process and others. But still there is a big question and enigma regarding the health hazards of wastewater discharged from various categories of pulp and paper mill on plant and animal in environment including the natural habitat. The detailed knowledge for the effect of pulp and paper mill wastewater on aquatic flora and fauna including the human is still not so known so far. Thus, this chapter covers all the pulp and paper mill-related problem and mitigation/removal approaches for the safe and green environment. Acknowledgments Authors are thankful to the Director of CSIR-IITR, Lucknow (India), for his encouragement and support. Financial support from the Department of Biotechnology (DBT), Government of India, New Delhi (Grant No.BT/PR20460/BCE/8/1386/2016), is highly acknowledged.

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Chapter 15

Cadmium as an Environmental Pollutant: Ecotoxicological Effects, Health Hazards, and Bioremediation Approaches for Its Detoxification from Contaminated Sites Sushila Saini and Geeta Dhania

Abstract Cadmium (Cd) is a toxic heavy metal that enters the environment through various natural and anthropogenic sources and is a potential threat to most organisms including human beings. Cadmium is nondegradable in nature and hence once released to the environment stays in circulation. With progressive industrialization, the amount of this polluting toxic metal is increasing at an alarming rate. Humans get exposed to cadmium by ingestion (drinking or eating) or inhalation (breathing). Ailments such as bone disease, renal damage, and several forms of cancer are attributed to overexposure to cadmium. Bioremediation is an innovative and promising technology for the removal of heavy metals in polluted water and lands and is very attractive in comparison with physicochemical methods because of its lower cost and higher efficiency at low metal concentrations. In microbial remediation, microorganisms can be used at the site of contamination (in situ) or off the contamination site (ex situ) for remediation. Combining both microorganisms and plants is an approach to bioremediation that ensures a possible solution for heavy metal pollution since it includes sustainable remediation technologies to rectify and re-establish the natural condition of soil. Keywords Cadmium · Carcinogen · Renal damage · Itai-itai · Bioremediation · Phytoremediation

S. Saini Department of Botany, Janta Vidya Mandir Ganpat Rai Rasiwasia (JVMGRR) College, Charkhi Dadri, Haryana, India G. Dhania (*) Department of Environmental Science, Maharshi Dayanand University, Rohtak, Haryana, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_15

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1 Introduction Industrial wastewaters are considered as the major sources of heavy metal pollution in the environment (Goutam et al. 2018; Bharagava et al. 2017a, b, c; Gautam et al. 2017; Saxena and Bharagava 2015). Contamination of heavy metal in the environment has become a serious global problem due to their nondegradation and persistence in the ecosystem leading to accumulation in different parts of the food chain (Saxena et al. 2019; Saxena and Bharagava 2017; Saxena et al. 2016; Chandra et al. 2015; Igwe et al. 2005). Cadmium (group IIB of the periodic table of elements) is considered one of the most toxic elements in the environment, posing severe risks to human health and a long elimination half-life. Weathering of cadmium-rich rocks, volcanic activities, mining, smelting, industrial processes (as an anticorrosive agent, as a stabilizer in PVC products, as a color pigment, in nickel-cadmium batteries, etc.), and overuse of phosphate fertilizers to agricultural soils contribute to aggravate the situation by increasing the levels of cadmium (Zoffoli et al. 2013). The worldwide production of cadmium in 2005 was estimated to be 20,000 metric tons. Exposure to cadmium occurs primarily through ingestion of contaminated water and food and to a significant extent through inhalation and cigarette smoking. Cadmium poses a great health risk to humans and animals even at very low concentrations in the body because the metal cannot undergo metabolic degradation to less toxic species and its excretion is poor as therapeutically effective chelating agents to enhance excretion of cadmium are lacking. The target organs for cadmium toxicity in animals include the liver, kidney, lungs, testes, prostate, heart, skeletal system, nervous system, and immune system (Waalkes 2003). The International Agency for Research on Cancer (IARC 1993) classified cadmium as a human carcinogen (group I). In plants Cd toxicity causes inhibition of photosynthesis, mineral nutrient imbalance, reduced growth, and oxidative stress (Khan et al. 2007; Feng et al. 2010). In recent years, the removal of toxic metals from soil and water poses a great challenge. Several types of chemical and mechanical treatments have been established, but these are not cost-effective and have less public acceptance. The use of microorganism such as algae, fungi, bacteria, and yeasts for the removal of heavy metals like cadmium from contaminated soil and water has a low cost, is efficient, and has greater acceptability from the public. Phytoremediation is the use of plants to clean up contamination from soils, sediments, and water. An attempt has been made in this chapter to review the challenges of cadmium toxicity in humans and plants and to discuss the most promising and alternative methods for remediation of cadmium from contaminated sites.

2 Cadmium and Its Properties Cadmium was discovered in 1817 by German chemist Friedrich Stromeyer. Cadmium is an odorless, soft, bluish-white malleable transition metal having atomic number 48 and belongs to group IIB of the periodic table. It shows resemblance with zinc in many of its physical and chemical properties (Hiatt and Huff 1975; IARC 1993).

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Cadmium is a rare element, and its abundance in the earth’s crust is about 0.1 to 0.2 parts per million. It is not found in its pure state in nature and occurs mainly as cadmium sulfide (CdS, or greenockite) in zinc deposits. Most cadmium is obtained as a by-product of zinc mining and refining (Llewellyn 1994). In nature eight isotopes of cadmium exist. The largest producers of cadmium in 1996 were Canada, Japan, Belgium, the United States, China, Kazakhstan, and Germany (NTP 2004).

2.1

Physical Properties

Atomic radius and ionic radius (Cd2+) of cadmium are 1.56 Å and 1.03 Å, respectively. Its melting point is lower than other transition metals, and when it combines with other metals in alloys, it lowers their melting point. Its melting point is 321
C (610
F) and its boiling point is 765
C (1410
F). The density of cadmium is 8.65 g at 25
C and vapor pressure is 1 mm Hg at 394
C. Cadmium has high ductility and high thermal and electrical conductivity and is an excellent corrosion resistance, making it suitable for a wide variety of industrial applications. As a bulk metal, cadmium is insoluble in water and is not flammable; however, in its powdered form, it may burn and release toxic fumes (IARC 1993).

2.2

Chemical Properties

Burning of cadmium in air leads to the formation of brown amorphous cadmium oxide (CdO). Cadmium oxide is a colorless amorphous powder, soluble in dilute acids and ammonium salts and insoluble in alkalis and water. Cadmium by reacting with hydrochloric acid, sulfuric acid, and nitric acid produces cadmium chloride (CdCl2), cadmium sulfate (CdSO4), and cadmium nitrate (Cd(NO3)2). Commercial cadmium chloride is a mixture of hydrates, soluble in water and acetone, slightly soluble in methanol and ethanol, and insoluble in diethyl ether. It occurs as small colorless-to-white rhombohedral or hexagonal crystals. Cadmium sulfate is soluble in water but insoluble in ethanol, acetone, and ammonia and occurs as colorless-to-white orthorhombic crystals. Cadmium nitrate is soluble in water, ethanol, acetone, and diethyl ether and occurs as a colorless solid (IARC 1973, 1993; HSDB 2009). The most common oxidation state of cadmium is +2, but it also exists in the +1 state. The oxidation state +1 is produced by dissolving cadmium in a mixture of cadmium chloride and aluminum chloride, forming the Cd22+ cation, which is similar to the Hg22+ cation in mercury(I) chloride (Holleman et al. 1985; Cotton 1999). Cd þ CdCl2 þ 2 AlCl3 ! Cd2 ðAlCl4 Þ2

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Applications

Cadmium is a common component of electric batteries, pigments, coatings, and electroplating (USGS 2008).

2.3.1

Batteries

The primary use of cadmium, in the form of cadmium hydroxide, is in electrodes for Ni–Cd(NiCad) batteries (Sahmoun et al. 2005). NiCad batteries are used in a large variety of appliances, including compact disc players, cellular telephones, pocket recorders, handheld power tools, cordless telephones, laptop computers, and scanner radios (Fig. 15.1).

2.3.2

Electroplating

Cadmium and cadmium alloys are used as electroplated coatings on iron, steel, aluminum, and other nonferrous metals. In electroplating a thin layer of one metal is deposited on the surface of a second metal by passing an electric current through a solution containing the coating metal. A thin layer of cadmium offers good corrosion resistance in alkaline or salt solutions, so it gives a high degree of safety or durability to electrical parts, automotive systems, military equipment, and marine/offshore installations (UNEP 2008; International Cadmium Association 2011). However discarded electroplated steel puts cadmium into the environment, so the use of cadmium in electroplating has been dropped now.

1% 4% 9% Ni-Cd Batteries Pigments Coatings and platings Alloys and PVC stabilizers 86%

Fig. 15.1 Use of cadmium

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Nuclear Fission

In a nuclear reactor during nuclear fission, neutrons are liberated that leads to chain reaction. Cadmium acts as a very effective neutron poison and is used in the form of control rods to control neutron flux in nuclear fission (Scoullos et al. 2001). 2.3.4

Cadmium Compounds

Cadmium compounds have various industrial applications. Cadmium oxide is used in Ni–Cd batteries, as a catalyst, in resistant enamels, in heat-resistant plastics, in the blue and green phosphors of color television cathode ray tubes, and in the manufacture of plastics (such as Teflon). Cadmium chloride is used in electroplating, photocopying, calico printing, dyeing, mirrors, analytical chemistry, vacuum tubes, and lubricants. Cadmium nitrate is used in photographic emulsions to color glass and porcelain, in nuclear reactors, and to produce cadmium hydroxide for use in alkaline batteries. Cadmium sulfate is used in electroplating, fluorescent screens, vacuum tubes, and analytical chemistry and as a chemical intermediate to produce pigments, stabilizers, and other cadmium compounds. Cadmium sulfide (CdS) and cadmium selenide (CdSe) are used to color paints and plastics as these cadmium compounds can withstand high temperatures and disperse well in polymers producing strong colors with high opacity and good tinting strength (Herron 2001; ATSDR 2008). 2.3.5

PVC Stabilizers

Cadmium is an important stabilizer for polyvinyl chloride (PVC) that helps retard degradation when these products are exposed to either (or both) ultraviolet light or heat. Cadmium-containing stabilizers are found in everyday PVC items as door and window frames, hoses, drain and water pipes, and electrical insulation (Herron 2001; UNEP 2008).

3 Sources of Cadmium Contamination in Environment Two major sources of cadmium emission in the environment are natural and anthropogenic activities.

3.1

Natural Sources

In nature cadmium occurs both in the earth’s crust and in ocean water. In the earth’s crust, concentration of cadmium is very meager (0.1–0.5 mg/kg), but its accumulation in sedimentary rocks, marine phosphate, and phosphorites may be much, as high as 500 mg/kg (WHO 1992). From these two reservoirs, cadmium is released by

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weathering of cadmium-containing rocks, volcanic activity, forest fires, sea spray, and mobilization of cadmium previously deposited in soils, sediments, landfills, etc. It is estimated that 15,000 metric tons of cadmium is transported to oceans through rivers due to weathering and erosion of rocks (WHO 1992; OECD 1994).

3.2

Anthropogenic Sources

The largest source of anthropogenic atmospheric cadmium emissions is ferrous and nonferrous metal production (smelting and refining of zinc ores) followed by waste incineration, production of nickel-cadmium batteries, fossil fuel combustion, iron and steel plants, electroplating wastes, fertilizer and fungicide application, and generation of dust by industrial processes such as cement manufacturing (Kazantzis 1987; Huff 1999; ATSDR 2012). From these human activities, an estimated 4000–13,000 tons of cadmium is released into the environment every year (EPA 2004). Cadmium present in soil from industrial emissions or other sources is selectively taken up by edible plants, resulting in levels much higher than those in the surrounding soils. Cadmium has also been shown to bio-concentrate in water plants and in fish. Of special concern are Mollusca and Crustacea; cadmium levels in crab may be as high as 30–50 parts per million (Schwartz and Reis 2000). Different ways by which people are exposed to cadmium are: • Non-smoking population is exposed to cadmium primarily via ingestion of food (approx. 90%) and, to a lesser extent, via inhalation of ambient air, ingestion of drinking water, contaminated soil, or dust (Fig. 15.2). Cigarettes smoking Ingestion of cadmium containing food and water

Breathing cadmium containing air

Mouthing or swallowing cadmium containing items such as inexpensive children jewelry

Cadmium exposure in human beings Consuming food and beverages served in glassware and tableware decorated with cadmium based pigments Occupational settings involving Cd production and utilization

Fig. 15.2 Different ways of cadmium exposure

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• Cigarettes are a significant source of cadmium exposure for the smoking population because tobacco leaves contain large amounts of cadmium (Morrow 2001). It has been estimated that tobacco smokers are exposed to 1.7 μg cadmium per cigarette, and about 10% is inhaled when smoked (Morrow 2001; NTP 2005). • Occupations in which the highest potential exposures occur include cadmium production and refining, Ni–Cd battery manufacture, cadmium pigment manufacture and formulation, cadmium alloy production, mechanical plating, zinc smelting, brazing with a silver–cadmium–silver alloy solder, and polyvinyl chloride compounding (IARC 1993). • Contamination of children’s jewelry with cadmium can have long-term effects by introducing a toxicant that can remain in a child’s body (Mead 2010).

4 Toxicity of Cadmium in the Environment Cadmium has no known biological function in animals and humans but mimics other divalent metals that are essential to diverse biological functions. Cadmium was found to be competitive with zinc and copper and to a lesser extent to iron (Pond et al. 1995). It can cross various biological membranes by different mechanisms (e.g., metal transporters) and once inside the cells binds to ligands with exceptional affinity (e.g., metallothioneins), thereby reducing the absorption of copper and, to a lesser extent, of zinc. Especially, the liver and kidneys contain metallothioneins, which accumulate cadmium throughout the animal life. However, cadmium is not easily cleared by the cells, with a biological half-life estimated to be 10–30 years, and the poor efficiency of cellular export systems explains the long residence time of this element in storage tissues such as the intestine, liver, and kidneys (EFSA 2009).

4.1 4.1.1

Health Impact on Human Beings Effect on the Kidney

The kidney is the critical organ of intoxication after long-term exposure to cadmium. Initial sign of renal lesion is tubular damage resulting in an increased urinary excretion of low molecular weight proteins, particularly α2-, β2-, and γ-globulins. In severe cases effects on renal function are manifested by glucosuria, aminoaciduria, hypercalciuria, hyperphosphateuria, and polyuria due to decreased concentration capacity (Friberg et al. 1974). All these conditions ultimately results in nephropathy that is the major cause of mortality in workers exposed to cadmium toxicity (Fig. 15.3). A number of studies in different areas over the years have looked at the effects of cadmium on the kidney in the environmentally exposed populations (Cai et al. 1990; Jung et al. 1993; Nordberg et al. 1997). In Japan many areas were contaminated with cadmium as a result of discharges from numerous nonferrous metal mines and

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Fig. 15.3 Health impact of cadmium toxicity in human beings

Gastrointestinal effects

Skeletal lesions

Cancer risk Reproductive health

Respiratory effects

Cadmium toxicity

Developmental effects Renal dysfunction

smelters. Rice grown in that area has high concentration of cadmium, and long-term consumption of it has been linked to tubular impairment with a loss of reabsorptive capacity for nutrients, vitamins, and minerals and nephropathy and proteinuria (Friberg et al. 1974; Jarup 2002). Besides this several investigators observed an increased prevalence of kidney stones in workers occupationally exposed to cadmium (Friberg et al. 1986; WHO 1992). As compared to control populations where 5% of cases of kidney stones were reported, prevalence rates of 18–44% have been noted in cadmium-exposed workers. Increased rate of stone formation is possibly related to cadmium-induced renal damage leading to hypercalciuria and hyperphosphateuria; but other contributing factors may include uric aciduria, reduced urinary citrate, and renal tubular acidosis (Kazantzis 1979).

4.1.2

Skeletal Lesions and “Itai-itai” Disease

Significant skeletal lesions have been observed in later stages in severe chronic cadmium poisoning (Alfvén et al. 2004). Cadmium interferes with the metabolism of calcium, magnesium, iron, zinc, and copper and leads to demineralization of the bones, intense bone-associated pain, and osteomalacia, osteoporosis, and other disorders of the bones. Cadmium also reduces the normal activation of vitamin D. Vitamin D is activated in the kidney to calcitriol (1,25-dihydroxy vitamin D), which plays an important role in the absorption of calcium from the gut and in the calcification of the bone (Aoshima and Kasuya 1991). Prolonged exposure to cadmium may give rise to itai-itai (“ouch-ouch”) disease. Itai-itai or “ouch-ouch” disease, so named from the painful screams due to severe pain in the joints and the spine, was caused by ingestion of runoff water containing cadmium released from mining companies in Toyama Prefecture, Japan (Kagawa 1994). Farmers in the region used the runoff for irrigating rice paddies and other

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crops. Cadmium concentrated in the crops, and local women began to experience pain in bones and joints, eventually becoming so excruciating that they were bedridden. Cadmium interfered with calcium metabolism, leading to reduction in calcium levels and thus reduced density and strength of bones, often causing the weakened bones to break (Watanabe et al. 2004). Symptoms and signs of “itai-itai” disease include: • Severe osteoporosis and osteomalacia with simultaneous severe renal dysfunction. • Normochromic anemia and low blood pressure sometimes also occur (Alfvén et al. 2002; Nogawa et al. 2004).

4.1.3

Cancer Risk

Cd is a multi-site carcinogen in both rodents and humans and has been classified as a Class B1 carcinogen by the International Agency for Research on Cancer and the US Environmental Protection Agency (Waalkes et al. 2000; Huff et al. 2007). Many studies have found a linkage between occupational exposure to cadmium and pulmonary cancer (Nawrot et al. 2006), as well as prostate (Zeng et al. 2004), renal (Pesch et al. 2000), liver and stomach (Waalkes and Misra 1996), and urinary bladder cancer (Kellen et al. 2007). Waalkes (2003) reported increased risks of lung cancer in workers exposed to Cd by inhalation. In another study a significant trend for lung cancer risk was found only by the combined exposure to arsenic and cadmium suggesting that cadmium is carcinogenic only in the presence of concomitant exposure to arsenic. Cd exposure in the industry has also been linked to prostate and renal cancer, but this linkage is much weaker than that for lung cancer. In a Cd-polluted region in China, an association between urinary Cd and raised serum concentration of prostate-specific antigen has been found suggesting a possible implication of Cd in prostate carcinogenesis (Zeng et al. 2004). Garcia-Morales et al. (1994) observed an association between cadmium exposure and human breast cancer, and it has been suggested that the effects of cadmium are mediated by the estrogen receptor independent of estradiol. However, the mechanism of Cd carcinogenesis remains largely unknown. Since the metal is not strongly genotoxic and does not cause direct genetic damage, epigenetic mechanisms and/or indirect genotoxic mechanisms such as a blockage of apoptosis, alterations in cell signaling, or inhibition of DNA repair might be involved (Waalkes 2003).

4.1.4

Respiratory Effects

There have also been studies examining the role of cadmium in the development of chronic obstructive pulmonary disease (COPD) in smokers. Inhalation causes respiratory stress and injures the respiratory tract (ATSDR 1999). Emphysema, anosmia, and chronic rhinitis have been linked to high cadmium concentrations in polluted air.

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A single acute exposure to high levels of cadmium can result in long-lasting impairment of lung function (Calabrese and Kenyon 1991). The most recent study showed that current and former smokers had higher body burdens of cadmium than non-smokers; the body burden of cadmium was related to lung injury (Mannino et al. 2004). Effect of chronic occupational exposure to cadmium includes chronic rhinitis, destruction of the olfactory epithelium with subsequent anosmia as well as the development of bronchitis (ATSDR 1999; Drebler et al. 2002).

4.1.5

Gastrointestinal Effects

Cadmium ingested in high doses irritates the gastric epithelium. The most common way that acute poisoning via cadmium ingestion occurs is consumption of acidic food or beverages improperly stored in containers with a cadmium glaze. The symptoms of severe cadmium ingestion are nausea, vomiting, abdominal cramps and pain, diarrhea, and tenesmus (Lewis 1997).

4.1.6

Reproductive Health

Cadmium causes disturbances of the male reproductive system involving changes in male sex hormones and related changes in accessory reproductive organs. Cadmium can cause acute necrosis of the testicles and may be involved in the development of prostate cancer (Goyer et al. 2004). In females high doses of cadmium inhibit ovarian progesterone biosynthesis. Cadmium may lead to hemorrhagic changes in ovaries, changes in ovarian hormonal production, and endothelial changes in the blood vessels (Henson and Chedrese 2004).

4.1.7

Developmental Effects

Cadmium does not induce birth defects in infants of women occupationally exposed to cadmium. However, there are reports in Japan of increased rates of preterm delivery in women that have higher urinary cadmium levels than mothers with lower levels. These preterm infants had birth weights that were lower than those of newborns of unexposed women, but this may be due to early deliveries (Nishijo et al. 2002). Tabacova et al. (1994) by studying a group of pregnant women residing in the vicinity of a copper smelter suggested that exposure to metals during gestation could enhance the development of pregnancy complications by increased lipid peroxidation. At this time, the evidence of cadmium’s effects on pregnancy is inconsistent and requires further investigation.

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Health Impacts of Cadmium on Animals

Cadmium is toxic to all animal species and is accumulated in the kidney and to a lesser extent in the liver. Alsberg and Schwartze (1919) and Schwartze and Alsberg (1923) reported various morphological changes in organs of a variety of vertebrates including birds and dogs and mentioned that Cd intoxication can lead to kidney, bone, and pulmonary damage. Cadmium exposure has been associated with nephrotoxicity, osteoporosis, neurotoxicity, carcinogenicity, genotoxicity, teratogenicity, and endocrine and reproductive effects (EFSA 2009). In ruminants high blood Cd concentrations have been reported in areas of high Cd exposure (Lane et al. 2015). Powell et al. (1964) reported that in ruminants, administration of a high dose of Cd (640 mg/kg) displayed severe sign of Cd toxicity. These included rough coat hair, dry scaly skin, dehydration, mouth lesions, shrunken scaly scrotum, sore and enlarged joints, and some atrophy of hind limb muscles. In rats mild testicular changes were seen after oral administration of 50 mg of cadmium per kg of body weight for 15 months (Krajnc 1987). Waalkes and Rehm (1994) concluded that cadmium given orally to rats caused tumors of the prostate, testis, and hematopoietic system. Manifestation of toxicity varies considerably, as depending on dose and time of exposure, species, gender, and environmental and nutritional factors. Subsequently, large differences exist between the effects of a single exposure to a high concentration of cadmium and chronic exposures to lower doses.

4.3 4.3.1

Toxicity of Cadmium in Plants Effect of Cd on Growth and Development

Cd toxicity results in inhibition of lateral root development, browning, rolling and twisting of main root and reduced stem elongation (Yadav 2010; Rascio and NavariIzzo 2011). Alteration in chloroplast ultrastructure, low contents of chlorophylls, and reduced photosynthetic activity has been observed in plants due to cd exposure (Liu et al. 2010; Miyadate et al. 2011). Treatment of rice seedlings with Cd led to inhibition of root growth and alterations in their morphogenesis (Rascio et al. 2008). Acute poisoning in roots of pea plant caused very rapid disorder in mitosis, resulting in unusual number of nucleus populations in the differentiated roots, even after 24 h of treatment (Fusconi et al. 2006, 2007). Seth et al. (2008) reported the inhibition of mitotic index, induction of chromosome aberration, mitotic aberrations, and micronucleus formation after 24 h of treatment with Cd in Allium cepa. In addition, damage to the DNA in root cap cells has been found.

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Effects of Cd on Photosynthesis

In many crop species, maize, pea, barley (Popova et al. 2008), mung bean (Vigna radiata) (Wahid et al. 2008), and wheat (Moussa and El-Gamal 2010), the evidence showed that photosynthesis was inhibited after both long-term and short-term Cd exposure. The primary sites of action of Cd are photosynthetic pigments, especially the biosynthesis of chlorophyll (Baszynski et al. 1980) and carotenoids (Prasad 1995). Cd caused chlorosis of leaves by decreasing chloroplast density, and the Cd-induced decrease in pigment content was more powerful at the leaf surface (stomatal guard cells) than it was in the mesophyll. Besides this, in Cd-treated leaves, change in cell size and reduction in stomatal density in the epidermis were observed. This showed that Cd might interfere directly with chloroplast replication and cell division in the leaf. Cadmium influences two key enzymes of CO2 fixation: ribulose-1,5-bisphosphate carboxylase (RuBPCase) and phosphoenolpyruvate carboxylase (PEPCase). Cd ions damage RuBPCase structure by substituting for Mg ions, which are important cofactors of carboxylation reactions and thus lower its activity. Cd can also shift RuBPCase activity toward oxygenation reactions thereby reducing the efficiency of photosynthesis (Siedlecka et al. 1998). Cd causes an irreversible dissociation of the large and small subunits of RuBPCase, thus leading to loss of enzyme activity and inhibition of C3 cycle (Stiborova 1988; Malik et al. 1992). Cd also affects the water-oxidizing complex of PSII by replacing the Ca2+ in Ca/Mn clusters of oxygen-evolving center (Sigfridsson et al. 2004). Cd induces changes in lipid and fatty acid composition of membranes and thus produces alterations in the functionality of membranes (Popova et al. 2009).

4.3.3

Effect of Cd on Mineral Nutrition

Cd has been shown to interfere with the uptake, transport, and use of several elements (Ca, Mg, P, and K) and water by plants (Das et al. 1997). In sugar beet, deficiency of Fe in roots induced by Cd was observed (Chang et al. 2003). Uptake of P, K, S, Ca, Zn, Mn, and B in pea plants was inhibited strongly by Cd exposure (Metwally et al. 2005). Guo et al. (2007) observed decrease in the concentrations of P, K, Ca, Mg, Cu, Fe, Mn, Zn, Mo, and B in roots of barley plants when they were treated with 1.0 μM Cd. Cd also inhibits nitrate reductase activity in the shoots and is thus responsible for reduced absorption of nitrate and its transport from roots to shoots (Hernandez et al. 1996). In soybean plants Cd treatments decreased nitrogen fixation and primary ammonia assimilation in nodules (Karina et al. 2003). Effect of Cd on mineral nutrition is due to inhibition of H+ATPase in maize root cells. H +ATPase functions as an ion transporter across the plasmalemma. Cd, which causes a decrease in activity of H+ATPase, might inhibit absorption of some essential elements (Astolfi et al. 2005).

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Effect of Cd on ROS Generation

Heavy metals cause oxidative damage to plants, either directly or indirectly through reactive oxygen species (ROS) formation. Cd is a non-redox metal that does not produce ROS such as the superoxide anion (O2•–), singlet oxygen (1O2), hydrogen peroxide (H2O2), and hydroxyl radical (OH•) but generates oxidative stress by interfering with the antioxidant defense system (Benavides et al. 2005; Gratão et al. 2005). However, Cd could lead to the generation of ROS indirectly by production of a disturbance in the chloroplasts. In addition, other reports suggested that Cd may stimulate the production of ROS in the mitochondrial electron transfer chain (Heyno et al. 2008). The manifestations of ROS damages in plants involve lipid peroxidation, protein peroxidation, and DNA damage. Cd produced an enhancement of lipid peroxidation in Phaseolus vulgaris (Chaoui et al. 1997), Helianthus annuus (Gallego et al. 1996), and Pisum sativum (Lozano-Rodriguez et al. 1997). DNA damage caused by Cd involved destruction of nucleic acids, cell membrane, lipids, and proteins, reduction of protein synthesis, and damage of photosynthetic proteins, affecting growth and development of the whole organism (Gill and Tuteja 2010; Kranner and Colville 2011). Cadmium toxicity leads to variation in the activity of scavengers of superoxide and hydrogen peroxide such as superoxide dismutase (SOD), catalase (CAT), ascorbate peroxidase (APX), peroxidases (POD), and glutathione reductase (GR) (Sandalio et al. 2001; Milone et al. 2003). Besides this proline accumulation appeared to be a suitable indicator of heavy metal stress. Islam et al. (2009) reported that tobacco cells exposed to Cd treatment accumulated high levels of proline, and by this way they can alleviate the inhibitory effect of Cd on cell growth.

4.3.5

Effect of Cd on Stress Proteins

Heat-shock proteins (HSPs) are presently known as proteins that have functions to resist stress in eukaryotes. In Cd-treated maize plants, a synthesis of 70 kDa phosphoprotein (HSP) was reported by Reddy and Prasad (1993). In Cd-treated pea plants, pathogen-related proteins PrP4A and HSP71 were found, and they probably serve to protect cells against damages induced by Cd (Rodríguez-Serrano et al. 2006). In wheat, seedlings treated with 50 μM CdCl2 for 48 h, a 51 kDa soluble protein was found. This protein was designated as a Cd stress-associated protein. It was generated mainly in the root tissue of treated and control seedlings and located below the plasma membrane and outer periphery of the tonoplast (Mittra et al. 2008). In poplar (Populus tremula ) exposed to Cd for a short-term (14 days) or a longerterm (56 days) treatment, it was found that stress-related proteins, like HSPs, proteinases, and pathogenesis-related proteins, increased in abundance in leaves, whereas most of the proteins from the primary metabolism (tricarboxylic acid cycle, nitrogen metabolism, and sulfur metabolism) were severely decreased in abundance (Kieffer et al. 2009).

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5 Bioremediation Approaches for Cadmium Detoxification from Contaminated Sites Cadmium can be effectively removed by various physicochemical methods such as absorption, oxidation/reduction, flocculation, cation/anion exchange, and precipitation. However they have some drawbacks like complexity, uneconomical, and lack of public acceptance. So efforts are focused on bioremediation process being a suitable method for removal of cadmium.

5.1

Bioremediation

Bioremediation is biological techniques that completely remove the toxic contaminants from the soil, water, and air by use of microorganisms (bacteria, fungi, algae, etc.) and green plants (phytoremediation). It is a natural method with wider application in the realm of environment. Mueller et al. (1996) define bioremediation as the process where organic waste is biologically degraded under controlled condition to an innocuous state or to level below concentration limits established by regulatory authorities. Bioremediation can be classified into in situ or ex situ. In situ techniques involves remediation at the site of contamination while in ex situ contaminants are removed from the original site and treated elsewhere, it is called as ex situ. Examples of bioremediation technologies are bioventing, bioleaching, composting, landfarming, bioreactor, bioaugmentation, biostimulation, and phytoremediation. It has high public acceptance due to its relatively low-cost and low-technology technique. In the process of bioremediation, microorganism may be indigenous to the contaminated site or may be isolated from other sources and brought to the contaminated site for heavy metal removal. To make the process effective, microorganism must enzymatically attack on the pollutants and convert them into harmless products (Vidali 2001). Phytoremediation is the use of green plants to remove pollutants from the environment or to render them harmless (Cunningham and Berti 1993). Mixing the use of resistant plants and the application of microorganisms with their beneficial effects to plants and to soil could represent a valid tool for soil remediation (Wenzel et al. 2003). The main goal of bioremediation is to remove pollutants from the soil and water and to restore their capacity and functional potential.

5.1.1

Bioremediation of Cadmium Using Bacteria

Microbial bioremediation is an effective approach for removal of heavy metals like lead, chromium, cadmium, and zinc. Bacteria are small in size and ubiquitous and grow in wide range of condition, and their ability to grow in controlled condition makes them suitable candidates for bioremediation process. Use of bacteria for

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bioremediation of Cd has been investigated by many researchers. Stationary cells of Bacillus thuringiensis exhibited high potential for Cd binding (El-Helow et al. 2000), while Ralstonia sp. showed high affinity for the removal of Cd+2 (Chompoothawat et al. 2010). Cupriavidus taiwanensis KKU2500-3 was isolated from cadmium-contaminated soil and found to be producing more thi-containing protein. It also decreases the accumulation of Cd in rice plant by 61% (Siripornadulsil and Siripornadulsil 2013). Pseudomonas sp. M3 (strain RZCD1) was found to be resistant against cadmium and useful for bioremediation of cadmium-contaminated industrial wastewater (Abbas et al. 2014). Priyalaxmi et al. (2014) reported that mangrove ecosystem contains a wide microbial diversity that is helpful in the biosorption and reduction of cadmium. Escherichia coli, Salmonella typhi, Bacillus licheniformis, and Pseudomonas fluorescens isolated showed promising results for the bioremediation of heavy metals including Cd from the textile dye effluents. Flavobacterium psychrophilum and Flavobacterium sp. ARSA-103 are hexavalent Cr- and Cd (II)-resistant bacteria collected from tannery effluents which were used to remove metal-contaminated effluents. Flavobacterium psychrophilum is a fish pathogen which is resistant to heavy metals and can grow efficiently in the heavy metal stressed environment. Therefore it can be used to remediate heavy metals so as to make the environment heavy metal pollution-free (Mukherjee et al. 2015). From activated sludge many different species of bacteria have been isolated that produce metabolic products called extracellular polymeric substances (EPS) which help in heavy metal removal by formation of complexes between carboxyl, hydroxyl, and phenolic surface functional groups of the extracellular polymeric substances and heavy metals (Yuncu et al. 2006). Biosorption of metal by bacteria is dependent on external factors such as pH, other ions in bulk solutions (which may be in competition), organic material in bulk solution, and temperature (Comte et al. 2008).

5.1.2

Bioremediation of Cadmium Using Fungi

Fungi show a great affinity toward metal ions as compared to other microorganisms. They can accumulate heavy metals by means of physicochemical and biological mechanisms from their external environment (Cabuk et al. 2005; Preetha and Viruthagiri 2005) (as shown in Fig. 15.4). Fungi have the capacity to bind with cadmium by the help of cysteine-rich protein called metalloprotein (Margoshes and Vallee 1957). Aspergillus fumigatus fungal isolate has good biosorption capacity toward removal of selected heavy metals from contaminated sites (Iram et al. 2013). Aspergillus fumigatus was isolated from the polluted soil, and its biosorption capacity against heavy metals, viz., Pb, Cd, Cr, Cu, Zn, and Ni, was checked. Maximum biosorption rate (76.07%) at 800ppm was exhibited by fungal isolate K3 of Aspergillus fumigatus and showed higher accumulation of metal lead (Shazia et al. 2013). Removal of Cd from wastewater by Aspergillus fumigatus was tested at different pH, inoculum sizes, and incubation times. It could remove 76% of Cd with

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Fig. 15.4 Showing metal–microbe interaction

inoculum size of 8% at pH 6 after incubation time of 144 h (Talukdar and Kumar 2015). All filamentous fungi belonging to Zygomycetes group showed a great ability for heavy metal removal. These fungi contain large amounts of polymers of chitosan, N-acetyl chitin, and deacetylated glucosamine on their cell wall. These polymers have large amounts of potential binding sites in the form of hydroxyl, amine, and carboxyl groups. The amine group containing nitrogen atom and the hydroxyl group containing oxygen atom have ability to bind a proton or a metal ion, respectively, and helps in heavy metal removal (Das et al. 2007).

5.1.3

Bioremediation of Cadmium Using Algae

Algae play an important role in bioremediation and act as important bioremediation agents. A number of microalgae, viz., Chlorella sp., Ankistrodesmus sp., and Enmosphaera sp., have been used for heavy metal removal from the aqueous medium (Rai et al. 1981). Aquatic plants mainly micro and macro algae have the ability to absorb metals and take up toxic elements from the environment or transform them into less harmful ones (Mitra et al. 2012). Heavy metals are removed by microalgae from polluted water by two mechanisms: first is metabolism dependent uptake into the cells at low concentration, and second is by non-active adsorption process known as biosorption (Matagi et al. 1998). Microalgae such as Chlorella, Scenedesmus, and Arthrospira also showed great potential to remove and incorporate heavy metals like lead (Aksu and Kustal 1991), cadmium, nickel, or

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mercury (Chen et al. 1998; Travieso et al. 1999) present in effluents. Chlorella is found to grow in a mixotrophic environment (Karlander and Krauss 1996). Sheng et al. (2007) used two locally harvested brown marine algae Sargassum sp. and Padina sp. for the removal of cations (Cd2+ and Cr3+) and an anion (Cr2O72-) from dilute aqueous solutions. Pithophora sp. showed bioaccumulation potential for Cr, Cd, and Pb removal at various concentrations (2, 5, 10, 20, and 30 ppm). The metal uptake was increased with increase in the incubation period. Marine algae such as Ascophyllum and Sargassum are effectively used in the biosorption of pollutants (Yu et al. 1999) and reduced the concentrations of heavy metals in the polluted environment to a very low level. In recent years, several species of the green algae Enteromorpha and Cladophora have been utilized to measure heavy metal levels in many parts of the world. Macroalgae have been extensively used to measure heavy metal pollution in marine environments throughout the world. Cyanobacteria (blue-green algae) are the most primitive photosynthetic prokaryotes and appeared on the planet during the Precambrian period. Cyanobacteria occupy a vast array of habitats and are susceptible to various physical and chemical changes (e.g., light, salinity, temperature, and nutrient composition) (Boomiathan 2005; Semyalo 2009). Cyanobacteria have great potential to be used in wastewater and industrial effluent treatment, bioremediation of aquatic and terrestrial habitats, chemical industries, food and feed, and fuel, as biofertilizers, etc. (Cairns and Dickson 1971). Chojnacka et al. (2005) reported the biosorption performance of Cr+3, Cu+2, and Cd+2 ions by blue-green algae Spirulina. The use of immobilized algae in the removal of heavy metal is an efficient method and offers significant advantages in bioreactors (Hameed and Ebrahim 2007). Huge loads of wastes from industries, domestic sewage, and agriculture practices find their way into rivers and ponds resulting in large-scale deterioration of water quality leading to the availability of potable water. There is an urgent need to screen and develop efficient alga for the bioremediation of wastewater.

5.1.3.1

Heavy Metal Accumulation and Tolerance Mechanisms in Algae

Synthesis of metallothioneins (MTs) or phytochelatins (PCs), complex formation, compartmentalization, exclusion, and translocation into vacuoles are various defense mechanisms that respond against heavy metals stress (Mejáre and Bülow 2001). Thalassiosira pseudonana and Thalassiosira weissflogii are marine algae that produce phytochelatins in large quantities due to the higher activity of phytochelatin synthase, which has greater affinity for the glutathione substrate or metal ions (Ahner et al. 2002). Ligands like histidine (His), citrate, oxalate, malate, nicotianamine, and phosphate derivatives (phytate) play an important role in detoxification and tolerance. Algae have high negatively charged surface (cell wall components) which leads to increased adsorption, phytosorption, and affinity of heavy metal cations in wastewater treatment (Sekabira et al. 2011). Micrasterias denticulata (Volland et al. 2012) and Spirogyra hyaline (Kumar and Oommen 2012) showed promising results for sequestration of cadmium ions from contaminated sites. The principal

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mechanism of metallic cation sequestration involves the formation of complexes between a metal ion and functional groups on the surface or inside the porous structure of the biological material. The carboxyl groups of alginate play a major role in the complexation. Different species of algae and the algae of the same species may have different adsorption capacities. Algae have many features which make them suitable candidates for removal and concentration of selective heavy metals. They have high tolerance toward heavy metals, ability to grow both autotrophically and heterotrophically, large surface area/ volume ratios, phototaxy, phytochelatin expression, and great potential for genetic manipulation (Cai et al. 1995). Therefore, there is urgent need to improve algae using genetic engineering to develop transgenic species with high accumulation potential for heavy metals like cadmium by overexpressing phytochelatins and metallothioneins which form complexes with cadmium and translocate them into vacuoles to increase phytoaccumulation. Algae offer a cost- effective, less invasive, and potentially more effective means of addressing existing heavy metal contamination than those currently practiced.

5.1.4

Phytoremediation

Phytoremediation is an environmentally friendly and cost-effective green technology which utilizes hyperaccumulator plants to extract heavy metals from soil. It is a suitable practice to remove both organic and inorganic pollutants present in soil, water, and the air (Gratão et al. 2005). Various mechanisms used to remediate soils contaminated with cadmium are phytoextraction, phytostabilization, and rhizofiltration.

5.1.4.1

Phytoextraction

The use of plants for extraction of heavy metals from contaminated soils is known as phytoextraction. Phytoextraction efficiency depends on the chemical property of the element removed, its uptake, translocation and accumulation by plants. It is one of the phytoremediation strategies based on the use of green plants to remove cadmium from soil (Salt et al. 1995). Strczynski (2000) recommended hemp and flax for cadmium phytoextraction. Sunflower (Helianthus annuus L.) shows high tolerance to heavy metals and is used in phytoremediation studies (Schmidt 2003; Tang et al. 2003; Pilon-Smits 2005). It can be used for cadmium removal from the high polluted environment. Cd removal efficiency of sunflower could be 57–72%. After phytoremediation sunflower plant must be not used in food and feed. Hadi et al. (2015) used Ricinus communis to treat cadmium contaminants in hydroponic condition. Cd content in plant tissue was found to be increased as the dose of Cd increased. The maximum uptake was found in root followed by leaf and then stem.

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5.1.4.2

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Phytostabilization

The use of indigenous hyperaccumulator plants to reduce the migration of Cd through the soil medium is known as phytostabilization (Robinson et al. 2006). Hamzah et al. (2016) also reported the potential of using indigenous hyperaccumulator plants to stabilize heavy metals, and such remediated contaminated soil could be used as agricultural soil again. Among the Cd hyperaccumulators, Solanum nigrum, Populus sp., Salix calodendron, and Arabis paniculata may be good candidates for field conditions due to their potentially higher biomass (French et al. 2006; Ji et al. 2011). Catharanthus roseus (a valued medicinal plant) was exposed to different concentrations of heavy metals like CdCl2 and PbCl2, and its bioaccumulation efficiency was observed. AAS analyses of leaves of treated plants show 5–10% accumulation of cadmium, but there was no accumulation of lead at all (Pandey et al. 2007). Mojiri (2011) found the potential of corn (Zea mays) for phytoremediation of soil contaminated with cadmium and lead. Result indicated that corn is an effective accumulator plant for phytoremediation of cadmium- and lead-polluted soils. Physalis minima a fast growing weed plant has been studied for extraction of heavy metals like Cd and Cr. The bioconcentration factor (BCF) and translocation factor (TF) showed that Physalis minima is a high heavy metal accumulator (Subhashini and Swamy, 2013). Alfalfa had a better ability of accumulating Cd, and results indicate that alfalfa had the higher biomass when grown in soil polluted by different concentrations of heavy metals (Wang et al. 2015). Kathal et al. (2016) found that Cd concentration was found highest in plant as compared to the seeds in Brassica juncea grown in Cd and control soil under greenhouse condition.

5.1.4.3

Rhizofiltration

It is a technique used to remove heavy metal (Cd) from aquatic environment using plants. Phytoremediation can be enhanced by using chelating agents such as citric acid, EDTA, DTPA, and EGTA (Luo et al. 2006). However, excessive use of chelating agents in the field results in secondary pollution of soils, and leaching of chelating agents may risk groundwater contamination and increase the cost of phytoremediation (Robinson et al. 2006).

5.1.4.4

Mechanism of Heavy Metal Detoxification and Uptake by Plants

Heavy metal chelation by high-affinity ligands in cytosol followed by compartmentalization of the metal–ligand complex is a very important mechanism of heavy metal detoxification and tolerance by plants. Chelators may be natural, i.e., amino acids, organic acids, metallothionein (MT), and phytochelatins (PC), or synthetic, i.e., ethylene glycol tetra acetic acid (EGTA) and ethylene diamine tetra acetic acid (ETDA). Addition of EDTA has been found to accelerate the uptake of Cd, Zn, Cu,

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and Ni by the plant (Raskin et al. 1997). It has been studied that Cd stimulates the synthesis of phytochelatins (PCs), which form a Cd–PC complex of low molecular weight. Cd–PC complex is then transported into vacuole by Cd/H antiport and ATP-dependent PC transport protein. The gene (Hmt1) is responsible for transporting the Cd–PC complex into the vacuole (Sriprand and Murooka 2007). Metallothioneins (MTs) are cysteine-rich metal-chelating proteins in plants that help to sequester toxic heavy metal such as cadmium, Hg, Cu, Zn, etc. However there is requirement that plant biomass must be harvested and removed, followed by metal reclamation or proper and safe disposal of the biomass.

5.2

Role of Environmental Biotechnology and Genetic Engineering in Improving Efficient Bioremediation

Various biotechnological approaches such as bioleaching, bioremediation, bioencapsulation, biostimulation (stimulating living microbial population), biomineralization (synthesis of mineral by living organisms or biomaterials), mycoremediation (stimulation of fungi), cyanoremediation (stimulating algal biomass), rhizoremediation (plant and microbes) biosorption (dead microbial and renewable agricultural biomass), and bioextraction are advanced bioremediation processes to remove heavy metals from the polluted environment. Environmental biotechnology plays an important role in bioremediation of cadmium from the environment. Understanding the mechanism through which microorganisms capture heavy metals, i.e., metabolism and detoxification pathways, may lead to the solution of environmental problem associated with heavy metals in a maximum efficiency. Genes responsible for metal uptake, translocation, sequestration, and bioaccumulation have been identified. Transfer of these genes into organisms will result in the generation of transgenic organism having greater ability for metal accumulation and detoxification. Large numbers of bacteria and algae have been genetically engineered to remove selective heavy metals from polluted soil and water by overexpressing a heavy metal-binding protein, such as metallothionein, along with a specific metal transport system (Chen and Wilson 1997; Liu et al. 2011). Various peptides consisting of metal-binding amino acids (histidine and cysteine residues) have been studied in bacteria to increase heavy metal accumulation (Liu et al. 2011). Bang et al. (2000) reported that expression of the thiosulfate reductase gene from Salmonella typhimurium in Escherichia coli led to increase of the efficiency of removal of heavy metals from solution and accumulation of cadmium up to 150 mM in 98% cells. In genetically engineered Ralstonia eutropha, a mouse metallothionein fusion protein (MTβ) was inserted that displays on the surface of the cell. The MTβ-bearing cells were able to sequester cadmium from a contaminated soil, partially ameliorating the effects of metal contamination and promoting growth of metal-sensitive tobacco plants (Lovley and Lloyd 2000) (Fig. 15.5).

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Fig. 15.5 Engineering a soil bacterium for enhanced sequestering of toxic heavy metals. (Adapted from Lovley and Lloyd 2000)

5.2.1

Role of Genetic Engineering in Developing Plants for Phytoremediation

Transfer of selective genes into candidate plants results in the development of transgenic plant with enhanced ability for metal uptake, transformation, and accumulation from the polluted environments. Genes from foreign source have been isolated and transferred into plants like Nicotiana tobacuum, Brassica juncea, Arabidopsis thaliana, Lycopersicon esculentum, etc. to enhance the phytoremediation potential (Abhilash et al. 2009). The pathway through which plants detoxify heavy metals is sequestration with heavy-metal-binding peptides called phytochelatins or their precursor, glutathione (GSH). Biosynthesis of phytochelatins (PCs) in plants has been modified to increase the heavy metal accumulation. Brassica juncea (Indian mustard) transgenic plants were developed with overexpressing gshII (gene encoding glutathione synthetase) which accumulated

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significantly more Cd than the wild type (Zhu et al. 1999). Transgenic plants expressing glutathione (GSH) have great efficiency for enhancing the Cd phytoextraction from polluted soils and wastewater. These plants may also show increased tolerance and accumulation of other heavy metals, because PCs are thought to play a role in tolerance of a wide range of heavy metals. Simultaneous expression of Allium sativum phytochelatins synthase (AsPCS1) and yeast cadmium factor 1 (YCF1) in Arabidopsis led to increase the tolerance to Cd and As and have higher amounts accumulation of these metals as compared with single-gene transgenic lines and wildtype (Jiangbo et al. 2012). Transfer of human MT-2 gene to Nicotiana tobacuum, Brassica sp., and Arabidopsis thaliana resulted in the transgenic plant with enhanced tolerance and accumulation of cadmium. Development of transgenic cauliflower (Brassica capitata) having 16-fold higher accumulation of cadmium by transferring CUP 1 gene (Vasavi et al. 2010).

5.3

Factors Affecting Bioremediation of Cadmium

Bioremediation processes are a complex system affected by many environmental factors such as nutrients, type of soil, pH, moisture content, temperature, and the presence of oxygen. (a) Nutrients: Carbon is the most important element for living forms and is required in large quantities than other elements. Hydrogen, nitrogen, and oxygen also constitute about 95% of the weight. Contaminated sites are rich source of organic carbon which is depleted during microbial metabolism. So there is need of nitrogen, phosphate, and potassium in the contaminated site which stimulate the cellular metabolism and growth of the microorganisms to augment the bioremediation. Bioremediation processes require 30:1 carbon-to-phosphorus ratio and 10:1(C:N) carbon-to-nitrogen ratio. However biodegradation of contaminated soil is much faster with higher C:N (25:1) ratio (Atagana et al. 2003). (b) Soil structure: Soil can be classified into silt, sand, and clay depending upon the textures. Granular and well-structured soil provides effective delivery of air, water, and nutrients to the microorganisms during in situ bioremediation process. To improve soil structure, materials like gypsum or organic matter can be added. Low soil permeability may not be suitable as it can impede movement of nutrients, water, and oxygen into soils. (c) pH: pH from 5.5–8.0 is found to be suitable for the growth of microbes as well to destroy the contaminants (Vidali 2001). (d) Moisture content: Water is the primary factor in determining the dielectric constant of soil and other mediums. Soil moisture content generally ranges from 25 to 28% (Vidali 2001)

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(e) Temperature: Temperature affects biochemical reaction rates, and the rates double for each 1
C rise in temperature; optimum temperature range is from 15 to 45
C. (f) Oxygen: It plays an important role in initial breakdown of contaminants and in the growth of microorganisms. It also determines whether the process is aerobic or anaerobic in nature.

6 Conclusion Cadmium due to its nonbiodegradable nature persists in the environment and becomes part of the food chain. Higher concentration of cadmium in environment leads to various types of toxicity to humans, plants, and animals. Bioremediation now a day is used for cleanup of metal-contaminated and metal-polluted ecosystem. Microorganisms like algae, bacteria, and fungi bioremediate heavy metals by various processes like complexation, biosorption, absorption, flocculation, cation/anion exchange, and oxidation–reduction reaction. The isolation of heavy metal-resistant microorganisms and the understanding of the mechanisms they use in order to remove pollutants may contribute to the development of improved bioremediation processes. The use of resistant plants along with application of microorganism in the soil would become a good mixing technique for better remediation. Biotechnological approaches like formation of genetically modified plants and microbes which tolerate high level of cadmium are key factors in the future to reduce the environmental burden of toxic substances. Due to industrialization, heavy metal pollution is increasing at an alarming rate, so development of cost-effective and eco-friendly remediation methods for heavy metal-polluted sites is an urgent need of the hour.

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Chapter 16

Cyanobacteria: The Eco-Friendly Tool for the Treatment of Industrial Wastewaters Sharma Mona, Virendra Kumar, Bansal Deepak, and Anubha Kaushik

Abstract As the earth’s human population has increased, an enormous industrial growth has taken place throughout the world. Industry is the most flagrant abuser of water quality. It discharges polluted water having the pollution strength of at least double the sewage of all municipalities combined. Industrial effluents are the most important sources of toxic contaminants in any environment. Discharge of untreated industrial wastewater into aquatic bodies is posing a serious threat to the water resources. It should be treated before discharge into the natural water bodies. Recently, there has been increasing interest in cyanobacteria for the treatment of industrial wastewater (phycoremediation) since they possess many advantages over other microorganisms. Cyanobacteria are photoautotrophic in nature and have the ability to fix atmospheric nitrogen enabling them to be productive. In this way cyanobacteria are inexpensive; they can maintain their growth without the addition of nutrients. They are known to inhabit in various aquatic and highly polluted environment and acquired natural resistance against environmental pollutants. Cyanobacteria are efficient in the assimilation of organic matter and have high biodegradation, transformation, and biosorption capability of pollutants present in industrial wastewater. In addition, cyanobacteria have a great potential as a source of biofuels, bio-fertilizers, animal feed, polysaccharide production, etc. which makes them a viable and sustainable approach for the treatment of industrial wastewater and can be improved through genetic engineering technologies. This chapter represents the biodiversity of cyanobacteria and their potential application for the removal of S. Mona (*) Department of Environmental Sciences, Central University of Haryana, Pali, Haryana, India Department of Environmental Science and Engineering, Guru Jambheshwar University of Science & Technology, Hisar, Haryana, India V. Kumar Department of Environmental Sciences, Central University of Haryana, Pali, Haryana, India B. Deepak JBM Group, Gurugram, Haryana, India A. Kaushik University School of Environment Management (USEM), Guru Gobind Singh Indraprastha University, New Delhi, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_16

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heavy metals, dyes, crude oil, and pesticides from the wastewaters of different industries followed by a critical overview of their utilization, suitability, biomass production, and potential in bioremediation of industrial wastewater. Keywords Phycoremediation · Biosorption · Cyanobacteria · Industrial wastewaters

1 Introduction Cyanobacteria are also called blue-green algae; they are simple, have primitive life forms, and are closely related to bacteria. Their size is larger than bacteria and they do photosynthesis like algae. Morphologically, cyanobacteria occurs in the form of single cell structure, filamentous or colonial forms. Some filamentous species have special cells called heterocysts, which fix atmospheric nitrogen (N2) into chemical forms. They produce energy through photosynthesis, but do not have nucleus or membrane-bound organelles, like chloroplasts. Cyanobacteria have cell structure like gram-positive bacteria, featuring a thick peptidoglycan layer that lies between the cell membrane and outer membrane. Cyanobacteria have layers of thylakoids, which are membrane-bound structures where the “light” photosynthetic reactions occur. Finally, at the center of the cell lies the DNA of the cell as well as the cyanobacteria’s carboxysomes. Cyanobacteria lack flagella, making it difficult for them to move under their own power. However, some species can move by gliding along surfaces. The cyanobacteria contain chlorophyll a and other pigments that are used for photosynthesis. The chlorophylls and pigments efficiently capture specific wavelengths of light, transferring the light energy to the cell. Phycocyanin, a blue color pigment, is the origin of the cyano in the name cyanobacteria, and this pigment, in conjunction with the green chlorophyll, is the source of the common name bluegreen algae. However, not all cyanobacteria are blue; some common forms are red or pink from the pigment phycoerythrin.

2 Classification of Cyanobacteria Cyanobacteria seem like algae and bacteria and have been classified in the kingdom Plantae and Monera. In kingdom Plantae, cyanobacteria are included in division Cyanophyta or class Cyanophyceae or Myxophyceae, while in kingdom Procaryotae, they are included in division Cyanobacteria (Murray 1968) or order

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Cyanobacteriales (Gibbons and Murray 1978). Stanier et al. (1971) coined the name cyanobacteria, and classification of cyanobacteria alongside the bacteria was considered as a truer representation of their phylogenic position. In 1977, Stanier and Cohen-Bazire stated that the only logical taxonomic treatment of cyanobacteria was to place them in the super kingdom Prokaryotae as a division, class, or order of bacteria: the Cyanobacteria. Traditional classifications are based upon morphological features of the organisms. The main morphological features used in classification of cyanobacteria are growth form (unicellular, colonial, filamentous), shape of the colonies, shape of the filaments, sheath (present or absent), cell differentiation (presence of heterocysts or not), size and shape of vegetative cells, branching (present or absent), and combinations of these different characters. Cyanobacteria include around 5 orders, 150 genera, and around 2000 species. The five orders of cyanobacteria recognized in the classic botanical taxonomic scheme are: 1. Chroococcales: These are the coccoid cells that are reproduced by binary fission or budding. The representative genera are Aphanocapsa, Aphanothece, Gloeocapsa, Merismopedia, Microcystis, Synechococcus, Synechocystis, etc. 2. Pleurocapsales: These are coccoid cell aggregates, or pseudo-filaments in nature that are reproduced by baeocytes; Chroococcidiopsis and Pleurocapsa are the representative genera of this order. 3. Oscillatoriales: They have uniseriate filaments, without heterocysts or akinetes for example Lyngbya, Leptolyngbya, Microcoleus, Oscillatoria, Phormidium, Planktothrix, etc. 4. Nostocales: Characteristics of this order are filamentous cyanobacteria that divide in only one plane, with heterocysts, false branching in genera such as Scytonema, Anabaena, Aphanizomenon, Calothrix, Cylindrospermopsis, Nostoc, Scytonema, Tolypothrix, etc. 5. Stigonematales: Stigonematales divide in more than one plane with true branching and multiseriate forms with specialized cells known as heterocysts. Mastigocladus and Stigonema are the representative genera of this group. The classical conventions used in the classification of cyanobacteria frequently proved unsatisfactory, and a symptom of this undesirable stage of affairs is that conventional taxonomy frequently appears to have been ignored by physiologists and biochemists (Whitton 1969). Because, they are morphological simple and occur in polymorphism which aggravated the problem of identification of the organisms based on morphological characters and sometimes variations (polymorphism) in response to changes in environmental conditions. Recent trends on classification of cyanobacteria are based on its dynamic characters, physiological properties, fatty acid composition, presence of pigments, isozymes, genome size, deoxyribonucleic acid base composition, sequence-specific deoxy ribonucleases, DNA-DNA hybridization, etc (Figs. 16.1 and 16.2).

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Bacteria

Purple bacteria Cyanobacteria

Flavobacteria Spirochetes

Thermotogales

Archaea

Eukarya

Euryarchaeota Methanosarcina Animals Methanobacterium Halophiles Gram-positive Fungi Methanococcus Thermo- Slime Chlamydiae Thermococcus molds plasma Plants Crenarchaeota Entamoebae Ciliates Thermoproteus Pyrodictium Flagellates Deinococci Trichomonads Green nonsulfur bacteria Microsporidia

Aquifex

Diplomonads

Tree of Life Fig. 16.1 Tree of cyanobacterial life. (Adapted from http://bioweb.uwlax.edu/bio203/2011/fedor_ kara/gallery.htm)

Fig. 16.2 Nostoc, filamentous cyanobacteria with heterocyst’s under a light microscope

2.1

Diversity of Cyanobacteria

Cyanobacteria can be found in nearly all environments, and they can generally survive anywhere where there is enough sunlight for photosynthesis. They are found in a wide variety of habitats which includes ponds, rivers, lakes, estuaries, oceans, soils, and snow. They can be free-floating in the water or attached to some substrate. Cyanobacteria have great diversity from fresh water to marine water and extreme environmental condition. Cyanobacteria with greater species richness make better use of niche opportunities in an environment, allowing them to capture a higher proportion of available resources (Cardinale 2011). The great variety of

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nutrition modes displayed by cyanobacteria for acquiring carbon or nutrients help to extend their niche opportunities in changing environments. Depending on the species, their physiology, cell size, and biotic and abiotic environmental growth conditions, they can use various forms of nitrogen (elemental nitrogen, ammonium, nitrate, organic nitrogen) or particulate organic substances for growth (Raven 1997; Graneli et al. 1999). Furthermore, cyanobacteria have the capability to oxidize the pollutants such as oils, pesticides, and dyes and produce peptides that are able to bind heavy metals (Wide and Bennan 1993). They are attractive bio-agent for the bioremediation of pollutants. The mixed culture of cyanobacteria, containing various species with different metabolic abilities may improve the overall remediation capacity of cultures when supplemented with multiple resources. Cyanobacteria have wide variety of metabolisms and are growing in all habitats even in extreme environments (high temperature and alkaline environment). In natural aquatic environments, cyanobacteria are associated with strong heterotrophic bacteria. This suggests a close interaction between the various microbial communities. However, very little information is known about the nature of such associations (independence, competition, mutualism, or symbiosis) which is still debated (Fouill and Mostajir 2011). Cyanobacteria have high genetic and metabolic variability and high biomass productivity. They play an important role as primary producers in these ecosystems (Oren and Seckbach 2001). Recent studies reported large genetic diversity among cultured cyanobacteria within and between aquatic environments. This genetic diversity is not represented well by traditional morphological microscopic identifications of cyanobacteria. To date, the majority of information on genetic diversity of cyanobacteria in lakes is based on analyses of cultured cyanobacteria. Janse et al. (2004) isolated 107 cyanobacteria from 15 lakes and found 59 distinct genotypes. Altogether, these studies suggest that populations of cyanobacteria are heterogeneous both within and across aquatic environments.

2.2

Benefits of Cyanobacteria in Bioremediation

In recent years, cyanobacteria attracted considerable scientific and technical interest for biological treatment of polluted wastewater. Because cyanobacteria have a great potential for the treatment of various types of environmental contaminants such as pesticides (Megharaj et al. 1994), crude oil (Al-Hasan and Radwan 2001), phenol (Shashirekha et al. 1997), heavy metals (Singh et al. 2011), textile dyes (Mona et al. 2011), and xenobiotics (Megharaj et al. 1987). They have high growth rate and metal sorption capacity that play a potential role in the detoxification of various industrial effluents such as from oil refinery, pulp and paper, brewery, distilleries, tanneries, textile, dye, and pharmaceuticals industries. There are several advantages of cyanobacteria that are used as a biological agent for the bioremediation of wastewater. Cyanobacteria have advantages over chemical treatment as they are widespread in all conceivable system and also able to

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accumulate high levels of metal or pollutants. Cyanobacteria have rapid kinetics of the metal removal and the possibility to treat water containing multiple metals. Cyanobacteria as bioremediators have some advantages over other microorganisms because of their photoautotrophic nature and ability to fix atmospheric N2 which makes them self-sufficient for growth and maintenance and adaptability to survive in heavily polluted environments (Sokhoh et al. 1992). In addition, the potentials of cyanobacteria in the removal of nutrients from wastewater rich in nitrogenous and phosphorus compounds have been demonstrated. The ability to remove nitrogenous and phosphate ions from wastewater was observed in cyanobacteria such as Oscillatoria, Phormidium, Aphanocapsa, and Westiellopsis (Blier et al. 1995; Pouliot et al. 1989). Cyanobacteria may be used for tertiary treatment of urban and agro-industrial effluents and can help mitigate eutrophication and metal toxicity problems in aquatic ecosystems. In natural environment, many cyanobacteria form symbiotic associations with other aerobic or anaerobic microorganisms that showed degradation of hydrocarbons present in oil. In such type of association, cyanobacteria are not directly responsible for the degradation of hydrocarbons and even facilitated the degradation process by providing oxygen and nutrients to the associated oil-degrading bacteria (Abed et al. 2009). Cyanobacteria are found in ponds, lakes, water streams, rivers, and wetlands. They can easily survive the extreme environments such as hot springs, hyper-saline waters, freezing environments, and arid deserts (Singh et al. 2013). Some of the cyanobacterial strain shows potential metabolic adaptation in diverse habitats such as in the highly saline condition (Reed and Stewart 1983). The ability of cyanobacteria to survive extreme environmental conditions can be exploited for better remediation of wastewater. Cyanobacteria are characterized by the ability to oxidize oil components, pesticides, and complex organic compounds and accumulate metal ions, for example, Zn, Co, and Cu (Mohapatra and Gupta 2005). Thus, cyanobacteria are promising tool for the secondary treatment of urban, agricultural, or industrial effluents. Many of the cyanobacteria synthesize metal-binding proteins “metallothioneins” that help organisms in the sequestration of toxic metal ions. The metal uptake was observed to be dependent on various nutritional factors and the tolerance to the intracellular detoxification mechanism (Verma and Singh 1995). Certain cyanobacteria produce bioflocculants that protect them from toxic effects of heavy metals (Bender et al. 1994). Bioflocculants are extracellular macromolecules capable of clarifying turbid water and may remove dissolved nutrients and heavy metals (Sharma et al. 2008) from water column thereby accelerating the heterotrophic activity in the sediment region. Bioflocculants are characterized by presence of numerous negatively charged binding sites that empowered the cyanobacteria in removal of heavy metals from polluted sites. Bar-Or and Shilo (1987) reported the production of bioflocculants by Phormidium sp. strain J-1 characterized by sulfated polysaccharides with attached fatty acids and proteins. The uptake of metal ions (Cu, Pb, Zn, Ni, Cd, and Cr) in Spirulina platensis was accompanied by liberation of protons suggesting an ion-exchange mechanism.

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Removal of mercury and lead by Spirulina platensis has been reported by Slotton et al. (1989). Cyanobacteria have been considered advantageous over traditional approaches which require large spaces and produce large amount of wet bulky sludge. In addition to this, cyanobacteria can be used in the immobilized form. Reports on immobilization of cyanobacteria in silica gel matrix and their potential application in removal of heavy metals have also been reported. Gardea-Torresdey et al. (1998) have experimentally worked out the potency of cyanobacterium, Synechococcus sp., to uptake copper, lead, cadmium, nickel, and chromium. Cyanobacteria have vast metabolic activities and hence can be able to treat complex and variety of pollutants and produced very high value-added products such as carotenoids, vitamins, biopolymers, and hydrogen during treatment can be recovered from cyanobacteria. They produce valuable substances for industrial application such as in medicine, and photosynthetic bacterial 5-aminolevulinic acid has been commercialized and diagnosed for the treatment of cancer (Abed et al. 2009). In addition to this, genetically engineered cyanobacteria have been devised with the novel genes for the production of a number of biofuels such as bio-diesel, bio-hydrogen, bio-methane, and syngas and, therefore, open new avenues for the generation of biofuels in the economically sustainable manner.

3 Cyanobacteria in Industrial Wastewater Treatment Major sources of toxic contaminants in any environment are effluents which discharge from industries containing a variety of organic and inorganic pollutants and causing serious toxicity in the living beings upon exposure (Goutam et al. 2018; Bharagava et al. 2017a,b; Gautam et al. 2017; Saxena et al. 2016; Saxena and Bharagava 2017, 2015; Mohana et al. 2008). Rapid industrialization, urbanization, intensive farming, and other human activities have resulted in land degradation, environmental pollution, and a decline in the crop productivity in agricultural sector (Trupti et al. 2009). Modern civilization with high-tech lifestyle depends upon the use of a myriad of natural and man-made compounds, many of which are hazardous in nature, causing adverse effects on ecosystems including humans (Divya et al. 2015). Several methods have been designed and developed, but more often, these processes produce secondary pollutants, which again are costing the environment for their disposal. Environmental management is crucial for sustainable growth and development. The use of microorganisms to clean up contaminated environment provides cheap alternative method to the conventional treatment methods. But the choice of easily grown, viable, and effective naturally occurring microorganism to do the cleaning is a major challenge. Bioremediation is an emerging and effective management tool to treat and recover the environment, in an eco-friendly way (Divya et al. 2015). A group of toxic chemicals released by various industries into the wastewater are heavy metals, dyes, pesticides, etc.

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Textile dyes and heavy metals cause irritation to the gastrointestinal tract, permanent eye injury, irritation, redness, cancer, and pain to the skin (Kauffman 1970; Katz and Salem 1993; www.lenntech.com/periodic/elements/co.htm). Since these heavy metals and dyes are known to cause several health problems, which may sometimes even be lethal, hence there is increasing emphasis on decontamination of such wastewaters to standard permissible limits. Decontamination or effective disposal of wastewater has always been challenging with a thrust on cost-effective treatment technologies showing minimal side effects. Conventional wastewater treatment technologies for removal of heavy metals and dyes include chemical precipitation, ion exchange, evaporation, membrane separation, electrodialysis, and chemical reduction. All these methods are ineffective at lower concentration of metal ions. These are cost intensive and also generate a large amount of toxic sludge (Robinson et al. 2001), disposal of which further requires energy and capital. Bioremediation is considered as a potential alternative to physicochemical technologies for removal and recovery of metals and dyes from effluents. Bioremediation involves the use of biological agents such as microbes and plants or their enzymes for the remediation purpose and has certain advantages over conventional methods (Saxena et al. 2019; Bharagava et al. 2017c; Chandra et al. 2015). It is costeffective, is easy to operate, and does not produce chemical sludge, hence eco-friendly. Biosorption is a common technique that has increasingly been investigated for removal of toxicants from wastewaters. It involves the use of nonliving (dead) and nongrowing biomass, which does not require nutrients and is metabolism independent. Biosorption is different from “bioaccumulation,” which is an active process, and removal of heavy metals requires the metabolic activity of a living organism (Davis et al. 2003). Both adsorption and absorption occur in bioaccumulation. A large number of microorganisms like bacteria, fungi, algae, and yeast have also been used for bioremoval of heavy metals, dyes, pesticides, COD, etc. There are reports on the removal of pollutants and contaminant by fungal strains like Rhizopus arrhizus, Mucor miehei, Penicillium chrysogenum, and Ganoderma lucidum (Lewis and Kriff 1988; Matheickal et al. 1991; Fourest et al. 1994) and bacterial strains like Arthrobacter sp. and Pseudomonas aeruginosa (Scott and Palmer 1990; Chang et al. 1997). Though relatively fewer reports are available on cyanobacterial utilization for heavy metal and dye removal, in recent years, some studies have reported on these aspects (Vijayaraghavan et al. 2005; Celekli et al. 2009; Mona et al. 2011). Microbial bioremediation can cost-effectively and expeditiously destroy or immobilize contaminants in a manner that protects human health and the environment (Heitzer and Sayler 1993; Gheewala and Annachatre 1997; Gadd 2000). Research is underway at a number of facilities using exogenous, specialized microbes or genetically engineered microbes to optimize bioremediation (Hassan et al. 2003). Cyanobacterial applications showed enormous potential in wastewater treatment, bioremediation, and detoxification of industrial effluents of chemical, bio-fertilizer, food, feed, and fuel industries. Nostoc sp., Anabaena sp., Oscillatoria sp., Synechococcus sp., Nodularia sp., and Cyanothece sp. are a group of cyanobacterial

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species having high biosorption, biodegradation, and dye decolorization capacity of industrial effluents (Dubey et al. 2011; Mona et al. 2011). Considering the biosorptive potential of cyanobacteria, many studies utilized their properties to derive maximum benefits from the system in an environmental friendly way. Bioremediation is a process in which beneficial microbiological agents, such as cyanobacteria, green algae, yeast, fungi, or bacteria, are used to clean up contaminated soil and water. It is defined as the elimination, attenuation, or transformation of polluted or contaminated substances by the application of biological processes. Many microbial species like bacteria (Pseudomonas, Flavobacterium, Alcaligenes sp., Acinetobacter sp., Zooglea sp. and Enterobacteriaceae), fungi (Trichoderma, Aspergillus), yeast, and their consortia also have potential applications for the bioremediation of heavy metals, municipal sewage water, paper and pulp mill effluents, textile effluents, tannery effluents, sugar mill effluents, COD and BOD removal, etc. (Kappesser et al. 1989; Pala and Sponza. 1996; Kamika and Mumba 2012; Sangitha et al. 2012; Buvaneswari et al. 2013; Mahmood et al. 2013; Krishnaveni et al. 2013). Contaminated water destroys aquatic life and reduces their reproductive ability of aquatic organisms, and water becomes unsuitable for human consumption or domestic usage. Cyanobacteria are gram-negative, planktonic, and non-motile organisms. They are capable of consuming CO2, a greenhouse gas which contributes to global warming and released O2 which is vital for cellular respiration (Badger and Price 2003). Their ability to use inorganic acid and phosphorus makes them grow well in wastewater. This indicates their potential ability in wastewater treatment. One of the major problems facing cyanobacterial wastewater treatment is the separation of the cyanobacteria from the wastewater. The use of immobilized technique does not only solve this problem but also has shown to have higher pollutant uptake than using free biomass. In the present section, species of cyanobacteria were reviewed for their tolerance to metals, dyes, crude oil, and pesticides. Cyanobacterial species with very good tolerance to the toxicants were selected as a model system for sequestration of contaminants, pollutants, and toxicants in biosorption studies (Mona et al. 2011).

3.1

Role of Cyanobacteria in Bioremoval of Metals and Dyes from Wastewaters

Bioremoval of pollutants from simulated or trade effluents can be due to bioaccumulation, biosorption, or biodegradation. Though several microorganisms have been studied for various purposes like bioremediation, however, cyanobacteria have been found to be of special advantages over other microorganisms: • Cyanobacteria have simple growth requirements and many of them can fix N2. • Their growth does not require energy-rich compounds. • These are environment-friendly and generally nontoxic microorganisms.

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• Due to their relatively larger size, separation is much easier than other microbial biomass. • Biomass production is larger. • Production of extracellular polysaccharides helps in binding of metal ions. Cyanobacteria can be used in live and immobilized form for the removal of dyes and heavy metals. Various techniques are available for biomass immobilization like adsorption on surfaces, flocculation, cross-linking of cells, encapsulation in a polymer gel, covalent bonding to carriers, and entrapment in a polymeric matrix. Among the different technologies available for the treatment of wastewater contaminated with heavy metals, bioremediation has proved to be cost-effective and eco-friendly, which involves the use of biological materials for the treatment. Different microorganisms like bacteria, fungi, yeast, and algae have been studied for their dye/metal removal capabilities because of their continuous growing nature. Both live and dead cells of alga are used for heavy metal and dye removal as biosorbent (Aksu 2002; Hashim and Chu 2001; Wong et al. 2000; Donmez et al. 1999; Sandau et al. 1966; Crist et al. 1981; Tam et al. 1998; Aravindhan et al. 2007; Robinson et al. 2001; Aksu 2005; Ozer et al. 2005). Due to easy availability, high surface area, and high binding affinity, algae are considered as suitable biosorbents (Darnall et al. 1986; Smith et al. 1994, Robinson et al. 2001; Aksu 2005). Ahuja et al. (1999) reported the copper adsorption capacity of a freshwater cyanobacterium Oscillatoria angustissima. Metal-removing capability has also been reported for the green microalgae, Chlorella vulgaris (Harris and Ramelow 1990; Aksu et al. 1992; Cho et al. 1994). Metal binding is, however, found to vary in different algal species (Wilde and Benmann 1993). Mohamed (2001) demonstrated the ability of both live and dry cells of Gloeothece magna, a nontoxic freshwater cyanobacterium, for the adsorption of cadmium and manganese. Corder and Reeves (1994) reported the accumulation of nickel by live Cylindrica, Anabaena flos-aquae, and Nostoc sp. The toxicity and uptake of iron, zinc, and copper have been studied for Oscillatoria perornata var. altenate and Scenedesmus quadricauda var. longispina by Mittal et al. (1992). Marine algae have also been found to adsorb heavy metals. Matheickal and Yu (1999) presented the uptake properties of two biosorbents (DP95Ca and ER95 Ca) that were developed by chemically modifying the marine algae Durvillaea potatorum and Ecklonia radiata for the lead and copper. There are certain other reports for feasibility of using cheaply available marine or freshwater algae for heavy metal removal (Darnall et al. 1986; Holan et al. 1993). The binding of single metals individually to the green alga, Chlorella vulgaris, and their dependence on pH, mass, and time have been studied by Darnall et al. (1986) and Kubiak et al. (1989). Removal of Cr6+ by Halimeda tuna, Sargassum vulgare, Pterocladia capillacea, Hypnea musciformis, and Laurencia papillosa has been studied with varying pH, initial metal ion concentration, and agitation time, and maximum sorption capacities of 2.3, 3.3, 6.6, 4.7, and 5.3 mg/g were observed, respectively (Baran et al. 2005). Gupta et al. (2001) have studied the biosorption of chromium VI from aqueous solution by a filamentous green algae Spirogyra sp. and found the maximum removal of 14.7 X 103 mg metal/kg of dry weight biomass at a pH of 2.0.

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But free cells are not suitable for column packing. They can provide valuable information in laboratory experimentation only (Ross and Townsley 1986). Free cells have low mechanical strength and small particle size. Cell immobilization can avoid the biomass-liquid separation requirement. Cell immobilization involves the different forms of cell attachment or entrapment (Lopez et al. 1997). Several techniques are used for immobilization like adsorption on surfaces, flocculation, cross-linking of cells, encapsulation in a polymer gel, and entrapment in a polymeric matrix. Entrapment inside a polymeric matrix, made as beads with optimum rigidity, mechanical strength, and porosity, is the most popular technique (Lu and Wilkins, 1995; Annadurai et al. 2000). Both natural polymers (alginate, agar, carrageenan, polysaccharides) and synthetic polymers are in use as gel-forming agents (Lozinsky and Plieva 1998). Use of silica gel for cell entrapment is called as sol-gel technique (Weller 2000). Rangsayatorn et al. (2004) studied the biosorption of cadmium by alginate and silica gel immobilized cells of Spirulina platensis TISTR 8217 with max biosorption capacities of 70.9 and 36.6 mg Cd/g biomass. There are several other reports on use of polymers such as polyacrylamide (Macaskie et al. 1987; Michel 1986; Sakaguchi and Nakajima 1991; Wong and Kwok 1992), polysulfone (Jeffers 1991 Bai and Abraham 2003), and calcium alginate (Babu 1993; Costa and Leite 1991; Peng and Koon 1993; Bayramoglu et al. 2002). Abu Al-Rub et al. (2004) have investigated the technical feasibility of using blank alginate beads and free and immobilized cells for nickel removal from aqueous solution and also found the effect of pH, shaking time, and initial metal ion concentration on adsorption capacity. Removal of copper by alginate beads of algae has been studied by Tam et al. (1998). Fungal biomass can also be used for metal removal. Kiff and Little (1986) utilized the immobilized fungal biomass for biosorption of heavy metals. Dong (2004) has compared the biosorption capacity of alginate beads and both live and inactivated spores of Cladosporium sp. for copper II from aqueous solution. The order for maximum biosorption capacity was found as: Immobilized live biomass > immobilized inactivated biomass > blank calcium alginate beads. All the biosorbents were regenerated by acid treatment and were reused in three biosorption and desorption cycles. Green algae Microcystis in gel has also been found to remove copper and cobalt ions by Jhang et al. (1991). Macaskie and Dean (1986) have used polyacrylamide gel for immobilization of Citrobacter sp., which then help in removal of large quantities of cadmium. Darnall et al. (1986) found that the biosorption of uranyl ions by Chlorella immobilized in polyacrylamide gel was independent of changes in pH in the range of 4–9. There are certain other reports indicating increase in removal of copper, zinc, cobalt, manganese, and mercury by algae when it is in an immobilized form (Singh et al. 1989; Wilkinson et al. 1990). Biodegradation of synthetic textile dyes by algae had been reported in a few studies. The biodegradation pathway was thought to involve reductive cleavage of the azo linkage followed by further degradation of the formed aromatic amines. Several species of green microalgae were capable of decolorizing azo and triphenylmethane dyes to simpler organic compounds or CO2 (Jinqi and Houtian 1992;

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Khataee et al. 2011). The potential of Spirogyra sp. freshwater green alga had been investigated as a biosorbent for biodegradation of reactive textile dye from its multicomponent textile effluent. Comparatively additional dye removal was observed from autoclaved algal biomass relatively gamma irradiated. The effect of process parameters, viz., pH, temperature, biomass concentration, and time, on dye removal was studied. There was 85% removal of dye from dried autoclaved biomass of Spirogyra sp. Khelifi et al. (2009) reported that the nonviable biomass acquired high steadiness and efficiency for dye removal. The capability of Chlorella pyrenoidosa, Chlorella vulgaris, and Oscillatoria tenuis to decolorize 30 different dyes was investigated by Jinqi and Houtian (1992). Fourteen of these azo dyes reported decolorization up to 50% or higher, with results indicating that degradation was directly related to the molecular structure of the azo compounds. Experiments showed high dye decolorization with high uptake yield ranging from 50% to 88% at all dye concentrations by Phormidium sp. (Ertugrul et al. 2008). Another study showed that Phormidium sp. has high uptake yields (13–97%) than Synechococcus at all tested dye concentrations (Sadettin and Donmez 2007). In open wastewater treatment systems, especially in stabilization ponds, algae may therefore contribute to the removal of synthetic dyes and aromatic amines from the water phase. While diffusion of dye molecules from aqueous phase onto the solid biopolymer layer of live microbial cells is reported to be quite efficient (Ozer et al. 2006), dry biomass of algae is reported to be more useful due to greater stability and efficiency (Khalaf 2008). Wastewaters usually contain more than one metal and dye. But most of the studies related to metal and dye revolve around single component system. The number of metals and dyes present in wastewater is expected to cause interactive effects on biosorption. These may be the concentration of metals/dyes, the numbers of metals/dyes competing for binding sites, and the nature and dose of biosorbent (Chong and Volesky 1995; Kandah et al. 2003; Aksu et al. 2009). Abu Al-Rub et al. (2006) have studied the biosorption of copper on Chlorella vulgaris from single, binary, and ternary metal solutions. They found that the presence of lead, zinc, or both suppressed the removal of copper ions. The immobilized biomass of Oscillatoria angustissima has been studied for the biosorption of zinc II, copper II, and cobalt II from single, binary, and ternary metal solutions, as a function of pH and metal concentration through response surface methodology by Mohapatra and Gupta (2005). Sag et al. (2001) evaluated the three metal biosorption equilibria using a fungal biomass. Biomass of marine alga Durvillaea potatorum has been studied for binary adsorption of copper II and cadmium II by Yu and Kaewasam (1999). Ceribasi and Yetis (2001) have also studied the biosorption of Ni(II) and Pb(II) by Phanerochaete chrysosporium from a binary metal system. Aksu et al. (2009) studied the effect of single and binary chromium(VI) and remazol black B biosorption properties of Phormidium sp. Biosorption of reactive dyes on the green alga Chlorella vulgaris was tested at different optimizing parameters like contact time, pH, temperature, and initial dye concentration by Aksu and

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Tezer (2005). Microbial consortium which could degrade azo dye was obtained from a wastewater treatment plant in Taiwan and immobilized in phosphorylated PVA gel beads (Chen et al. 2003). A very few studies are available on response surface methodology (RSM) for optimization of various physicochemical parameters by algae. Neural network modeling of biotreatment of triphenylmethane dye solution by a green macroalgae by Khataee et al. (2011) is very important study for checking the bisorption potential of algae.

3.2

Degradation of Pesticides

In effort to enhance food production to meet the growing human population, the use of pesticide in agriculture is on increased nowadays. The use of this pesticide released toxic substances that are hazardous to man. Pesticides can also bioaccumulate or biomagnify and cause danger to the environment (Ragini and Bisen 2011). Several cyanobacterial species were revealed to be capable of reducing the toxicity of this hazardous substance. Lindane, a highly chlorinated aliphatic pesticide, was found to be degraded by two filamentous cyanobacteria Anabaena sp. PCC 7120 and Nostoc ellipsosporum strain B1453–7 (Kuritz and Wolk 1995). The same species was also found by Kuritz (1998) to degrade the same pollutant, first to pentachlorocyclohexane and then to trichlorobenzenes. Because the degradation occurred in the presence of nitrate and the process is inhibited by both darkness and ammonia, the mechanism of degradation was proposed to be a nitrate reduction system of cyanobacteria. Fenamiphos, an organophosphorus pesticide, was also found to be degraded by five different cyanobacterial species (Ca0 ceres et al. 2008). These species are Anabaena sp., Nostoc muscorum, Nostoc sp. MMI, Nostoc sp. MM2, and Nostoc sp. MM3. The degradation of this pollutant was found to be through hydrolysis and oxidation. The products of oxidation were toxic, while hydrolysis was referred to be detoxification. Pesticide degradation process involves three phases. Phase I involves hydrolysis, oxidation, or reduction. This phase converts the pesticides into less toxic and watersoluble substances. Phase II involves the binding of pesticide metabolites or pesticides to the sugar, glutathione, or amino acids. Toxicity is reduced, and solubility is increased in this phase compared to phase I. Phase III involves the conversion of metabolites of phase II to form nontoxic substances (Hatzios 1991; Shimabukuros 1985). Because of the complexity and toxicity of pesticides, organisms must possess certain features to be able to degrade them to nontoxic substances. According to Ragini and Bisen (2011), organisms can possess the ability to degrade pesticide via adaptive enzymes that can be induced in the presence of pesticides. The constitutive enzymes may be induced by the adapted organisms in the polluted environment before degrading pesticide via random mutation.

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Degradation of Crude Oil

Another serious cause of environmental pollution that makes the soil and water bodies unsuitable for plant growth and difficulty for the survival of aquatic organisms is crude oil. Furthermore, due to health-related problems associated with oil spill, studies are directed toward finding a suitable solution to this problem. Bioremediation offers a better solution to treat oil spills due to its cost-effective nature. Cyanobacteria are suitable candidates in bioremediation of crude oil due to their ability of growing under various conditions and forming bloom with other organisms. Naphthalene, a main constituent of water-soluble part of crude oil, was reported to have been degraded by the cyanobacteria Oscillatoria sp. strain JCM (Narro et al. 1992). The cyanobacteria species Oscillatoria salina, Aphanocapsa sp., and Plectonema terenbans were also found to degrade crude oil (Raghukumar et al. 2001). Within 10 days, 50–65% of pure hexadecane, 45–55% of total crude oil fractions, and 20–90% of aromatic compound were removed. After forming a cyanobacterial mat by the mixed culture of the three cyanobacterial species, about 40% of the crude oil was removed. Phormidium corium and Microcoleus chthonoplastes species of cyanobacteria also degraded crude oil by consuming the individual n-alkanes (Al Hasan et al. 1994). When the growth was measured in terms of biliprotein content and dry biomass, the two species grew better in crude oil suggesting the utilization of these hydrocarbons. Contrary to the above studies, other investigations revealed that cyanobacteria are not directly involved in the degradation of petroleum compound but rather play a major role by enhancing the activity and growth of the actual degraders. For example, Abed and Köster (2005) concluded that the aerobic heterotrophic bacteria associated with the cyanobacteria were responsible for the degradation of oil compounds and not the cyanobacteria. Oil-degrading consortium was also analyzed using 16S rRNA (Sa’nchez et al. 2005). The organisms obtained were found to be related to agrobacterium and rhizobium nitrogen-fixing bacteria. The result indicates that the heterotrophic bacteria associated with cyanobacteria are the ones responsible for degradation of the hydrocarbon, while the cyanobacteria provide oxygen, organic matter, and habitat for the heterotrophic bacteria. The same conclusion was made by Radwan et al. (2002) and Sa’nchez et al. (2005). While the actual degraders degrade the hydrocarbon, the cyanobacteria provide fixed nitrogen and simple organic compound and oxygen for the breakdown of aromatic rings (Abed and Köster 2005) which are also necessary for the growth and the activity of the actual degraders. The evidences for degrading and not degrading hydrocarbon by cyanobacteria are tentative; however, Radwan and Al- Hasan (2002) concluded that some species/strains are capable of utilizing aliphatic and aromatic compounds, while others that occur together with other natural hydrocarbon-degrading bacteria forming cyanobacterial mat only enhance the growth and activities of the degrading bacteria. The association between the two appeared to be significant and crucial in remediation of crude oil.

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4 Cyanobacteria as Sources of Value-Added Products Cyanobacteria are known to produce various fine chemicals, and there are considerable interests in the production of these chemicals from cyanobacteria on a commercially viable scale. Two important cyanobacterial pigments, phycobiliproteins and carotenoids, are extensively used in bio-industry and have high commercial value. On the other hand, the major carotenoids accumulated by cyanobacteria are betacarotene, zeaxanthin, nostoxanthin, echinenone, and canthaxanthin. These pigments are commonly used as food colorants, food additives, and supplements for human and animal feeds. Carotenoids are well known for their antioxidant properties, and their possible role in the prevention and control of human health and disease conditions, for example, cancer, cardiovascular problems, cataracts, and muscular dystrophy, has been reported (Guedes et al. 2011). Some marine cyanobacteria are valuable sources of vitamins, and they are being used for the large-scale production of vitamins of commercial interest such as vitamins B and E. For example, Spirulina is known to be a rich source of vitamin B12, beta-carotene, thiamine, and riboflavin. Cyanobacteria are found to secrete a broad spectrum of enzymes that can be exploited for commercial applications. These industrially important enzymes include protease, amylase, and phosphatases. Cyanobacteria have gained a lot of attention in recent time due to their potential applications in bioremediation and production of value-added product. They have been considered as good source of biological products such as:

4.1

Biopolymers (Polyhydroxyalkanoate)

In conditions of excess essential nutrients, cyanobacteria usually assimilate and store nutrients for future consumption. Various storage materials have been identified in microorganisms, which include glycogen, sulfur, polyamino acids, polyphosphate, and lipid. Polyhydroxyalkanoate (PHA) is a lipoidic material accumulated by cyanobacteria in the presence of abundant carbon sources. The assimilated carbon sources are biochemically processed into hydroxyl alkanoate monomer units, polymerized, and then stored in the form of water-insoluble inclusions (or granules) in the cell cytoplasm. PHA is a type of biodegradable polymer that can serve as substitute for chemical plastics and a biocompatible material that has promising applications in biomedical or pharmaceutical field. Several cyanobacteria including Aphanothece sp. (Capon et al. 1983), Oscillatorialimosa (Stal et al. 1990), some species of the genus Spirulina (Vincenzini et al. 1990), and the thermophilic strain Synechococcus sp. MA19 (Miyake et al. 1996) are natural producers of PHA.

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Biofuels

Cyanobacteria are the unique group of photosynthetic bio-agents that can grow at a fast rate due to their simple cell structure and minimum requirement of nutrients accompanied by the capacity to produce bio-energy including bio-hydrogen (Mona et al., 2011) and bioethanol (Rawat et al. 2011). Cyanobacteria have been used to produce hydrogen gas that constitutes an alternative future energy source to the limited fossil fuel resources (Dutta et al. 2005). Cyanobacteria produce hydrogen either as a by-product of nitrogen fixation, when nitrogenase-containing heterocystous cyanobacteria are grown under nitrogen-limiting conditions, or by the reversible activity of hydrogenase enzymes. Many genera of cyanobacteria such as Anabaena, Oscillatoria, Cyanothece, Synechococcus, Nostoc, Microcystis, Calothrix, Gloeobacter, Aphanocapsa, Chroococcidiopsis, and Microcoleus produce hydrogen gas under various culture conditions. The production of ethanol via biological route has received widespread attention in recent years. Traditionally, bioethanol is produced in a two-step route to first collect plant-derived biomass, and subsequent conversion of the biomass to fuels by microbial fermentation is employed (Stephanopoulos 2007). This two-step bioethanol production is inefficient in comparison than photosynthetic microbes who directly convert carbon dioxide to fuels. Although some cyanobacterial strains naturally produce low quality ethanol but, it could be enhance the production efficiency of ethanol production by genetic engineering. An attempt to introduce the pyruvate decarboxylase (pdc) and alcohol dehydrogenase II (adh) genes from Zymomonas mobilis into the chromosome of Synechocystis sp. PCC 6803 was reported (Deng and Coleman 1999). Further engineering of Synechocystis sp. by overexpressing endogenous alcohol dehydrogenase and disrupting polyhydroxyalkanoate biosynthetic pathway increased ethanol production up to 5500 mg/L (Gao et al. 2012).

4.3

Bio-fertilizer

Cyanobacteria can fix atmospheric N2 and are good source of bio-fertilizer. They can fix atmospheric nitrogen in the form of free-living or in the symbiotic associations with partners such as water fern Azolla, cycads, Gunnera, etc. Some cyanobacteria have specialized cells known as heterocyst which are considered site of nitrogen fixation by nitrogenase enzyme. Nitrogenase enzyme is a complex in nature and catalyzes the conversion of the molecular N2 into reduced form like ammonia (Singh et al. 2011). The fixed nitrogen may be released in the form of ammonia, polypeptides, and free amino acids, either by secretion or by microbial degradation after the cell death. Nitrogen-fixing ability is not limited by heterocyst-containing cyanobacteria but also by several non-heterocystous unicellular and filamentous genera which can also fix nitrogen. Cyanobacteria as bio-fertilizer is advantageous over the chemical fertilizer as it is cheap and eco-friendly (Issa et al. 2014).

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Food Supplements

Cyanobacteria as food supplements are available in the market in different forms such as tablets, capsules, and liquid. They enhance the nutritive value of foods and beverages (Liang et al. 2004). They can act as the nutritional supplement or represent a source of natural food colorants (Soletto et al. 2005). Spirulina is the most commercial cyanobacterial strain used for human nutrition because of its highprotein content and excellent nutritive value (Soletto et al. 2005). It contains more than 60% proteins and is rich in beta-carotene, thiamine, and riboflavin and is considered to be one of the richest sources of vitamin B12 (Prasanna et al. 2010). Spirulina contains a wide spectrum of prophylactic and therapeutic nutrients including B-complex vitamins, minerals, proteins, γ-linolenic acid, and super antioxidants such as β-carotene, vitamin E, trace elements, and a number of unexplored bioactive compounds (Kulshreshtha et al. 2008).

4.5

Bio-control Agents

Cyanobacteria produce a variety of biologically active compounds of antibacterial, antifungal, antialgal, and antiviral potential (Dahms et al. 2006). These bioactive compounds belong to the group of polyketides, amides, alkaloids, fatty acids, indoles, and lipopeptides (Burja et al. 2001). In addition cyanobacteria produce a broad spectrum of antialgal compounds which inhibit growth of pathogens by disturbing their metabolic and physiological activities (Dahms et al. 2006). The excretion of bioactive compounds by cyanobacteria into the aquatic environments is a possible allelopathy strategy used by cyanobacteria to outcompete other microorganisms within the same ecosystem (Gross 2006). To date, several allelochemicals from cyanobacteria have been identified, and they include cyanobacteria produced by Scytonema hofmanni, enediyne-containing photosystem II inhibitor synthesized by Fischerella muscicola, hapalindoles that are isolated from Hapalosiphon and Fischerella sp., and nostocyclamides from Nostoc sp.

4.6

Emulsifiers

Some halophilic cyanobacteria produce large amounts of exopolysaccharides (EPS), which are used for oil recovery by decreasing its surface tension and increasing its solubility and mobility. EPS, when gelated under alkaline conditions, was employed to remove dyes from textile effluent. The halophilic cyanobacterium Aphanocapsa halophytica was used for the production of EPS (Matsunaga et al. 1996).

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5 Conclusion and Future Prospective Cyanobacteria are excellent accumulators or degraders of various environmental contaminants such as dye, heavy metals, pesticides, and oil-containing compounds. They are fast-growing and ubiquitous bio-agents, which can also be used for capturing and storing of CO2 that may also lead to climate change mitigations through photosynthesis and biological calcification. They are also the ideal source of variety of bioactive compounds with marked antagonistic properties. The cyanobacteria are multifunctional bio-agents for safe and eco-friendly environmental sustainability, along with several other uses. To improve their utility in bioremediation and associated sectors needs more attention. Thus there is an urgent need to address certain key issues of exploiting cyanobacteria in the better way. Since the use of cyanobacteria to produce valuable chemicals including food supplements is still little explored, there seems a long way to go. In addition to product developments, future research must address the strain improvement of useful cyanobacteria to achieve high-quality food and fuel products and maintain high growth rates and survival under harsh environmental conditions. The utility of cyanobacteria in sustainable agriculture and environment can be enhanced by genetic manipulations. However, the application of genetic engineering to improve bio-fuel production in cyanobacteria is still in its infancy. In future, genetic and metabolic engineering of cyanobacteria are likely to play important roles in improving the economics of cyanobacteria-mediated bio-fuel production. However, from lab to field condition shift will not be as easy as it will require more efforts and strategies according to the field condition. Additionally, new methods need to be developed to allow the cultivation of previously uncultivable strains. The methods should consider the organism’s requirements in the field, and these conditions should be mimicked in the laboratory. Numerous compounds that have antibacterial and antialgal activity have been extracted from different species of cyanobacteria; however, only limited compounds have been screened for their antifouling, antiviral, and antitumor activities. These compounds are often present in small quantities, and in order to obtain sufficient amounts, it is needed to harvest and extract many marine organisms. Most of the commercial compounds were isolated from freshwater cyanobacteria. Marine environments with different environmental conditions ranging from shallow euphotic zone to deep-sea hydrothermal vents are likely to be a good source for a variety of cyanobacterial species that may have high biotechnological and environmental significance.

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Shashirekha S, Uma L, Subramanian G (1997) Phenol degradation by the marine cyanobacterium Phormidiumvalderianum BDU-30501. J Ind Microbiol Biotechnol 19:130–113 Shimabukuro RH (1985) Detoxification of herbicides. In: Duke SO (ed) Weed physiology, vol 2. CRC Press, Boca Raton Singh DP, Khattar JIS, Kaur M, Kaur G, Gupta M, Singh Y (2013) Anilofos tolerance and its mineralization by the cyanobacterium Synechocystis sp. strain PUPCCC 64. PLoS One 8 (1):53445 Singh DP, Khattar JIS, Nadda J (2011) Chlorpyrifos degradation by the cyanobacterium Synechocystis sp. strain PUPCCC 64. Environ Sci Pollut Res 18(8):1351–1359 Singh SP, Verma SK, Singh RK, Pandey PK (1989) Copper uptake by free and immobilized cyanobacteria. FEMS Microbiol Lett 11:193–196 Slotton DG, Goldman CR, Frank A (1989) Commercially grown Spirulina found to contain low levels of mercury and lead. Nutr Rep Int 40:1165–1172 Smith LA, Alleman BC, Copley-Graves L (1994) Biological treatment options. In: Means JL, Hinchee RE (eds) Emerging technology for bioremediation of metals. Lewis Publishers, London, pp 1–12 Sokhoh NA, Al-Hasan RH, Radwan SS, Hopner T (1992) Self-cleaning of the Gulf. Nature 359:109 Soletto D, Binaghi L, Lodi A, Carvalho JCM, Converti A (2005) Batch and fed-batch cultivations of Spirulina platensis using ammonium sulphate and urea as nitrogen sources. Aquaculture 243:217–224 Stal LJ, Heyer H, Jacobs G (1990) Occurrence and role of poly-hydroxy-alkanoate in the cyanobacterium Oscillatorialimosa in Novel Biodegradable Microbial Polymers. Springer, Dordrecht, pp 435–438 Stanier RY, Kumizawa R, Mandel M, Cohen-Bazire G (1971) Purification and properties of unicellular blue-green algae (order Chroococcales). Bact Rev 35:171–205 Stephanopoulos G (2007) Challenges in engineering microbes for biofuels production. Science 315 (58):801–804 Tam NFY, Wong YS, Simpson CG (1998) Repeated removal of copper by alginate beads and the enhancement by microalgae. Biotechnol Tech 12:187–190 Trupti D, Eapen CS, Fulekar MH (2009) Characterization of industrial waste and identification of potential micro- organism degrading tributyl phosphate. J Toxicol Environ Health Sci 1 (1):001–007 Verma SK, Singh SP (1995) Multiple metal resistance in the cyanobacterium Nostoc muscorum. Bull Environ Contam 54:614–619 Vijayaraghavan K, Jegan J, Palanivelu K, Velan M (2005) Biosorption of copper, cobalt and nickel by marine green alga Ulva reticulata in a packed column. Chemosphere 60:419–426 Vincenzini M, Sili C, de Philippis R, Ena A, Materassi R (1990) Occurrence of poly-β-hydroxybutyrate in Spirulina sp. J Bacteriol 172(5):2791–2792 Weller MG (2000) Immunochromatographic techniques-a critical review. Fres J Anal Chem 366:635–645 Whitton BA (1969) The taxonomy of blue-green algae. Br Phycol J 4(l):121–123 Wide EW, Bennan JR (1993) Bioremoval of heavy metals by the use of microalgae. Biotechnol Adv 11:781–812 Wilde EW, Benmann JR (1993) Bioremoval of heavy metals by the use of microalgae. Biotechnol Adv 11:781–812 Wilkinson SC, Goulding KH, Robinson PK (1990) Mercury removal by immobilized algae in a batch culture system. J Appl Phycol 2:223–229 Wong PK, Kwok SC (1992) Accumulation of nickel ion by immobilized cells of Enterobacter species. Biotechnol Lett 14(7):629–634 Wong PK, Wong YS, Tam NFY (2000) Nickel biosorption by two Chlorella species, C. vulgaris (a commercial species) and C. miniata (a local isolate). Bioresour Technol 73:133–137 Yu Q, Kaewsarn P (1999) Binary adsorption of copper (II) and cadmium (II) from aqueous solutions by biomass of marine alga Durvillaea potatorum. Sep Sci Technol 34:1595–1605

Chapter 17

Plant-Microbe Interactions for Bioremediation and Phytoremediation of Environmental Pollutants and Agroecosystem Development Akash Mishra, Shraddha Priyadarshini Mishra, Anfal Arshi, Ankur Agarwal, and Sanjai Kumar Dwivedi

Abstract Development in both the industrial and agricultural sectors has resulted in excess production of hazardous substances which is ruining our environment. However several physicochemical technologies are available to treat such substances but require extra setup to deal with eco-friendly manner. Phytoremediation and bioremediation has emerged as a substitute of such technologies which is brought by the interaction among plant and microorganisms. PGPR (plant growth-promoting rhizobacteria) has an important contribution in remediation of environmental pollutants as well as agro-ecosystem development. Along with PGPR, several fungi, endophytes, mycorrhiza, and algae also form association with plants and contribute in sustainable development. Application of genetic engineering has resulted tremendous effect in increasing their efficiency of pollution control and plant growth regulation. Keywords Environmental pollutants · Bioremediation · Phytoremediation · Plantmicrobe interactions

1 Introduction In the present era, our world is suffering through various economic and environmental problems, among which conventional energy depletion, global warming, and water pollution are of more concern. These problems are affecting the whole society in different ways. As per a report by WHO 2013, drinking water pollution is a major problem of half of the population worldwide. Such pollution is responsible for A. Mishra · S. P. Mishra · A. Arshi (*) · A. Agarwal · S. K. Dwivedi Defence Institute of Bio-Energy Research (DIBER), Defence Research and Development Organization (DRDO), Ministry of Defence, Government of India, Haldwani, Uttarakhand, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_17

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around 250 million cases of waterborne disease and 0.005–0.01 billion deaths in a year. A vigorous development in agricultural as well as industrial sector has resulted in excess production of chemicals and its entrance into the environment as toxic contaminants (Sharma et al. 2014). Due to extreme presence of these potential toxicants, clean water and healthy soil have become scarce resulting in limited crop production (Kamaludeen et al. 2003). There are a number of toxic agents which pose serious hazardous effect to our environment resulting in water, air, and soil pollution (Goutam et al. 2018; Bharagava et al. 2017a; b; Goutam et al. 2017; Saxena et al. 2016; Saxena and Bharagava 2015). The remediation of such pollution in water and soil often involves some technologies that are expensive, cost-effective, and labor-intensive and require site restoration either with physical or chemical methods. Due to drawbacks of these technologies, scientists have started to develop some new technology as an alternative to using plant and microorganism or both in an interaction for the removal of toxic contaminants in soil (Glick 2003). Use of certain plants for removal or destruction of hazardous toxicants from environment for its cleanup is the recently developed method and termed as phytoremediation. The plants used in this method are called hyperaccumulators which grow best in metal concentration-rich soil (Glick 2003). Alkorta and Garbisu (2001) have reported phytoremediation to be an effective, nonintrusive, in situ, aesthetically pleasing, low-cost, and socially accepted technology for the remediation of polluted soil. This technology may remediate the pollutants in several forms: phytostabilization, rhizofiltration, phytoextraction, and phytovolatilization. Bioremediation is the process which uses microorganisms like bacteria or fungi and yeast for the cleaning of polluted water and soil (Bharagava et al. 2017c; Saxena and Bharagava 2017; Kishor et al. 2018). In this technology, the growth of indigenous microbial consortia of polluted site is promoted for desired activity (Agarwal 1998) by controlling biotic and abiotic stresses. Not only remediation of environmental pollutions but plant-microbe interaction contributes to sustainable development of agriculture also. Nowadays, there is a big challenge in crop production with reduced use of pesticides and chemical fertilizers. Therefore the use of PGPR for increase in crop yield has proved environmentally friendly approach as an alternative to such problems. The direct and indirect mechanism of plant growth promotion by these PGPR includes nutrient regulation and hormonal regulation in plants resulting in induced resistance against phytopathogens. The current chapter is based on the elastration of available physicochemical technologies for environmental remediation and plant microbial interaction with special reference to phytoremediation and bioremediation for the sustainable development of agroecosystem.

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2 Environmental Pollutants and Their Toxicity in Environment Pollutants in the environment are of several categories like organic, inorganic, and radioactive and some other metals. Inorganic pollutants are mainly nitrate, sodium, arsenic, or ammonia, whereas metallic pollutants are characterized by cadmium, copper, mercury, chromium, and selenium. Uranium, strontium, and cesium are the main radioactive substances causing pollution in the environment. Organic pollutants are the main source of environmental pollution. It includes various compounds like bentazon and atrazine as pesticides; polycyclic aromatic hydrocarbons (PAHs); petroleum hydrocarbons such as toluene and benzene; and trichloroethylene which is a chlorinated solvent. There are some other very hazardous pollutants being released into the environment unintentionally or intentionally and posing global concern for their remediation. These lipophilic chemicals are called as persistent organic pollutants (POPs) because they get accumulated in different biological systems present in the environment like animal tissue and are resistant to photochemical and biodegradation resulting in longtime presence in the environment (Buccini 2003; Wong et al. 2005; Sharma et al. 2014). Recent advancement in day to day life of humans has led to the increase in the utilization of nanoparticles in cosmetics, but in most of the cases, besides their benefits, the negative effects observed on the environment requires mineralization or removal of these toxic chemicals (Landis and Yu 2003). According to Kuppusamy et al. (2016), there is a list of toxic pollutants mainly inorganic and organic (as demonstrated in Fig. 17.1.) which exert risk to health of more than 100 million people if exposure occurs. The exposure to these toxic pollutants may have adverse health impacts like organ dysfunction, cancer, mental and physical disorders, neurological disorder, and reduced immune system and ultimately causes death (Godduhn and Duffy 2003; Perera and Herbstman 2011; Mates et al. 2010; Yu et al. 2011; Huang et al. 2012). Since the adverse effects of various kinds of environmental pollutants, the demand has increased to develop a suitable technology for lowering the cost of pollutant treatment because the remedial sector plays an important role in strengthening the GDP. In this respect, the selection of already available technology for better applicability directly depends upon the characteristics of polluted site and task objectives. Therefore, a combination of all the adopted methods for remediation of pollutants such as biological and physicochemical means is the most promising option in present scenario (Kuppusamy et al. 2016).

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Fig. 17.1 Major organic and inorganic pollutants (Adapted from Kuppusamy et al. 2016)

Major pollutants

Inorganic

Organic

Aldrin chlordane Dieldrin Endrin Heptachlor Diioxins Furans

Heavy Metals

Non-Metals

Cadmium Lead Antimony Arsenic Mercury Nickel Selenium Silver Thallium Beryllium Zinc Copper Chromium

Nitrate Ammonia Sulphate Phosphate Cyanide

3 Environmental Remediation Technologies With the increase in human population, industries based on food production, health stability, automobiles, etc. have also expanded which results in more natural resources utilization like water, land, and air (Kumar et al. 2011). Various kinds of environmental pollutants are being used for this purpose and are having adverse effect on the environment. Hence, cleanup of the environment is necessary. Several biological and physicochemical remediation technologies (Fig. 17.2) can be adapted to cure the environment. These technologies are categorized as ex situ or in situ on the basis of site of treatment. The transport and treatment of polluted media (soil, water) from contaminated site to a different location is called ex situ, whereas on-site treatment of pollutants is called in situ mode of remediation. Both ex situ and in situ methods of remediation have some advantages and disadvantages in their uses. The major advantage with the use of in situ is it does not require excavation and transport of contaminated soil from its site and also the cost of the treatment process and risk of exposure to pollutant are

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Remediation Technologies

Physicochemical

Incineration Advanced oxidation processes(AOPs) Dehalogenation Chemical Reduction/Oxidation Solidification Chemival extraction

Biological

Bioremediation

Phytoremediation

Biosparging Bioventing Bioaugmentation Land farming Composting Biopiles Bioreactors

Phytoextraction Phytotransformation Phytostabilization Rhizodegradation Rhizofiltration Phytovolatilization

Fig. 17.2 Physicochemical and biological technologies for environmental remediation

minimal. The main disadvantage of this technique is less efficiency in pollutant removal than ex situ. Generally, the cost of ex situ treatment process is very high, but the time requirement for this process makes it more applicable than in situ method. The soil treated by ex situ method can be further used for landscape purposes (Kuppusamy et al. 2016).

3.1

Physicochemical Remediation

Pollutant removal from water and soil can be achieved by several physical and chemical means which are as follows:

3.1.1

Incineration

It is the process which involves disposal of hazardous waste through exposing them to a very high temperature (750–1200
C). Burning can be achieved in different types of experimental setup such as infrared combustors (infrared energy as heat source), fluidized bed combustors, circulating bed combustors, and rotary kilns,

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where temperature ranges in between 850 and 1010
C depending upon the type of incineration chamber used (FRTR 2012). Prime benefits of incineration are reduced bulky solids or wastes and the less amounts of greenhouse gases (CH4 and CO2) generation. Recovery of energy and aid into the economy is the other advantage in use of the incinerator. The main drawback of this technology is that it is very expensive both in construction and operation of this facility (Kuppusamy et al. 2016).

3.1.2

Advanced Oxidation Processes (AOPs)

This technology is basically based on the use of ozone along with UV or hydrogen peroxide and on the other hand UV with hydrogen peroxide. High cost of reagent (energy source, ozone, hydrogen peroxide) used in this process is the main disadvantage of the technology.

3.1.3

Dehalogenation

The technology is also called as dechlorination. In this process, halogen molecule like chlorine is replaced by hydrogen or a reducing radical containing a hydrogen donor for decomposition of contaminants in organic compounds. There are two dehalogenation processes: 1. Base catalyzed decomposition (BCD) – Where screened contaminated soil is crushed and mixed with sodium bicarbonate followed by its introduction in reactor for heating of mixture above 330
C (630
F) and volatilization or partial decomposition of pollutants. 2. Alkaline polyethylene glycol process – In this APEG process, polyethylene glycol (an alkaline reagent) is used to form glycol ether and/or a hydroxylated compound. An alkali metal salt also forms as by-products which are watersoluble. 3.1.4

Chemical Reduction/Oxidation

Chemical reduction/oxidation or redox reactions are the conversion of hazardous contaminants (viz., metals and inorganic, pesticides, cyanides, triazines, and formaldehyde-contaminated soils) to non-hazardous or less toxic compounds which are less mobile and so more stable. In this reaction, electrons are transferred from one compound to another, where the first compound losing electron is oxidized and the other one gaining the electron is reduced. The most commonly used oxidizing agents are hypochlorite, chlorine, chlorine dioxide ozone, hydrogen peroxide, etc. To make the process more effective, mixture

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of the reagents can be used combining them with ultraviolet oxidation. In the reduction processes of unsaturated organic contaminants or high oxidation state metals like Cr (VI), metals with low oxidation potential or sodium borohydride are generally used. The chemistry involved in this method is generally well known, and it has been used for years in related chemical processes. But the main drawback of this method is the requirement of excessive amount of reagents making it costly for high contaminant concentrations, and partial decontamination may result along with formation of intermediate contaminants.

3.1.5

Solidification

This is the method of stabilizing the contaminant by physical bound or enclosing within a low permeability mass, i.e., solidification. The mobility is reduced by the induction of chemical reaction between contaminant and stabilizing agent. This technique can be applied ex situ as well as in situ but requires additional setups. Inorganics, including radionuclides, are mostly treated by this method, whereas it has less effectiveness against organics and pesticides. The main disadvantage in the application of this technique is generation of higher final mass of pollutants than the original contaminated soil, and contaminants are neither eliminated nor transformed into less toxic form, and only mobility is reduced.

3.1.6

Chemical Extraction

In this process, contaminants are separated from the soil to reduce the volume of contaminant. On the basis of the type of contaminants, two major chemical extraction processes are as follows: 1. Acid extraction – Acids are used to extract contaminants from soils. Additionally after decontamination, residual acid is neutralized by dewatering of soil followed by mixing it with fertilizer and lime. 2. Solvent extraction – To remove mixtures of metals and organic compounds, different solvents are used in the treatment of soil. Physical separation is generally required prior to chemical extraction, which can enhance the process by separating out particulate heavy metals. An advantage of this technology is that it can be used for the extraction of a range of selected organic contaminants for the treatment such as SVOCs, VOCs, some fuels, explosives, and inorganics, heavy metals, etc. However, the effectiveness of this technology is limited on organics with high molecular weight (eugris.info).

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Biological Remediation

Biological remediation is mainly of two types, i.e., bioremediation and phytoremediation.

3.2.1

Bioremediation

Bioremediation is the process for removal of environmental contaminants with the use of biological agents mainly microorganisms (Saxena and Bharagava 2016; Bharagava et al. 2019). Therefore, it is one of the best management tools for remediation and recovery of contaminated environment. Most importantly, for the success of various bioremediation technologies, the nature of contaminated site and complexity of organisms being used must be strategized prior to treatment process. Here, a list (Table 17.1) is being presented consisting of various microorganisms like fungi, anaerobes, and aerobes which have been used in environmental remediation. These microorganisms used in bioremediation may be of indigenous nature to polluted site, or they may be isolated from elsewhere and introduced to the site Table 17.1 Microbial agents reported in the degradation and detoxification of environmental pollutants Microorganism Organic pollutants Leifsonia Scenedesmus obliquus, Euglena gracilis Chlamydomonas sp.

Toxic chemicals

References Anhalt et al. (2007) Ardal (2014)

Chlamydomonas sp. Chlorococcum sp. Dunaliella sp. Heavy metals Bacillus cereus strain XMCr-6 Kocuria flava Bacillus cereus

Imidacloprid DDT, parathion Lindane, naphthalene, phenol Toxaphene, methoxychlor Mirex

Cr (VI) Cu Cr (VI)

Sporosarcina ginsengisoli Pseudomonas veronii Aspergillus versicolor Aspergillus fumigatus Spirogyra spp. and Spirulina spp.

As (III) Cd, Zn, Cu Ni, Cu Pb Cr Cu, Fe, Mn, Zn

Hydrodictyon, Oedogonium, and Rhizoclonium spp.

As, V

Dong et al. (2013) Achal et al. (2011) Kanmani et al. (2012) Achal et al. (2012) Vullo et al. (2008) Tastan et al. (2010) Kumar et al. (2011) Mane and Bhosle (2012) Saunders et al. (2012)

Chlorella sp.

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(Vidali 2001). Now, scientists from all over the world have started to put their energy to select or search new organism with more biodegradation ability for a number of pollutants from different environmental locations (Kumar et al. 2011). On the basis of applicability, bioremediation also can be categorized ex situ and in situ depending upon the experimental process involved. In situ Biodegradation It is a type of bioremediation in which nutrients and oxygen are supplied into the contaminated site in the form of aqueous solution and degradation of organic contaminants is stimulated by native bacteria. This process is best applicable in case of polluted groundwater and soil. (A) Biosparging In this process, concentration of groundwater oxygen is increased by injecting pressurized air in the water, and contaminants are degraded biologically by native microorganisms. The injected air increases the contact between groundwater and soil so that saturated zone gets mixed. The requirement of less capital input in construction of the air injection system makes this process more flexible. (B) Bioventing It is the most commonly used in situ method which involves air supply with less flow rate than biosparging. Here, the nutrients and necessary oxygen are provided to indigenous bacteria through wells to stimulate biodegradation and minimize the chance of release of volatile contaminants into the environment. This process is best used for treatment of contaminants deep below the surface (C) Bioaugmentation It is the addition of potential microorganisms to the contaminated site with better degradation ability. The organism may be indigenous or exogenous. Ex situ Bioremediation This technique involves the physical removal or excavation of polluted soil from a location. It involves: (D) Land Farming It is a very simple process which involves excavation of contaminated soil and spreading over a bed followed by periodical turning for complete degradation of contaminants by indigenous microorganisms through aerobic degradation. Advantage with this process is its reduced monitoring and maintenance cost, whereas limitation to treat superficial 10–35 cm of soil is the main drawback of this process. (E) Composting A rich microbial population for biodegradation of pollutants with a characteristic temperature of compost can be achieved by the mixing of contaminated soil with nonhazardous organic contents like agricultural wastes or manure. These organic contents support the growth or survivability of microorganisms which degrade the pollutants.

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(F) Biopiles A technique consisting property of both composting and land farming where cells are constructed to aerate the composted pile. Here, physical loss such as volatilization and leaching of contaminant is reduced. (G) Bioreactors In this ex situ technology of removal of pollutants, water and contaminated soil are treated in aqueous reactors or slurry reactors. Here, polluted materials are more manageable than in situ methods. The main disadvantage of this process is that it requires pretreatment like washing of soil.

3.2.1.1

Advantage of Bioremediation

It is the most natural and publically accepted process to treat pollutants. The residue produced after the practices is harmless which may include bacterial cells, water, and carbon dioxide. It is less expensive and possesses almost complete degradation of pollutants.

3.2.1.2

Disadvantage of Bioremediation

The main disadvantage of this technology is the limitation of treatment of only biodegradable materials. Sometimes the bacterial metabolic process involved in the process is highly specific; hence a controlled environment is required for the successful degradation of contaminants. The longer time consumption and pretreatment of target media like soil contribute to the disadvantages of the process.

3.2.2

Phytoremediation

Plants are natural filter and metabolize naturally generated substances, and therefore, the use of plants for the removal of contaminants in water and soil is the emerging technology known as phytoremediation (EPA 1999, 2000; Raskin and Ensley 2000; Chandra et al. 2015; Saxena et al. 2019). This technology is further categorized in six different types of techniques (Table 17.2) which are classified on the basis of types of contaminants: phytovolatilization, phytotransformation, phytostabilization, rhizofiltration, rhizodegradation, and phytoextraction.

3.2.2.1

Phytoextraction or Phytoaccumulation

In this process, accumulation of contaminants in plant takes place in the root system and shoot or leaves present above the ground which ultimately saves economy

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Table 17.2 An overview of different phytoremediation strategies for environmental decontamination Phytoremediation techniques Phytoextraction

Phytotransformation Phytostabilization Rhizodegradation

Rhizofiltration Phytovolatilization

Mechanism Uptake and concentration of metal via direct uptake into the plant tissue with subsequent removal of the plants Plant uptake and degradation of organic compounds Root exudates cause metal to precipitate and become less available Enhances pollutant degradation in rhizosphere Uptake of metals into plant roots Plants evaportranspirate metals such as selenium, mercury, and volatile hydrocarbons

Surface medium Contaminated soils and wastewaters Surface water and groundwater Contaminated soils, groundwater, mine tailing waste Remediation of contaminated soils and groundwater within rhizosphere Surface water Contaminated soils and groundwater

Adapted from Vidali (2001)

invested in various costly remediation technologies. Contaminant like metals present even in low level can be removed from the site and accumulated in plants and further recovered by recycling from the biomass before its disposal.

3.2.2.2

Phytotransformation or Phytodegradation

It is the process in which highly toxic organic contaminants from a polluted site like water body or soil can be taken up and transformed to less toxic forms via plant system.

3.2.2.3

Phytostabilization

This process is the reduction of mobility of a contaminant and its migration into the ground level. Pollution-causing substances are leached out and absorbed to the surface of plant roots making it stable and avoiding its reentrance into the environment.

3.2.2.4

Rhizodegradation

A mutual relationship between plant root and microorganisms like bacteria and fungi residing in rhizosphere makes an environment where contaminants are broken down through the metabolic activity and secretion of enzymes and proteins from the root system of plant.

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Rhizofiltration

With this remediation technology, contaminants present in the water can be filter out by the help of plant roots.

3.2.2.6

Phytovolatilization

This process is the uptake and conversion of contaminant into the gaseous state through plant system and release into the environment. Volatile organic compounds are the best candidate for this process to be treated by the help of evapotranspiration.

3.2.2.7

Disadvantages of Phytoremediation

There are several disadvantages associated with this technology. The foremost disadvantage is time consumption in treatment of a polluted site, whereas growth of plant used is also inhibited by the increase in metal level inside the plant body. Bioavailability of metal or any other contaminant is one of the major contributors in the disadvantages of this technology.

4 Plant-Microbe Interaction for Sustainable Agricultural Development and Environmental Cleanup The interaction between plant and microbes has a very significant role in the development of agriculture as well as remediation of environment. Microbial interaction with plants may be both negative and positive, resulting in disease development or stimulation in growth of plant along with stress tolerance by the help of beneficial microbiota (Abhilash et al. 2012). In addition to it, the communication system to form an interaction between plant and microorganism also helps in resource distribution in below or above the ground across plant body and provides resistance against its competitors. This communication system modifies the physicochemical property of soil and diversity of biotic life which helps in plant growth promotion and pollutant removal from soil (Fig. 17.3). The secondary metabolites produced in form of exudates by the plant root and shoot system are responsible for the development of such communication system for the plant-microbe interaction. However, understanding the exact mechanisms of interaction is more or less difficult and complex as it takes place at different spheres of the plant system such as endosphere, phyllosphere, and rhizosphere. Therefore it is vital to understand the exact mechanisms of interaction between plant and microbes for the assessment of contribution of plant beneficial microbiota in sustainable agriculture, environmental cleanup, and restoration of ecosystem (Saleem and Moe 2014; Dubey et al. 2015).

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Phytofiltration

Phytovolatilization

Blastofiltration

Phytoevaporation Phytostimulation

Phytohydraulics Phytoaccumulation

Phytodegradation

Phytoextraction Phytoassimilation

Bioremediation

Phytoconcentration Phytotransfer

Phytotransformation

Phytomining

Phytoremediation

Phytocontainment

Phytoreduction

Phytoimmobilization Phytostabilization

Phytooxidation

Phytosequestration

Rhizosphere Nitrilase

Desorption Adsorption

Respiration

Nitroreductases

Cytochromes P450

Precipitation

Acidification Redox reactions

Hydrolases Oxidases

Chelation

Glucosyltransferases

Exudation

Peroxidases

Complexation

Rhizoremediation

Transportation Bacterial ACC deaminase

Leakage

427

Laccase Fungi

Bacteria

Fig. 17.3 Plant-microbe interaction and strategies to act against pollutants (Adapted from Ma et al. 2011)

Nowadays, sustainability of agriculture without a polluted environment is of major concern worldwide (Singh et al. 2011), and therefore the beneficial impact of plant-microbe associations can be a best alternative for this problem. In respect to this, several bioagents have gathered the attention of researchers for their use in biofertilizer and some other valuable effects such as healthy crop promotion and sustainable development of agroecosystem by alteration in physical, chemical, and biological factors involved in establishment of better interaction between plant and soil (Barea et al. 2005). For this purpose, studies are going on, and some species of bioagents have been commercialized like Bacillus, Azospirillum, Pseudomonas, Azotobacter, Klebsiella, Enterobacter, Variovorax, and Serratia sp. (Glick 2012). On the other hand, microbes have role in environmental bioremediation as it is a natural process which is carried out by association of plants and microorganisms.

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Therefore, bioremediation is a cost-effective tool for destruction of contaminants with the help of biological activity of microbes (Kamaludeen et al. 2003). This activity can be enhanced by the supplementation of nutrients (P and N) and other substrates like phenol, methane, and toluene (Baldwin et al. 2008; Akhtar et al. 2013). According to Weyens et al. (2009), plant-microbe interaction has a crucial role in phytoremediation by plant growth promotion and sequestration of pollutants, its detoxification, and degradation. These microbes have certain metabolic abilities and degradation pathways which results in degradation of organic pollutants and evapotranspiration of volatile organic contaminants in more effective way (Weyens et al. 2009). Some root endophytes are equipped by metal sequestration/resistance and can enhance the accumulation of these toxic metals in plant tissue even if they are present as trace element in soil (Rajkumar et al. 2012). Therefore, it is expected that microbes present in phyllosphere of the plant can resist the stress due to particulate matter contamination and promote the phytoremediation ability of plant. Growing indoor plants can increase humidity level in the air, and allelochemicals released into the environment through them can inhibit airborne harmful microbes (Berg et al. 2014; Wolverton 2008).

4.1

Role of Plant Growth-Promoting Rhizobacteria

Plant growth-promoting rhizobacteria have a very significant role in growth promotion of plants resulting in sustainable agricultural development. They aid in growth promotion of plant by two mechanisms, viz., direct and indirect mechanism. Fixation of atmospheric nitrogen, solubilization of phosphorus, synthesis of siderophore for iron chelation, and supplying siderophore-iron complex to plant so that plant may synthesize various phytohormones like gibberellins, cytokinins, and auxins come under direct mechanism, whereas indirect mechanism is brought about by control of disease-causing phytopathogens by producing antibiotics, depletion of iron in the soil, and ultimately stimulation of plant growth. On the basis of interaction with host plant, PGPR are categorized into two groups: (1) symbiotic rhizobacteria, which invade and infest the interior of the plant cell to survive (known as intracellular PGPR, e.g., bacteria forming nodule), and (2) free-living rhizobacteria that reside outside the plant and are also known as extracellular PGPR, e.g., Azotobacter, Pseudomonas, Burkholderia, and Bacillus (Babalola and Akindolire 2011; Khan 2005). Microorganisms like plant growth-promoting rhizobacteria can enhance the nutrients availability in rhizosphere (Choudhary et al. 2011). For example, nitrogen is the most limiting factor in plant growth as it is not easily available for plant, but Azospirillum present in cereal ecosystem can fix the free nitrogen and improve the crop yield (Tejera et al. 2005). Additionally, phosphate is also solubilized by PGPR (Wani et al. 2007), which is further taken up by plants. Vejan et al. (2016) decribe that Lavakush et al. (2014) conducted a study on PGPR strain like Pseudomonas putida and Pseudomonas fluorescens for their effect on nutrient uptake in rice.

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Role of Endophytes

According to Schulz and Boyle (2006) and Lodewyckx et al. (2002), bacteria that colonize the intimate niche of plant (internal tissues) without any negative effects or infection to host are called endophytic bacteria. Except seed endophytes, the primary site to gain entry (or route of colonization) by endophytes into plants is via the roots which is now confirmed by several microscopic studies (Pan et al. 1997; Germaine et al. 2004). After getting entrance into the plant, endophyte resides in xylem or root cortex or transports through the vascular system to colonize the plant systematically (Mahaffee et al. 1997; Quadt-Hallmann et al. 1997). With interaction to plants, endophytes get carbohydrates, and in return, they provide resistance to plant from various abiotic and biotic stresses (Hamilton and Bauerle 2012; Hamilton et al. 2012). Endophytes can alter the structural community of plant (Clay and Holah 1999; Yuan et al. 2011), as well as they regulate the interaction between competitors and their host plant (Omacini et al. 2001; Clay and Holah 1999; Hyde and Soytong 2008; Guo et al. 2008) A study of Chen et al. (2010), Shin et al. (2011), and Luo et al. (2011) describes that many endophytes are being used in phytoremediation because of metal resistance or organic pollutant degradation and plant growth promotion ability.

4.3

Role of Mycorrhiza

For plant growth development, mycorrhizae fulfil its role by providing mineral nutrients exclusively the uptake of phosphate to the plants (Moose 1972). According to Bagyaraj (1984), Entry et al. (2002), and Fomina et al. (2005), this effect is because of several key features of mycorrhiza such as (i) extra radical mycelium increases the absorbing surface and exploits large soil volume; (ii) hyphal diameter is small which leads to increase in P-absorbing surface area; (iii) P concentration is low in mycorrhiza by the formation of polyphosphates (poly P); and (iv) release of P is catalyzed by the production of phosphatases and organic acids. Hence mycorrhiza can help in sustainable development of agroecosystem by increasing plant survival rate and plant nutrients acquisition, and also it helps in increasing carbon and nitrogen deposition into the soil and reduces plant stress (Almas et al. 2004). The infection by mycorrhiza helps in the increase of uptake of Pb and Mn by plant from soil solutions even in low concentrations (Heggo et al. 1990; Malcova and Gryndler 2003), and thus, they may play crucial role in the phytoremediation of contaminated site (Liao et al. 2003; Gohre and Paszkowski 2006; Orlowska et al. 2011; Zarei et al. 2010; Chanda et al. 2014). The association formed by ectomycorrhiza can perform a significant resistance against metallic toxicity in contaminated soil (Leyval et al. 1997) and petroleum-like compounds (Sarand et al. 1999) or polycyclic aromatic hydrocarbons (Leyval and Binet 1998).

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Mycorrhiza enhance the growth of plant by improving nutrition, resistance, and tolerance against various stresses (Clark and Zeto 2000; Turnau and Haselwandter 2002). They can also be used as bioprotectants, biodegraders, and biofertilizers (Xavier and Boyetchko 2002). Several studies also report their phytoremediation potential for heavy metal-polluted soil (Chaudhry et al. 1998; Khan et al. 2000; Khan 2001; Jamal et al. 2002; Hayes et al. 2003; Khan and Ahmad 2006).

5 Genetically Engineered Microorganisms in Environmental Remediation Genetic engineering technique which is also called as recombinant DNA technology is based on the natural genetic interchange in between microorganisms, and the organism formed is called genetically engineered microorganism (GEM) or modified microorganism (GMM). These engineered microorganisms have the capacity to bioremediate the soil, activated sludge, and groundwater by degrading varied chemical. Several researchers also suggest that the genetically modified organisms may have more potential to remediate the environment than wild ones. Several gene complexes or plasmids are responsible for degradation of various environmental pollutants and generally for every compound, and one separate plasmid is required. According to Ramos et al. (1994), for better understanding, four categories are described: 1. 2. 3. 4.

OCT plasmid for degradation of hexane, decane, and octane. Camphor is decomposed by CAM plasmid. XYL plasmid can degrade xylene and toluenes. NAH plasmid for naphthalene degradation.

The best GMM, for example, Pseudomonas putida, contains the NAH and XYL plasmid and a hybrid of CAM and OCT plasmid which can degrade camphor, salicylate, octane, and naphthalene. It can metabolize hydrocarbons more effectively and grow quickly on crude oil (Markandey and Rajvaidya 2004). This organism formed by the technology of genetic engineering is known as superbug (oil-eating bug). According to Huang et al. (2004), there may be three recommended criteria for gene recombination and selection as a suitable strain: (1) after cloning, stability and expression of target gene should be confirmed for selected strains; (2) the strain should be contaminant tolerant or insensitive; and (3) strains should survive in plant rhizosphere. Many rhizobacteria have only limited capability in degrading organic pollutants. With the use of advance molecular biology, rhizoremediation may get achieved by the construction of genetically engineered rhizobacteria with the contaminantdegrading gene (Glick 2010).

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6 Future Prospects and Challenges This chapter describes and presents lucrative information about available technologies especially for environmental bioremediation and agro-ecosystem development. However, information about organisms involved and the interaction formed with indigenous microbiota are less. In addition to it, study based on characterization of all the other microorganisms as well as increasing efficiency of these organisms with the help of new recombinant DNA technology is required. The matter of bioavailability of pollutants and longer time period required in the process like bioremediation and phytoremediation is of more concern.

7 Conclusion The present chapter has mainly focused on the plant-microbe interaction in maintaining ecosystem sustainability along with remediation of environmental pollution. Accordingly, we first summarize key information on available remediation technologies. Then, we have discussed microorganism responsible for remediation of environmental pollutants and their interaction with plant to maintain sustainable ecosystem. Further, we have discussed on how to discover and manipulate the efficiency of these organisms with the help of genetic engineering. This chapter has a special emphasis on biological remediation technology and its superiority over physicochemical methods to treat environmental pollutants. Finally, we also outline the research required in the future. Acknowledgment Authors are extremely grateful to Director, Defence Institute of Bio-Energy Research (DIBER), DRDO, Ministry of Defence, Government of India (GOI), India, for providing financial and infrastructural support.

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Chapter 18

Molecular Technologies for Assessment of Bioremediation and Characterization of Microbial Communities at PollutantContaminated Sites Sudhir Kumar Shekhar, Jai Godheja, and Dinesh Raj Modi

Abstract Among the various microbial biodegradation techniques, molecular microbiology methods have revolutionized microbial biotechnology, thus leading to rapid and high-throughput methods for culture-independent assessment and exploitation of microbes present in polluted environments. Whether organic or inorganic, pollutants present in contaminated sites can cause an imbalance in the ecosystem by affecting the flora and fauna. The efficiency of naturally occurring microorganisms for field bioremediation could be significantly improved by the microbial molecular biology approach for its comparatively high efficiency and safety. Many techniques, including polymerase chain reaction (PCR), fluorescent in situ hybridization (FISH), denaturing gradient gel electrophoresis (DGGE), ribosomal intergenic spacer analysis (RISA), amplified ribosomal DNA restriction analysis (ARDRA), terminal-restriction fragment length polymorphism (TRFLP), single-strand conformation polymorphism (SSCP), and ribosomal intergenic spacer analysis (RISA) can be selectively employed in microbial flora and ecology research. Recent methods such as genotypic profiling, metagenomics, ultrafast genome pyrosequencing, metatranscriptomics, metaproteomics, and metabolomics have provided exemplary knowledge about microbial communities and their role in the bioremediation of environmental pollutants. Only 1% of the microbial diversity can be cultured by traditional techniques. Thus, the application of molecular techniques in studying microbial populations in polluted sites without the need for culturing has led to the discovery of novel and previously unrecognized microorganisms. Such complex microbial diversity and dynamics in contaminated soil offer a clear opportunity for bioremediation

S. K. Shekhar · D. R. Modi (*) Department of Biotechnology (DBT), School for Bioscience and Biotechnology (SBBT), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India J. Godheja School of Life and Allied Sciences, ITM University, Naya Raipur, Chhattisgarh, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_18

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strategies. These techniques not only prove the existence of microbes in polluted sites but also reveal the undetectable complex relationships among them. This book chapter presents an overview of the different applications of molecular methods in bioremediation of hydrocarbons and other pollutants in environmental matrices and an outline of recent advances in the applications of such techniques. Keywords Bioremediation · Metagenomics · Metaproteomics · Metabolomics · Pyrosequencing · Biodegradation · PCR · DGGE · DNA hybridization

1 Introduction The tremendous increases in industrialization and in the extraction of natural resources have created extreme environmental contamination and pollution. Many toxic compounds have been dispersed in numerous contaminated sites. Varied evidence shows higher risks to human health: cocktails of pollutants in nature are causing a global epidemic of cancer and other degenerative diseases. These types of pollutants are mainly classified as inorganic or organic. To make environments safe for human habitation and food consumption, developing innovative and economically low-cost solutions for decontaminating polluted environments is a challenging task. In developed countries such as the UK, USA, Canada, Australia, Japan, and European countries, much progress has been made as compared to India. In India, evaluation of the developments in laboratories is urgently needed. New molecular techniques provide a novel opportunity to proceed with the required microbial culturing and have greatly increased the available methods for delineating bacterial diversity and functionality during the bioremediation of hazardous industrial waste. Currently, only a fraction of the potent microorganisms involved in biodegradation can be cultured by using standard laboratory agars and different pollutants from various ecosystems (Chikere 2000; Malik et al. 2008). Comparative study between molecular/metagenomic and culture-dependent methods states that only about 1% of all microorganisms are amenable to culture (Mallik et al. 2008). In many cases, the concept of phylogenetic diversity has been key for the development of more effective microbial culture methods by the improvement of media, extended incubation period, and various growth factors such as temperature, pH, and atmospheric conditions (Rajendhran and Gunasekaran 2011). However, it is the subject of great efforts to evaluate the total microbial diversity/dynamics in various environments, a persisting challenge especially during bioremediation of industrial hazardous waste (Chikere and Ughala 2011; Chikere et al. 2011a, b). For evaluation of microbial communities, research experiments rely on new rapid methods for characterization of cellular constituents such as proteins (enzymes), nucleic acids, and specific compounds (Maila 2005; Maila et al. 2005, 2006; Montiel et al. 2009). These molecules can be extracted directly from the polluted soil and used to elucidate microbial community composition during bioremediation (Simon and Daniel 2011). This book chapter presents recent techniques

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Table 18.1 Advantages and disadvantages of bioremediation Advantages of bioremediation Natural process acceptable to the public as an efficient waste treatment process for contaminated sites Residual components are usually harmless products such as water, simpler hydrocarbons, carbon dioxide, and cell biomass Useful for complete detoxification and biodegradation of many hazardous contaminants Complete destruction of target pollutants is possible More economic than other physicochemical technologies used for hazardous waste cleanup

Disadvantages of bioremediation Limited to biodegradable compounds Residual components may be more toxic than the parent toxic compound These processes need specific site factors for complete success Many techniques are difficult to perform in large-scale field operations Longer treatment time than other, physicochemical options

for molecular microbiology applications in the assessment of microbial diversity and dynamics in polluted environments to identify the predominant microbial communities or derivative genes for bioremediation.

2 A General Overview of Bioremediation Bioremediation offers one of the best ways to destroy toxics or render them harmless using natural biological activity. Bioremediation is important for two reasons: 1. Chemicals are not used: Use of natural microbes and no chemicals is the most important advantage of bioremediation, because the use of chemicals in the treatment and removal of toxics can lead to further contamination of the environment. Initially, many chemicals were used for remediation but slowly their side effects to the environment were observed, which opened the way for use of microbes in remediation. 2. Option of waste recycling: Bioremediation is also preferred in that once the waste is neutralized or removed, it can be recycled, whereas in a chemical remediation some amount of the remaining waste shows incomplete neutralization and thus cannot enter the recycling process. Some of the advantages and disadvantages of bioremediation are given in Table 18.1.

3 Overview of Environmental Pollutants and Their Toxicity 3.1

Organic Pollutants

The use of synthetic chemicals has increased tremendously because of demand, their easy manufacturing, and their applications in various industries. These chemicals

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include pesticides, plastics, paints, hydrocarbon fuels, soaps, detergents, and many more substances. Each chemical has a different effect on the environment and its organisms. Dichlorodiphenyltrichloroethane (DDT) affects the endocrine system of birds, leading to eggshell thinning and reduced breeding. Other chlorohydrocarbons are also highly toxic to aquatic organisms, being persistent, soluble in fat, and having very low solubility in water; thus, they quickly accumulate in the food chain, a process described as bioaccumulation. Many countries have banned the use of these chemicals and have promoted the use of other chemicals such as organophosphates, carbamates, pyrethrins, and pyrethroids. Having relatively less solubility in water and fat, these chemicals do not bioaccumulate and thus are less toxic, having a much lower impact on the environment. However, these chemicals can be highly toxic to biota including mammals and aquatic organisms. Soaps and detergents, containing surfactants, have relatively low toxicity, persistence, and bioaccumulation. Although discharged to the environment in large quantities their effects are generally low. Plastics, on the other hand, enter the environment as solid wastes with little toxicity but have other adverse environmental effects. Proper management of pollutants through bioremediation is the demand of the present time. Some important pollutants, their sources, and their effects are highlighted in Table 18.2.

3.1.1

Hydrocarbons

Hydrocarbon contamination in the environment poses a serious problem whether it comes from petroleum products, pesticides, or other toxic organic matter. Of the hydrocarbons, petroleum products are a great concern because these are toxic to all forms of life. Environmental contamination by crude oil is relatively common because of its widespread use and the associated disposal operations and accidental spills.

3.1.2

Polychlorinated Biphenyls (PCBs)

Another group of organic pollutants, polychlorinated biphenyls (PCBs), shares a common structure but differ in the number of attached chlorine atoms. The international treaty on Persistent Organic Pollutants, drafted by 122 nations in Johannesburg in December 2000, emphasized phase-by-phase removal of targeted PCBs from the world. PCBs can be carcinogenic according to the International Agency for Research on Cancer, the U.S. Environmental Protection Agency, and the National Toxicology Program. According to the National Institute for Occupational Safety and Health, PCBs are a potential occupational carcinogen leading to increased rates of melanoma, liver cancer, gallbladder cancer, biliary tract cancer, gastrointestinal tract cancer, and brain cancer. PCBs are reported to cause various types of cancer in rats, mice, and other study animals.

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Table 18.2 Pollutant sources, toxicity, and health effects Pollutants Hydrocarbons (Saturated alkanes, branched alkanes, alkenes, napthenes, aromatics, napthenoaromatics, resins, asphaltenes, carboxylic acids, ethers, and others released mainly by human activities Polychlorinated biphenyls (PCBs)

Toxicity Humans are exposed to hydrocarbon pollution directly or indirectly

Organs affected Nervous system, immune system, respiratory system, circulatory system, reproductive system, sensory system, endocrine system, liver, kidney, etc.

More than 90% of human exposure to PCBs is through food, mainly meat and dairy products, fish, and shellfish

Agricultural supplements Hazardous pesticides, herbicides, synthetic insecticides Dyes Can be acidic, basic, disperses, azo- or anthraquinonebased and metal complex dyes Arsenic Pesticides; gold, copper, nickel, iron, lead mining; coal burning; wood preservatives; pharmaceutical and glass industries; pigments; poison bait; agrochemicals; antifouling paint, electronics industry Asbestos Mining of raw asbestos around refineries, power plants, shipyards, steel mills, vermiculite mines, building demolition Cadmium Rechargeable batteries, zinc smelting, mine tailings, burning coal or garbage containing cadmium, pigments, televisions, solar panels, phosphate fertilizer, metal plating, sewage sludge

Accumulate in the food chain

Short-term exposure results in skin lesions and altered liver function Long-term exposure results in impairment of the immune system, nervous system, endocrine system, and reproductive functions Liver and kidney functions

Azo dyes enter the body by ingestion; metabolized by intestinal microorganisms causing DNA damage

Skin, nervous system, liver and kidney

Exposure mostly through consumption of groundwater containing high levels of inorganic arsenic, food prepared with this water, or food crops irrigated with this water

Chronic arsenic poisoning (arsenicosis); gastrointestinal tract, skin, heart, liver, and neurological damage, diabetes, bone marrow and blood diseases, cardiovascular diseases

Exposure through release of microscopic asbestos fibers into the air via inhalation; can also be ingested or adsorbed on the skin

Causes parenchymal asbestosis, pleural abnormalities, lung carcinoma, and pleural mesothelioma

Cadmium can enter plant crops, depending on soil characteristics and pH; cadmium can enter animals at levels that are not harmful to them, but can affect humans consuming those animal products

Liver and kidney damage, low bone density

(continued)

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Table 18.2 (continued) Pollutants Lead Mining, batteries, solder, ammunition, pigments, paint, hair colour, fishing equipment, leaded gasoline, plumbing, coal burning, water pipes Mercury Thermometers, electrical switches, fluorescent light bulbs, batteries, dental fillings, mining, pesticides, medical waste, chlor-alkali industry Radionuclides Nuclear weapons program; nuclear weapons testing; nuclear power plants; uranium mining and milling; commercial fuel reprocessing; geological repository of highlevel nuclear wastes; nuclear accidents

3.1.3

Toxicity Accumulation of lead in topsoil from leaded fuel and mining activities

Organs affected Nervous system, hand–eye coordination, encephalopathy, bone deterioration, hypertension, kidney disease

Humans are exposed via eating contaminated seafood; children are exposed via direct ingestion of contaminated soil

Central nervous system and gastric system, brain development, coordination, eyesight and sense of touch, liver, heart, kidney

Exposed to living beings by radiation poisoning

Skin redness and hair loss, radiation burns, acute radia tion syndrome; prolonged exposure can be carcinogenic

Pesticides

Pesticides have been commonly used to control pests during recent decades (Timmons 1970; Chauvel et al. 2012), leading to widespread deposition of these xenobiotics into the environment (Toccalino et al. 2014). The intensive use of pesticides causes harmful effects on biodiversity, food security, and water resources (Malaj et al. 2014; Queyrel et al. 2016). Agricultural producers are using these pesticides to increase food production because the world population is expected to increase by 30%, to 9.2 billion, by 2050, with the further demand to increase food production by 70% (Popp et al. 2013). Nonpesticidal tools have been developed and will be important in the near future, but chemical pesticides are currently the best solution to pest control and food security (Popp et al. 2013; Fisher et al. 2012).

3.1.4

Dyes

Dyes are important in human lives because they impart color to our clothes, are used as food colors, and are even used in our medicines. Scientists have done much research to produce these artificial dyes: more than 10,000 dyes are available commercially and 7 lakh tons of dyes are produced annually (Zolinger 1987).

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Approximately 10% to 15% of dyes are released into the environment during the dyeing process, and many of them are highly colored and aesthetically unpleasant. Dyes are difficult to remove by conventional water treatment procedures, and they persist in water because they are highly soluble. Biodegradation is highly applicable for detoxifying the toxic and carcinogenic components of dyes (Rindle and Troll 1975).

3.2

Inorganic Pollutants

Inorganic pollutants include metals, metal compounds, mineral acids, inorganic salts, trace elements, organic metal complexes, cyanides, and sulfates, which have long-term adverse effects on aquatic flora and fauna as well as on terrestrial organisms. High concentrations of heavy metals and other inorganic pollutants from various industries contaminate the water. These compounds are nonbiodegradable and persist in the environment. Metals at concentrations above the threshold value are toxic to biota; for example, copper is toxic to microbes at concentrations greater than 0.1 mg/l.

3.2.1

Radionuclides

Radioactive pollutants are toxic to all life forms because they accumulate in the bones and teeth, causing serious disorders. Radioactive isotopes such as 131I, 32P, 60 Co, 45Ca, 35S, and 14C originate from mining and processing of ores, research labs, agriculture, medical, and industrial activities, and as radioactive discharge from nuclear power plants and nuclear reactors (90Sr, 137Cs, 248Pu, 238U, 235U) and during testing of nuclear weapons. The safe concentration for lifetime consumption is 1
107 microcuries per milliliter (μCi/ml).

3.2.2

Heavy Metals

Heavy metals are categorized as environmental pollutants because of their toxic effects on plants, animals, and humans. Heavy metals belong to a very heterogeneous group of elements that vary widely in their chemical properties and biological functions. Heavy metal contamination of soil results from both anthropogenic and natural activities. Anthropogenic activities such as mining, smelting operations, and agriculture activity increase the levels of heavy metals such as Cd, Co, Cr, Pd, As, and Ni in soil above the threshold limit. Because heavy metals are persistent in nature, they are easily accumulated in soils and plants, with long-term detrimental effects on human health when ingested as a part of the diet when consuming vegetables and plant foods. The impact of heavy metals on aquatic organisms results from the movements of pollutants from various diffuse or point sources, giving rise

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to coincidental mixtures in the ecosystem. Thus, these metals pose a great threat to aquatic fauna, especially to fish, which constitute one of the major sources of protein-rich food for mankind. Thus, we studied the acute and sublethal toxic effects of heavy metals in Krishna River sediment, water, and aquaculture using atomic absorption spectroscopy. The aquatic environment becomes contaminated with a variety of pollutants generated from diverse sources (industry, agriculture, domestic). Among these pollutants, pesticides, heavy metals, and detergents are the major causes of concern for the aquatic environment because of their toxicity, persistency, and tendency to accumulate in organisms. The 19 elements that constitute the heavy metal group have many similar physical and chemical properties and are remarkably different from the other 97 known elements. Among these 19 heavy metals, lead, cadmium, and mercury are extremely toxic. Other metals, such as chromium, copper, manganese, nickel, tin, and zinc, when once dispersed in the biosphere cannot be recovered or degraded and thus cause permanent environmental damage. Metal pollution has harmful effects on biological systems and does not undergo biodegradation. Toxic heavy metals such as Pb, Co, Cd, and Hg can be differentiated from other pollutants, because they cannot be biodegraded but can be accumulated in living organisms, thus causing various diseases and disorders even at relatively lower concentrations.

4 Molecular Fingerprinting Techniques in Microbial Identification Fingerprinting methods rely on sequence variations in the genome of different organisms. Those differences result in different melting behaviors as well as different restriction enzyme recognition sites. Different species have sites at unique positions along the whole genome. Two commonly used methods to analyze environmental samples of unknown microbial community composition are denaturing or temperature gradient gel electrophoresis (DGGE or TGGE) and restriction fragment length polymorphism (RFLP). Both these methods, which were introduced by Muyzer et al. (1993), make use of sequence variations of polymerase chain reaction (PCR) products amplified from environmental DNA on a gradient of either increasing denaturants or temperature. Both methods rely on differences in the sequencedependent melting behavior of double-stranded DNA. For this, the extracted nucleic acids must be amplified using primers that target specific molecular markers such as the 16S rRNA gene. For DGGE/TGGE, the use of a 50-GC clamped (30–50 nucleotides) forward primer is essential to avoid complete dissociation of the double strands. After loading the PCR product onto the gel, an electric current is applied, pulling the DNA fragments through the gel. Depending on the sequence of the amplicon, a higher or lower denaturant concentration or temperature is needed for a complete dissociation of the double strands into single strands, causing the amplicons to stop moving through the gel at different positions. To determine the

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identities of bands separated on the gel, those bands can be excised from the gel and further analyzed via re-amplification, cloning, and sequencing, or by hybridization with molecular probes specific for particular taxonomic groups. Although being valuable for the identification of changes in dominant species within a community, fingerprints generated from DGGE/TGGE can be very complex, especially when using universal bacterial primers. Those bands in which the fragments have similar melting points can be difficult to separate. Also, the rRNA operons of the same bacterium can show heterogeneity, leading to multiple bands and an overestimation of the microbial diversity; further, quantification of the extracted bands is not possible. Compared to DGGE/TGGE, T-RFLP analysis has the advantage that fragments can be relatively quantified by the intensity of the fluorescent signal, and the method is quicker and less labor intensive. The inherent difficulty with T-RFLP, however, is that a collection or recovery of the fragments and thereby a subsequent analysis and identification of the microorganism via sequencing is not possible. To overcome this, fragments have been identified via comparison against databases of fragments produced by known gene sequences (Kent et al. 2003). Combining T-RFLP with clone library construction and sequencing, Huang et al. (2011) were able to find close associations of the four most dominant operational taxonomic units detected in T-RFLP to phylum or genus level, when analyzing spatial and temporal variations of the microbial community in a tailings basin of a Pb–Zn mine. Different studies have used DGGE/TGGE and T-RFLP to assess microbial community composition in different contaminated environments. Spatial and temporal variations of microbial community composition were analyzed in different mining environments such as an acidic stream draining across a pyrite mine in China (Tan et al. 2009) or in a low-temperature, acidic, pyrite mine, where Kimura et al. (2011) were able to highlight the importance of bacterial species in iron transformation using T-RFLP and fluorescent in situ hybridization (FISH). Kim et al. (2009) used DGGE to examine the effects of mine tailings and waste rocks on the hydrogeochemistry and microbiology of a stream and groundwater near an abandoned copper mine. The effects of metal pollution on microbial community structure and composition in a salt marsh were analyzed by Cordova-Kreylos et al. (2006), who used T-RFLP to aid the development of bioindicators of toxicant-induced stress and bioavailability of contaminants for wetland biota. In another study, Gough and Stahl (2011) used T-RFLP to follow microbial community changes in lake sediments along a metal contamination gradient. In a recent study by Thavamani et al. (2012), the authors employed a holistic approach. To determine the soil microbial activity affected by a mix of polyaromatic hydrocarbons (PAHs) and heavy metals, they combined physicochemical, biological, and advanced molecular methods to analyze the activities of the soil microbial community in long-term mixed contaminated soils collected from a former manufactured gas plant (MGP) site. The study highlighted the difficulties of implementing remediation strategies when studying mixed contaminations, as well as the importance of combining different analysis methods.

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Other studies used those fingerprinting techniques when monitoring the effects of different remediation techniques such as amending mine tailings with a mixture of organic carbon sources to treat pore water and drainage (Lindsay et al. 2011), incorporating compost into a heavy metal-contaminated acidic soil (Farrell et al. 2010), or testing the effects of phytoremediation approaches (Martinez Inigo et al. 2009) or land-farming on oil refinery sludge (Ros et al. 2010). In laboratory-based studies, fingerprinting techniques have been used when testing the effect of different contaminants/elements on microbial communities. Jakobs-Schonwandt et al. (2010) investigated the shift of soil microbial communities when subjected to a biocide frequently found in wood preservatives. Brandt et al. (2010) compared a Cu-adapted and a corresponding nonadapted soil microbial community for the ability to resist experimental Cu pollution. Other studies investigated the abilities of indigenous bacteria on arsenic mobilization (Corsini et al. 2011) or the ability of specialized mixed communities to selectively precipitate transition metals from acidic mine waters (Nancucheo and Johnson 2011), or the acid tolerance response of a bioremediation system based on sulfate reduction (Lu et al. 2011a, b). Fingerprinting techniques are extremely useful when starting to investigate an unknown microbial community; however, as mentioned previously, they have their limitations and can be time and labor extensive. On their own, both techniques can either quantify or identify fragments but not both.

5 Molecular Techniques for In Situ Monitoring of Microbial Communities, Bioremediation Processes, and Environmental Pollution Traditional methods for characterizing microbial communities have been based on analysis of those bacteria that can be cultured. The overall structure of the community has been difficult to interpret as most of the bacteria are not culturable (Dokic et al. 2010). Modern methods focus towards molecular techniques, which do not require culturing the microorganisms but provide measures based on genetic diversity. The molecular-phylogenetic perspective is a reference framework within which microbial diversity is described; the sequences of genes can be used to identify organisms. A variety of approaches (Table 18.3) that have been developed to study molecular microbial diversity include DNA–DNA and mRNA–DNA hybridization, DNA re-association, DNA cloning and sequencing, and other PCR-based methods such as denaturing gradient gel electrophoresis (DGGE), temperature gradient gel electrophoresis (TGGE), and ribosomal intergenic spacer analysis (RISA). Other advanced techniques, such as DNA microarrays, have also improved specificity to a great extent.

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Table 18.3 Advantages and disadvantages of molecular-based methods to study microbial diversity Method Mol% guanine plus cytosine (G + C)

Nucleic acid reassociation and hybridization

qPCR and qRT-PCR

Denaturing- and temperaturegradient gel electrophoresis (DGGE, TGGE)

Single-strand conformation polymorphism (SSCP) Restriction fragment length polymorphism (RFLP) Terminal restriction fragment length polymorphism (T-RFLP)

Ribosomal intergenic spacer analysis (RISA)/amplified ribosomal DNA restriction analysis (ARDRA) DNA microarrays and DNA hybridization

5.1

Advantage Quantitative method, not affected by polymerase chain reaction (PCR) biases; includes all DNA extracted and also includes rare members of microbial community Can be studied in situ, not influenced by PCR biases; can be used to study both DNA and RNA Use of gels eliminated; allows sample to be analyzed in real time Large number of samples can be analyzed simultaneously; reliable, reproducible, rapid

Same as DGGE/TGGE; use of GC clamp and gradient eliminated Detect structural changes in microbial community Simpler banding patterns than RFLP, can be automated, highly reproducible; ability to compare differences between microbial communities Highly reproducible community profiles

Thousands of genes can be analyzed; increased specificity

Disadvantage Requires large quantities of DNA, highly dependent on lysing activities and extraction efficiency; level of resolution is low Sequences need to be in high copy number; lack of sensitivity; depends on lysing and extraction efficiency

Only detects dominant species, PCR biases, dependent on lysing and extraction efficiency; one band can represent more than one species PCR biasing; some ssDNA can form more than one stable conformation PCR biasing, banding patterns often too complex Dependent on extraction and lysing efficiency; type of Taq can increase variability, choice of restriction enzymes will influence community fingerprint Requires large quantities of DNA (for RISA), PCR biases

Only detects the most abundant species; need to culture organisms; only accurate in low-diversity systems

Mole Percentage Guanine + Cytosine (Mol% G + C)

The base composition of DNA was used for taxonomic purposes as mole percentage guanine + cytosine (mol% G + C). Much diversity in base composition occurs within bacteria (25–75%) but certain microorganisms have constant GC values. Mol% G + C can be determined by thermal denaturation of DNA, and it has been reported that closely connected organisms have fairly similar GC profiles whereas

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taxonomically connected groups only differ from 3% to 5% (Tiedje et al. 1999). However, similar base composition is not a confirmation of relationship, although it suggests that the difference can be evidence of a missing relationship.

5.2

Nucleic Acid Hybridization and Reassociation

Nucleic acid hybridization is an important qualitative and quantitative tool in molecular bacterial ecology (Clegg et al. 2000). Valuable spatial distribution information on microbial communities in natural environments is often provided by hybridization methods, mostly using extracted DNA or RNA, or in situ. Probes (oligonucleotide or polynucleotide) can be designed from well-known sequences, ranging in specificity from domain to species, and can be tagged with markers at the 50 -end (Goris et al. 2007). After lysing the sample to release all nucleic acids, dot blot hybridization with specific and universal oligonucleotide primers is used to quantify rRNA sequences of interest relative to total rRNA. The relative abundance may represent changes in the abundance in the population or changes in the activity and hence the amount of rRNA content (Theron and Cloete 2000). Another approch is in situ cellular level hybridization. The only limitation is the lack of sensitivity of hybridization of nucleic acids, which requires the sequences to be present in high copy number, such as those from dominant species; otherwise, the probability of detection is low. The kinetics of DNA reassociation estimates diversity by measuring the genetic complexity of the microbial community (Torsvik et al. 1996). Total DNA is extracted from environmental samples, purified, denatured, and allowed to reanneal. The rate of reassociation depends on the similarity of sequences present, and as the complexity of DNA sequences increases, the reassociation rate decreases (Theron and Cloete 2000). Two parameters controlling the reassociation reaction are the concentration of DNA product (Co) and time of incubation (t), usually described as the half-association value, Cot1/2 (the time needed for half the DNA to reassociate). The value of Cot1/2 can be used as a diversity index in special conditions, as it takes into account both the amount and distribution of DNA reassociation (Torsvik et al. 1998). Thus, the similarity between communities of two different samples can be studied by measuring the degree of similarity of DNA through hybridization kinetics (Griffiths et al. 1999).

5.3

Quantitative PCR and RT-PCR

DNA amplification by PCR has been used in many studies to detect, characterize, and identify pollutant-degrading microbial populations. PCR detection of genes encoding for microbial monooxygenases and dioxygenases such as NAHA, PHNAc, NIDA, and NARB are useful for the detection of pollutant-degrading

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microbial populations (Lu et al. 2011a, b). However, one of the most commonly used approaches for the detection and identification of microorganisms is the PCR amplification of microbial ribosomal RNA (rRNA) genes (e.g., 16S, 18S, 23S rRNA). The rRNA genes are the basis for microbial phylogenetic analyses, as several million sequences have been published in the GenBank database. During bioaugmentation treatments, rRNA of introduced microorganisms can be easily amplified by PCR and detected by gel electrophoresis. In most cases it is necessary to analyze the rRNA amplification products by additional techniques, such as terminal-restriction fragment length polymorphism (T-RFLP), or fully sequencing the amplified product, to increase the specificity of detection and identification. On the opposite end, quantitative PCR (qPCR) or real-time PCR (RT-PCR) has been also used to quantify microorganisms after introduction to different environmental matrices (Kikuchi et al. 2002). One of the most specific and popular ways to perform qPCR is with the use of Taqman probes. In this technology, Taq polymerase cleaves a fluorogenic Taqman probe that binds to an internal site within the sequence being amplified during the extension step, which releases a fluorescent molecule (fluorophore), resulting in fluorescence. The cycle threshold value (Ct) is determined at the point where a significant increase in the fluorescence emission occurs, as compared with a background baseline. A larger initial concentration of target DNA results in a lower Ct value. qPCR eliminates the use of gels and allows the sample to be analyzed in real time, in less time than conventional PCR. qPCR has been used in bioremediation studies to calculate the copy number of the benzyl succinate synthase gene (BSSA) and naphthalene dioxygenase (NAHAc) in hydrocarbon-contaminated soils, bioaugmentated with degrading microbial consortia (da Silva and Alvarez 2004; Nyyssonen et al. 2006). Cebron et al. (2008) reported the use of qPCR to detect and quantify PAH-ring hydroxylating dioxygenases (PAH-RHDα) in soil and sediment samples contaminated with PAHs. The results of these studies highlighted a positive correlation among the PAH-degrading gene copy levels and microbial biodegradation potential, as well as the contamination levels in the studied soils.

5.4

DGGE and TGGE

Molecular fingerprinting techniques such as denaturing gradient gel electrophoresis (DGGE), terminal restriction fragment length polymorphism analysis (T-RFLP), and temperature gradient gel electrophoresis (TGGE) are the most popular today for analysis of microbial communities in environmental issues. In these methods, PCR-amplified fragments are separated by their differing mobility in a capillary or on a gel. The generated patterns based on each band or peak reflect the diversity of the microbial community. DNA fragments of each band can be separated and sequenced to give phylogenetic information for microbial strains. In nucleic acid hybridization techniques, short oligonucleotide probes specific to target microorganisms are used. The applied probe contacts the specific extracted nucleic acids.

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The dot blotting technique frequently has been used in metabolic study of a microbial community by investigation of the changes in the gene expression level. By the use of DGGE/TGGE, complex communities of bacteria and fungi can be analyzed after PCR amplification with the use of primers designed to specifically bind to conserved regions of bacterial 16S rRNA genes or fungal 18S rRNA. By the usage of group-particular primers, it is likewise possible to investigate bacterial communities, together with actinomycetes or Archaea. Although the 16S and 18S rRNA genes are most often used, other genes along with the β-subunit of bacterial RNA polymerase (rpoB) also can be used to analyze the microbial range and survival of inoculated microorganisms in soil (Dahllof et al. 2000). DGGE has been proven to be useful in monitoring the bioremediation of freshly and aged PAH-infected soils (Cunliffe and Kertesz 2006), allowing monitoring of the survival of inoculated Sphingobium yanoikuyae in addition to tracking adjustments within the local bacterial communities over time. In some studies (Zhou et al. 2009; Wang and Tam 2011; HuiJie et al. 2011), DGGE has also been used to observe microbial community dynamics and biodegradation throughout PAH biodegradation in soils and sediments. DGGE/TGGE is a preferred method although the desired information is no longer as phylogenetically exhaustive as that furnished by 16S rRNA gene clone libraries; however, these methods can determine the dominant participants of microbial communities with average phylogenetic definition (Sanz and Kochling 2007). For environmental or contaminated supply samples wherein microbial diversity is basically unknown, the DGGE/TGGE method may provide identification of the microbial population through the excision of selected bands followed through their reamplification, cloning, and sequencing, which can indicate the phylogenetic affiliation of the ribotypes (Evans et al. 2004; Forney et al. 2004; Van Elsas et al. 2007). DGGE specifically has been widely used for the evaluation of microbial community shape in infected soil and water (Chang et al. 2000; Ralebitso et al. 2000; Watanabe et al. 2000; Kleikemper et al. 2002; Cummings et al. 2003; El-Latif Hesham et al. 2006; Mahmoud et al. 2009). Apart from microbial network profiling, the DGGE technique has been used to study gene clusters such as dissimilar sulfite reductase β-subunit (dsrB) genes in sulfate-decreasing bacterial communities (Geets et al. 2006) and benzene, toluene, ethylbenzene, and xylene (BTEX) monooxygenase genes from bacterial traces obtained from hydrocarbon-polluted aquifers (Hendrickx et al. 2006). Coulon et al. (2012) used DGGE evaluation of opposite-transcribed bacterial 16S rRNA from the upper 1.5 cm of a hydrocarbon-polluted sediment in coastal mudflats to examine the essential function of dynamic tidal biofilms by using aerobic hydrocarbonoclastic bacteria and diatoms in the biodegradation of hydrocarbons. Their research found phylotypes related to straight chain and polycyclic hydrocarbon degradation including Cycloclasticus, Alcanivorax, Oleibacter, and Oceanospirillales strain ME113.

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Single-Strand Conformation Polymorphism (SSCP)

Single-strand conformation polymorphism (SSCP), which is also dependent on electrophoretic separation based on differences in DNA sequences, allows differentiation of DNA molecules having the same length but different nucleotide sequences. This technique was originally developed to detect known or novel polymorphisms or point mutations in DNA (Peters et al. 2000). In this method, single-stranded DNA separation on polyacrylamide gel was based on differences in mobility resulting from their folded secondary structure (heteroduplex). As formation of folded secondary structure or heteroduplex and hence mobility is dependent on the DNA sequences, this method reproduces genetic diversity in a microbial community. All the limitations of DGGE are equally applicable for SSCP. Again, some singlestranded DNA can exist in more than one stable conformation, so that the same DNA sequence can produce multiple bands on the gel (Tiedje et al. 1999). However, it does not require a GC clamp or the construction of gradient gels and has been used to study bacterial or fungal community diversity (Stach et al. 2001). SSCP has been used to measure the succession of bacterial communities (Peters et al. 2000), rhizosphere communities (Schmalenberger et al. 2001), bacterial population changes in an anaerobic bioreactor (Zumstein et al. 2000), and arbuscular mycorrhizal fungi (AMF) species in roots (Kjoller and Rosendahl 2000).

5.6

Restriction Fragment Length Polymorphism (RFLP)

Restriction fragment length polymorphism (RFLP), another method for reading microbial range, depends on DNA polymorphisms. During previous years RFLP packages have also been implemented to estimate diversity and network structure in extraordinary microbial groups (Moyer et al. 1996). In this approach, electrophoresed digests are blotted from agarose gels onto nitrocellulose or nylon membranes and hybridized with suitable probes prepared from cloned DNA segments of associated organisms. RFLP can be very useful, especially in a mixture with DNA–DNA hybridization and enzyme electrophoresis, for the differentiation of carefully associated lines (Palleroni 1993); the technique also appears to be useful for detection of intraspecies variants (Kauppinen et al. 1994). RFLPs can also provide a easy and effective device for the identity of bacterial lines at and below species level. This method is helpful for detecting structural modifications in microbial groups, although not for degree of diversity or detection of precise phylogenetic agencies (Liu et al. 1997). Banding patterns in various communities are too complicated to analyze by RFLP because a single species may have four to six limit fragments (Tiedje et al. 1999). However, one must be aware that a similar banding pattern does not necessarily indicate a very close connection among the organisms compared.

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Terminal Restriction Fragment Length Polymorphism (T-RFLP)

Terminal restriction fragment duration polymorphism (T-RFLP) overcomes a number of the limitations of RFLP (Thies 2007). T-RFLP is an extension of RFLP analysis that offers an alternative approach for rapid analysis of microbial network range in diverse environments. It follows the same precept as RFLP except that one PCR primer is categorized with a fluorescent dye, together with TET (4,7,20 ,70 -tetrachloro-6-carboxyfluorescein) or 6-FAM (phosphoramidite fluorochrome 5-carboxyfluorescein). PCR carried out on sample DNA uses regularly occurring l6S rDNA primers, one of which is fluorescently categorized. Fluorescently labeled terminal restriction fragment duration polymorphism (FLT-RFLP) patterns can then be created by digestion of labeled amplicons using restriction enzymes. Fragments are then separated by gel electrophoresis on an automated collection analyzer. Each precise fragment length can be counted as an operational taxonomic unit (OTU), and the frequency of each OTU may be calculated. The banding sample may be used to measure species richness and evenness in addition to similarities between samples (Liu et al. 1997). T-RFLP, as every totally PCR-based approach, may underestimate authentic range because numerically dominant species are detected by the massive amount of template DNA (Liu et al. 1997). Incomplete digestion by means of restriction enzymes can also cause an overestimation of diversity (Osborn et al. 2000). Despite these obstacles, some researchers believe that when standardized, T-RFLP can be a useful tool to observe microbial variety in the environment (Tiedje et al. 1999), although others consider it to be inadequate (Dunbar et al. 2000). T-RFLP has been used to measure spatial and temporal modifications in bacterial groups (Lukow et al. 2000), to study complex bacterial groups (Moeseneder et al. 1999), to screen populations (Tiedje et al. 1999), and to evaluate the diversity of arbuscular mycorrhizal fungi (AMF) within the rhizosphere of Viola calaminaria in a metal-contaminated soil (Tonin et al. 2001). Tiedje et al. (1999) reported five instances of greater success at detecting and tracking specific ribotypes by using T-RFLP rather than DGGE. Dojka et al. (1999) monitored microbial variety in a hydrocarbon- and chlorinated solvent-infected aquifer undergoing intrinsic bioremediation with T-RFLP, determining sequence types characteristic of Syntrophus spp. and Methanosaeta spp. They hypothesized from their findings that the terminal step of hydrocarbon degradation inside the methanogenic zone of the aquifer became aceticlastic methanogenesis, with these organisms existing in a syntrophic relationship. Bordenave et al. (2004) studied bacterial network modifications in microbial mats following crude oil pollution using T-RFLP. Their results indicated a clean succession of various bacterial populations with operational taxonomic units that were associated with the genera Chloroflexus, Burkholderia, Desulfovibrio, and Cytophaga. T-RFLP has also been used to characterize microbial groups recovered from surrogate minerals incubated in an acidic uranium-contaminated aquifer (Reardon et al. 2004) and dechlorinating microorganisms from a basalt aquifer

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(Macbeth et al. 2004; Fahy et al. 2005). The use of T-RFLP showed that the chronic presence of benzene in groundwater reduced bacterial range and network composition in comparison with that of available groundwater resources. The use of automatic detection systems and capillary electrophoresis in T-RFLP analysis permits high throughput and greater correct quantitative analysis of microbial community samples than with any of the opposite genetic fingerprinting strategies.

5.8

Ribosomal Intergenic Spacer Analysis (RISA) and Amplified Ribosomal DNA Restriction Analysis (ARDRA)

RISA makes use of the length and sequence heterogeneities present within the intergenic spacer (IGS) among the small (SSU) and large subunit (LSU) rRNA genes inside the rRNA operon (Van Elsas et al. 2007). Similar to RFLP and T-RFLP, RISA and ARDRA provide ribosomal-based fingerprinting of the microbial network. In RISA the IGS location between the 16S and 23S ribosomal subunits is amplified through PCR, denatured, and separated on a polyacrylamide gel under denaturing conditions. This location may encode tRNAs and can differentiate between bacterial lines and closely related species by the heterogeneity of the IGS duration and series (Fisher and Triplett 1999). Sequence polymorphisms are detected through silver staining in RISA. RISA has been used to evaluate microbial range in soil (Borneman and Triplett 1997), inside the rhizosphere of plant life (Borneman and Triplett 1997), in contaminated soil (Ranjard et al. 2000), and in response to inoculation (Yu and Mohn 2001). RISA is a very rapid and easy fingerprinting approach but its application in microbial community evaluation from contaminated assets is restricted, partly because the database for ribosomal intergenic spacer sequences is not as large or as complete as the 16S collection database (Spiegelman et al. 2005). Limitations of RISA include requirement of massive portions of DNA, the longer time requirement, insensitivity of silver staining in some instances, and acrylamide choice (Fisher and Triplett 1999). Banding patterns in ARDRA may be used to display clones or the degree of bacterial network shape (Kirk et al. 2004). ARDRA is easy, rapid, and cost-efficient and thus has been utilized in microbial identification (Vaneechoutte et al. 1995; Kita-Tsukamoto et al. 2006; Krizova et al. 2006) and microbial network research (Weidner et al. 1996; Bai et al. 2006; Babalola et al. 2009). Microbial network composition and succession in an aquifer exposed to phenol, toluene, and chlorinated aliphatic hydrocarbons have been assessed by means of ARDRA to identify the dominant microbial community involved in the biodegradation of trichloroethylene (TCE) following biostimulation (Fries et al. 1997). In another study, (Gich et al. 2000) used ARDRA to examine the microbial differences in activated sludge from remedial plants consumed at home or from industrial wastewater. Bacterial groups in activated sludge could be distinguished between commercial and domestic wastewater treatment plant life. Hohnstock-Ashe

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et al. (2001), using ARDRA as a fingerprinting method, also discovered that microbial community composition in waters contaminated with TCE had shifted towards a fairly diverse community dominated by Dehalococcoides ethenogeneslike microorganisms. Babalola et al. (2009) used ARDRA to study the phylogenetic relationships of actinobacterial populations associated with Antarctic valley mineral soils. Further sequencing of the amplicons restricted singly with endonucleases RsaI, BsuRI, or AluI determined that the phylotypes were most closely related to uncultured Pseudonocardia and Nocardioides spp. In contrast, complementary traditionally established research had isolated more species of Streptomyces that were detected at a low frequency in metagenomic analysis. ARDRA is useful in detecting structural adjustments in microbial groups but cannot display microbial variety or detect unique phylogenetic groups within a network fingerprinting profile (Liu et al. 1997). Optimization with restriction enzymes is required, which frequently is difficult if sequences are unknown. Thus, further optimization can be required to produce fingerprinting styles characteristic of the microbial community (Vaneechoutte et al. 1995; Spiegelman et al. 2005). In addition, banding styles in diverse groups appear to be too complicated to use ARDRA (Kirk et al. 2004). In recent research, ARDRA has been blended with different molecular techniques including T-RFLP and DGGE to signify microbial groups from contaminated sources (Watts et al. 2001; Haack et al. 2004). An important assignment in the usage of ARDRA lies within the interpretation of the fingerprints obtained from complicated microbial communities.

5.9

Fluorescent In Situ Hybridization (FISH)

A critical step towards determining the variety of microbes in environmental samples is to harness the statistics received from the direct sequencing of rRNA genes extracted from such samples. The full-cycle rRNA method essentially makes use of the series statistics of cloned, rRNA-encoding genes from environmental habitats to broaden phylogenetic oligonucleotide probes that permit specific hybridization to the goal region of the ribosomal RNA in constant permeabilized cells. Referred to as fluorescent in situ hybridization (FISH) (Van Elsas et al. 2007), this is a speedy and sensitive approach that allows the direct visualization and identity of environmental microorganisms without culturing (Bakermans and Madsen 2002). Microbial cells are penetrated with fixatives, hybridized with precise probes (generally 15- to 25-bp oligonucleotide fluorescently labelled probes) on a tumbler slide, then visualized with epifluorescence or confocal laser microscopy (Malik et al. 2008). Hybridization with rRNA-targeted probes enhances the characterization of uncultured microorganisms and also enables the description of complicated microbial communities (Edgcomb et al. 1999). FISH is a taxonomic approach used more often than not to determine whether participants of a particular phylogenetic affiliation are present; it offers direct visualization of uncultured microorganisms and also facilitates the quantification of precise microbial organizations (Sanz and Kochling 2007). FISH

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use alone does not provide any perception of the metabolic function of microorganisms. However, it may be combined with different techniques such as microautoradiography to describe the practical properties of microorganisms in their natural surroundings (Wagner et al. 2006). Two kinds of FISH probes based on conserved or precise areas of 16S rRNA genes may be advanced: domain- or organization-specific probes and stress-unique probes. Domain- or institutionspecific probes discriminate or indicate participants of larger phylogenetic assemblages, even as stress-particular probes quantify or investigate the abundance of a specific species or presence inside a microbial network (Dubey et al. 2006). Richardson et al. (2002) blended institution-unique FISH and T-RFLP in the characterization of microbial groups involved in TCE biodegradation. From the FISH analysis, the authors found that a wide variety of organisms consisting of Cytophaga, Flavobacterium, and Bacteroides had greater abundance than the TCE degrader Dehalococcoides ethenogenes in the microbial consortium. However, the lack of useful gene evaluation meant that the relative abundance of these organisms and their ecological importance for TCE biodegradation could not be established. FISH has been used to analyze microbial community composition in PAH-infected soils, in particular detecting the 16S and 23S rDNA genes (Van Herwijnen et al. 2006; Hesham et al. 2012; Chang et al. 2014). However, unique care must be taken using FISH on soil or sediment samples. FISH techniques are frequently used with different genetic fingerprinting strategies including DGGE (Onda et al. 2002; Collins et al. 2006) and T-RFLP (Richardson et al. 2002; Kotsyurbenko et al. 2004; Jardillier et al. 2005; Collins et al. 2006) for the enumeration and characterization of microbial populations from contaminated resources. The downside of FISH is that a confined variety of probes may be utilized in a single hybridization experiment and the fluorescence history thus can be complex in some samples (Dubey et al. 2006; Sanz and Kochling 2007). An earlier examination of the sample and the maximum microorganisms that may be detected is important (i.e., rRNA sequence) for the selection of specific probes. A predominant obstacle of the same FISH technique is its restricted sensitivity because bacterial cells with decreased ribosome content that frequently appear in oligotrophic environments, such as maximum soil habitats, are not satisfactorily stained for microscopic analyses. Modifications of FISH, such as the catalyzed reporter deposition fluorescence in situ hybridization (CARD-FISH) and fluorescence in situ hybridization micro-autoradiography (FISH-MAR) are being used as alternatives to circumvent these issues and measure the potential for PAH degradation in soils (Teira et al. 2007; Lekunberri et al. 2010). In CARD-FISH, signal intensities of hybridized cells are multiplied by enzymatic signal amplification using horseradish peroxidase (HRP)-categorized oligonucleotide probes in combination with tyramide signal amplification (TSA). TSA is based on the patented catalyzed reporter deposition (CARD) method using derivatized tyramide. In the presence of small quantities of hydrogen peroxide, immobilized HRP converts the categorized substrate (tyramide) into a short-lived, extremely reactive intermediate. The activated substrate molecules then very swiftly react with, and covalently bind to, electron-rich areas of adjacent proteins. This binding

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of the activated tyramide molecules occurs best without delay adjacent to the websites at which the activating HRP enzyme is certain. Multiple deposition of the categorized tyramide takes place in a brief time (normally less than 3–10 min). Subsequent detection of the label yields a successfully massive amplification of sign (Van Elsas et al. 2007). The mixture of FISH with micro-autoradiography allows the identification of bacteria and concomitantly illustrates their particular in situ interest for the usage of appropriate isotope-categorized substrates (in particular, β-emitters such as 14C and 3H) (Van Elsas et al. 2007). The substrate uptake patterns in FISHclassified bacteria can be investigated in situ in combined natural groups at an unmerged cellular stage although the bacteria are not yet culturable.

5.10

DNA Microarray and Reverse Sample Genome Probing (RSGP)

DNA–DNA hybridization has been used collectively with DNA microarrays to assess bacterial species (Cho and Tiedje 2001) or to evaluate microbial range (Greene and Voordouw 2003; DeSantis et al. 2007) with great specificity. This is a chip technology containing nucleic acids as probes that are ideal for the highthroughput analysis of the collection range of 16S rRNA genes as well as of other useful genes in environmental samples (Van Elsas et al. 2007; Malik et al. 2008). In contrast to FISH, it offers a method for simultaneous evaluation of many genes (Cho and Tiedje 2002). The DNA microarray is a miniaturized array of complementary DNA probes (~500–5000 nucleotides in length) or oligonucleotides (15–70 bp) connected to a strong support, which allows simultaneous hybridization against a very large set of probes complementary to their corresponding DNA/RNA targets in a pattern. The application of microarrays in environmental microbiology, especially in the examination of microbial populations engaged in biodegradation, can identify organisms in addition to defining their ecological position (Wu et al. 2001). However, extra-rigorous and systematic evaluation and improvement are needed to comprehend the overall ability of microarrays for microbial detection and community analysis (Zhou 2003). Microarrays detect the simplest dominant populations in many environments (Rhee et al. 2004). In addition, probes designed to be precise to acknowledge sequences can move or hybridize to similar or unknown sequences and might produce deceptive alerts (Gentry et al. 2006). Moreover, soil, water, and sediments regularly contain humic acids and different organic substances that may also inhibit DNA hybridization on microarrays (Saleh-Lakha et al. 2005). Finally, limits in RNA extraction from many environmental samples imply that advances in RNA extraction and purification and amplification techniques are needed to make microarray gene expression evaluation possible for a broader variety of samples (Gentry et al. 2006).

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Another DNA microarray-based approach for reading a microbial community is reverse sample genome probing (RSGP). This technique uses genome microarrays to research the microbial network composition of the most dominant culturable species in an environment. RSGP has three steps: isolation of genomic DNA from pure cultures, hybridization checking to attain DNA fragments with much less than 70% cross-hybridization and preparation of genome arrays on a strong support, and random labelling of a defined mixture of general network DNA and general inner DNA (Greene and Voordouw 2003). This approach has been used to investigate microbial groups in oil fields and in polluted soils (Greene et al. 2000). As in DNA– DNA hybridization, RSGP and microarrays have the advantage of not being confounded via PCR biases. Microarrays can incorporate many target gene sequences, but are handiest in detecting the maximum plentiful species. In this fashion, species need to be cultured; however, in precept cloned DNA fragments of uncultured bacteria may also be used. The variety must be minimum, or enriched cultures will be needed for this approach. Otherwise, pass-hybridization can be intricate. Using genes or DNA fragments rather than genomes in the microarray removes the need to maintain cultures of live organisms, because genes can be cloned into plasmids or PCR can be consistently used to amplify the DNA fragments (Gentry et al. 2006). In addition, fragments would increase the specificity of hybridization over using genomes, and useful genes within the community might be assessed (Greene and Voordouw 2003).

6 Omics in Bioremediation Various ‘omics’ strategies have opened the way for environmental probing at the molecular level and have also created a new paradigm in bioremediation design and control. Ecogenomics – the utility of genomics to ecological and environmental sciences – defines phylogenetic and functional biodiversity on the DNA, RNA, and protein levels, with emphasis on the capabilities and interactions of organisms in the surroundings relative to ecological and evolutionary methods. For effective bioremediation of halo-organic pollution in anaerobic environments, we need the information from catabolic ability and in situ dynamics of organohalide-consuming and cometabolizing microbes. Current methods of the OMICS pathways are outlined in Fig. 18.1.

6.1

Genomics

Genomics is a powerful computer-based technology used to understand the structure and feature of all genes in an organism. Study of pure cultures have become very easy using the software of genomics to bioremediation. Next-era genome sequencing strategies will be critical in elaborating the physiological and genomic features of

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Fig. 18.1 Bioremediation monitoring and “OMICS”

microorganisms applicable to bioremediation. Complete genome sequences should be known for such microbes, which can be important in bioremediation (Table 18.4). The views of researchers changed after observing the utility of bioremediation to the advanced sciences such as genomics, which have given exclusive results to many questions. For instance, molecular analyses have indicated that Geobacter species are crucial for the bioremediation of organic and metal contaminants in subsurface environments. This realization led to the sequencing of several species of the genus Geobacter, in addition to other related organisms, which has altered the concept of how Geobacter species help in reducing the contaminants in subsurface environments. Earlier it was thought that Geobacter species were nonmotile, but genes encoding flagella were subsequently determined in the Geobacter genomes (Childers et al. 2002). Later it was found that Geobacter metallireducens in particular produces flagella only when the microbe is growing on insoluble ferrous and manganese oxides. Pili genes were also found to be expressed during culture on insoluble oxides (Childers et al. 2002). Similarly, other microbes with bioremediation capability have been screened for complete genome sequencing. With the completed genome sequences, it is better to design whole-genome DNA microarrays to analyse the expression of all the genes under various environmental situations. When bioremediation approaches are researched in detail, trials are commonly made to isolate the concerned microorganisms (Rogers and McClure 2003). The isolation and characterization of such

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Table 18.4 Genomes of microorganisms pertinent to bioremediation

Microorganism Dehalococcoides ethanogenes Geobacter sulfurreducens Geobacter metallireducens Rhodopseudomonas palustris Pseudomonas putida Dechloromonas aromatica Desulfitobacterium hafniense Desulfovibrio vulgaris Shewanella oneidensis Deinococcus radiodurans

Relevance to bioremediation Dechlorination of solvents to ethylene in reduced conditions Oxidation of aromatic hydrocarbons anaerobically and precipitation of uranium in reduced conditions

Anaerobic metabolism of aromatic compounds Aerobic degradation of organic contaminants and capabilities for genetic engineering Perchlorate-reducing microbes capable of anaerobic oxidation of benzene Dechlorination of chlorinated solvents and phenols under reductive conditions Reductive precipitatation of uranium and chromium Reduction of uranium in vitro Resistant to radiation and thus can be used in genetic engineering for bioremediation of highly radioactive environments

Website for genome documentation http://www.tigr. org http://www.jgi. doe.gov http://www.tigr. org http://www.jgi. doe.gov http://www.tigr. org http://www.jgi. doe.gov http://www.jgi. doe.gov http://www.tigr. org http://www.tigr. org http://www.tigr. org

microbes has been and will remain critical for the development and interpretation of molecular analyses in microbial ecology. However, before the utility of molecular techniques to bioremediation, it was not sure whether the isolated microbes had been critical in bioremediation under natural conditions, or whether they grew as “weeds” unexpectedly in the laboratory, but were now not the expected organisms responsible for bioremediation.

6.2

Transcriptomics

The transcriptome is the subset of genes transcribed in any given organism, a dynamic link among genome, proteome, and cellular phenotype. The control of gene expression is the key in modifying changes in environmental situations and hence for survival. Transcriptomics describes this expression procedure in a genome-extensive manner. DNA microarrays allow detection of RNA expression of each gene of an organism (Diaz 2004; Golyshin et al. 2003; Gao et al. 2004; Muffler et al. 2002; Schut et al. 2003). One important challenge in this method is elucidation of the huge amounts of data recorded (Dharmadi and Gonzalez 2004). A global gene expression pattern analysis revealed the numerous to-date-unknown

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genes at some stage in the degradation of alkyl benzenes (Kuhner et al. 2005). Besides this, DNA microarrays were used to decide bacterial species, in quantitative programs of stress gene analysis of microbial genomes and in genome-wide entranscriptional profiles (Greene and Voordouw 2003; Muffler et al. 2002).

6.3

Proteomics

Proteomics is the study of all the proteins in an organism. Among many functions these proteins serve as catalytic enzymes in metabolic pathways and in signal transduction of regulatory pathways of cells (Graham et al. 2007; Zhang et al. 2010). Environmental conditions can lead to differences in their cellular expression in an organism, and the toxicity present in the environment acts as a trigger in adjusting to the physiological response. Proteomics has made it possible to acquire an in-depth overview of global modifications and the abundance of proteins, apart from identification of major proteins involved in bioremediation at a given physiological state (Vasseur et al. 1999; Wilkins et al. 2001). Various reports have suggested that proteins are up- or downregulated in response to the contamination (Vasseur et al. 1999; Wilkins et al. 2001; Kim et al. 2002). The downregulated proteins were found to be a part of nucleotide biosynthesis and cellular motility (Santos et al. 2004). A proteomic evaluation also revealed the participation of energy- and stress-related proteins when Pseudomonas putida DOTT1E was exposed to toluene (Segura et al. 2005). Such detailed facts are important for the development of bacteria with greater solvent or contaminant tolerance that can be used for bioremediation.

7 Metabolomics Metabolomics is a young and vibrant discipline of technological advancement in its exponential increase. Metabolome evaluation has become very popular currently, and novel strategies for obtaining and reading metabolomics statistics that are useful for different biological studies are increasing. Bioremediation is one such field which gains from the advances in this emerging field. Various bioremediation research options such as finding strategies for elucidation of biodegradation pathways, using isotope distribution analysis and molecular connectivity analysis, assessing the mineralization system using metabolic footprinting evaluation, and improving the biodegradation system through metabolic engineering can be analysed in a more comprehensive manner using metabolomics.

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Metabolomics, Metametabolomics, and Fluxomics

Metabolomics is analysis of all the cellular metabolites in a cell and their interactions in the microbial community. Analysis of cellular metabolites within a cell and community over a time period in real time is known as fluxomics (Wiechert et al. 2007). Information on factors regulating growth and metabolism of microbial communities can be accessed by metabolomics, and fluxomics can provide the missing links in the regulatory pathways involved in metabolism of environmental pollutants. A microbial cell releases a number of low molecular weight primary and secondary metabolites in response to an environment challenge or stress. The influence of the local environment on the metabolome of an organism or community can be exploited in bioremediation to characterize the effects of xenobiotic compounds. Metabolites can be characterized by mass spectrometry and various spectroscopic techniques. Villas-Boas and Bruheim (2007) have discussed the scenario of metabolome analysis in bioremediation. The application of metabolomics can significantly extend and enhance the power of existing bioremediation approaches by providing a better understanding of the biodegradation process (Villas-Boas and Bruheim 2007). Keum et al. (2008) studied the metabolic profiles of Sinorhizobium sp. C4 using mass spectrometry and gas chromatography during the degradation of phenanthrene in comparison to natural carbon sources. Tang et al. (2009) also performed a fluxomics analysis on Shewanella sp. known to have co-metabolic pathways for bioremediation of halogenated organic compounds, toxic metals, and radionuclides. In addition, they analyzed that a mixed bacterial consortium can degrade benzene and its derivatives and other aromatic ring organic members more than 97%. Maphosa et al. (2010) studied organ halide-catabolizing bacteria and its molecular diagnostics with mass-balancing and kinetic modeling in an in situ dechlorinating bioreactor and showed its application in monitoring bioremediation.

7.2

Improvement of the Biodegradation Process via Metabolic Engineering

Recombinant DNA technology can drastically improve metabolic engineering (ME) and thus improve cellular properties or introduce new ones (Stephanopoulas et al. 1998). The metabolic engineering (ME) approach has been successfully used for the improvement of industrial microorganisms. However, the use of ME tools and principles is also relevant for bioremediation, as potential natural microbial populations to degrade contaminations are found in fewer numbers (Urgun-Demitras et al. 2006; Pieper and Reineke 2000; Rui et al. 2004). Accumulation of contaminants in the environment represents a potential pollution problem because many such are highly toxic, mutagenic, or carcinogenic. Metabolic engineering can lead to both the introduction of novel biodegradation pathways and to modification and extension of a substrate range of enzymes in existing degradation pathways in the

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development of microorganisms with improved bioremediation properties (Dua et al. 2002). A psychrotolerant nonpathogenic Pseudomonas fluorescens strain that was genetically engineered for degradation of 2,4-dinitrotoluene was studied by transferring the Burkholderia sp. strain DNT pJS1 mega plasmid containing the dnt genes (Monti et al. 2005). Thus, a stable 2,4-dinitrotoluene-degrading phenotype was first achieved by integrating the dnt genes into the P. fluorescens chromosome. Haro and Lorenzo (2001) combined two catabolic segments from the toluene (TOL) and toluene dioxygenase (TOD) pathways of Pseudomonas putida to create an upper hybrid pathway for bioconversion of 2-chlorotoluene into 2-chlorobenzoate. This pathway was integrated into the pathway of two 2-chlorobenzoate-degrading Pseudomonas strains, expecting to provide complete degradation of 2-chlorotoluene. The two strains were able to co-metabolize 2-chlorotoluene to 2-chlorobenzoate with citrate as co-substrate but failed to grow on 2-chlorotoluene alone. Analysis of the cultures with gas chromatography–mass spectroscopy (GC-MS) showed that citrate-grown cells accumulated 2-chlorobenzoate and other intermediates in the upper pathway during exposure of 2-chlorotoluene. However, no reason was found for the lack of degradation of 2-chlorobenzoate. This study clearly emphasizes that an unbiased and nontarget analysis of the performance and physiology of the biological model system via metabolomics is necessary. Metabolomics and other genome-wide methodologies as transcriptomics, proteomics, interactomics, and fluxomics have became important tools for describing how a phenotype is generated from its genotype and the environmental conditions. These technologies became important for both metabolic engineering and system biology because the focus is on integrated networks of metabolic pathways and not on individual isolated reactions (Friboulet and Thomas 2005). The strategy of inverse metabolic engineering is through comparisons between mutated strains and reference strains to gain insight into the metabolism (Bro and Nielsen 2004). In bioremediation, this approach is also relevant (e.g., comparison of reference strains with recombinant strains modified with a new degradation pathway). Also important is the exploration of the response of the microorganisms to pollutant exposure. A metabolome analysis can supply information regarding which pathways are activated under expression of the new heterologous pathway and under pollutantexposure conditions. The strength of metabolome analysis is that it points out the pathways that need to be targeted by metabolic engineering (Trethewey 2001). Because the other OMICS-related approaches provide invaluable complementary information, the future of metabolic engineering is to integrate the different sets of data to develop optimal strains for use in both bioprocessing and bioremediation (Park et al. 2005). Lee et al. (2006) used a quantitative proteomics approach to provide important insights into the metabolic and physiological changes that occur upon cis-dichloroethylene degradation by engineered Escherichia coli strains. Because the number of identified central proteins in the degradation pathways was less than the number of genes, a transcriptome analysis may complement the proteomics approach as would the metabolomics and fluxomics analyses on the physiological characterization.

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8 Bioinformatics Bioinformatics is an important part of modern biotechnology that combines biology and information technology. Bioinformatics seek information and keep all the biological data, helping to investigate and decide relationships among organic molecules, such as macromolecular sequences, structures, biochemical pathways, and expression profiles. Biology and computer systems are running parallel, which might be collectively respecting, supporting, and influencing each other and synergistically merging as greater than ever (Fulekar 2008). Enormous amounts of data and statistics from biology, especially within the form of DNA, RNA, and protein sequences, are slowly accumulating, thus placing heavy demand on computers and scientists. Bioinformatics has also led to finding possible cures for detoxification of the environment. Scientists expert in analyzing biological data can use the computational tools to solve the problems of bioremediation (Westhead et al. 2003). The important branches of bioinformatics are genomics, transcriptomics, proteomics, organic databases, molecular phylogenetics, and microarray informatics, which are crucial in understanding bioinformatics. These bioinformatics-associated tools are very imporant for bioremediation of hazardous wastes.

8.1

Bioinformatics in Bioremediation-MetaRouter

MetaRouter is an application for laboratories working in bioremediation that need to maintain public and private data, linked internally and with external databases, and to extract new information from it. The system has a very versatile open and modular architecture adaptable to different customers. It is a multiplatform program, implemented in Postgre SQL (standard language for relational databases) and using SRS as an indexing system (used to connect and query Molecular Biology databases). MetaRouter uses a client/server architecture that allows the program to run on the user station or on the company server, so it can be accessed from any place in a secure way just by having a web browser. Bioinformatics will facilitate and quicken the analysis of cellular processes and also the investigation of the cellular mechanisms used to treat wastes. Coming decades will yield greater understanding of molecular mechanisms and cellular metabolism coupled with bioinformatics.

8.2

Bioinformatics Resources

Many bioinformatic resources are available today exclusively for bioremediation studies and analyses. In 1955 University of Minnesota Biocatalysis/Biodegradation Database (UM-BBD, http://umbbd.msi.umn.edu/) was developed and has been

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regularly updated since its inception. It has a collection of databases related to organic compounds, enzymes, reactions, biotransformation rules, and different microorganisms. Further, it contains a biochemical periodic table (UM-BPT) and a pathway prediction system (UM-PPS) that predicts hypothetical pathways for microbial degradation. Other scientists are accessing UM-BBD data for their research purposes. Also, UM-BBD compound data are now contributed to PubChem and ChemSpider, which are other public chemical databases. Another database is being developed at ETH Zürich to improve the speed and reliability of online access from anywhere in the world (Gao et al. 2010). Urbance et al. (2003) have developed a database to provide detailed information on degradative bacteria and the hazardous substances they degrade (Biodegradative Strain Database, BSD) within the phylogenetic framework of the Ribosomal Database Project II (RDPII: http://rdp.cme. msu.edu/html). Pazos et al. (2003, 2005) developed a unique system, MetaRouter, for maintaining the variety of information related to bioremediation in a way that allows its query and mining. Arora et al. (2009) compiled a database of biodegradative enzyme oxygenases (OxDBase), taken from primary literature and converted in the form of a web-accessible database, that consists of two separate search engines for mono- and dioxygenases, respectively. For each enzyme queried the database shows its common name and synonym, family and subfamily, reaction in which enzyme is involved, structure and gene link, and literature citation. It can be also linked to several other external databases including ENZYME, BRENDA, KEGG, and UM-BBD, providing more background information. OxDBase, the first database specially dedicated to oxygenases, consists of more than 200 databases that provide comprehensive information about them. Moriya et al. (2010) developed PathPred, a web-based server focused on predicting pathways for microbial biodegradation. The server provides a platform to study the transformation patterns and reference transformation patterns in each predicted reaction. These patterns are then displayed in a tree-shaped graph. The transformation of xenobiotic compounds by microorganisms is essential for the bioremediation, and there is no single resource available that provides information about environmental contaminants as well as microorganisms with biodegradative capabilities. Thus, a database that consolidates the detailed information about xenobiotics, biotransformation methods, and hypothetical pathways would be a handy tool for academic and industrial researchers (Urbance et al. 2003).

9 Future Challenges Advances in molecular biology techniques have opened a new era in bioremediation processes. Further, the development of various OMICs approaches offers efficient research avenues for analyzing biological systems at different levels. However, the utility of genome sequence facts and associated numerous OMIC approaches to link microbial genes and their products with bioremediation potential is an incredible mission. The goal of this chapter was to review a number of the most current

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advances in molecular techniques and OMICs methods and point out some future challenges. 1. Can these metagenomics approaches be powerful in tracking microbial diversity and community dynamics at contaminated sites? If so, what are the essential goals to enforce metagenomics methods to bioremediation? 2. Can proteomic techniques be used as a quantitative tool for tracking in situ microbial metabolism? What are the technologically demanding situations in making efficient use of such equipment in the field? 3. Can these microarrays be used as a excessive throughput, particular, sensitive, and quantitative device for tracking microbial populations and activities? 4. Can OMICs tactics be used to clarify relationships among biomarker estimations and in situ activities? 5. What is the level of correlation between the abundance and expression of the practical genes and the rates of degradation or transformation of contaminants? Can methods primarily based on OMICs offer insights into the costs of kinetic predictions? 6. Can the collected data and information be scaled from molecules to subsurface ecosystems for improving our predicative capability of field bioremediation?

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Conclusions

Molecular biology approaches have shown an exponential rise in the past few decades and, together with the latest developments in molecular nanotechnology, will lead to further insight into bioremediation processes. Scientists are trying to develop new bioremediation approaches by designing artificial pathways, developing databases for specific microbes or enzymes, and other inter- and intradisciplinary metabolic pathways related to bioremediation. During this entire chapter, one common point is that in spite of the disadvantages of all the molecular-based approaches, the idea that such cellular systems work has been clarified at some stage. In précis, the OMICs tactics should be developed in a manner to assist finding solutions to ecological questions in the context of biochemistry, physiology, microbiology, geochemistry, and ecology.

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Chapter 19

Biochar: A Sustainable Tool in Soil Pollutant Bioremediation Chhatarpal Singh, Shashank Tiwari, and Jay Shankar Singh

Abstract Soil is a vital reservoir of living being likely bacteria, fungi, algae, protozoa, etc. They dynamically standardize ecosystem functioning but, due to some imbalance and unstoppable anthropogenic activities, for instance, industrialization, urbanization, and wrong agricultural practices, cause soil pollution, eventually resulting in various environmental health hazards. Although there is no any single factor that is responsible for leading these challenges, many more other activities are involved in a direct and indirect manner to creating environmental pollution. Hence newly developed sustainable, cost-effective, and different feedstock-mediated carbon-rich by-product is a unique and multifunctional sorbent called “biochar” that can play a vital role in bioremediation of several highly hazardous petroleum refinery wastes containing different types of aliphatic, aromatics, other complex hydrocarbons, and heavy metals in contaminated soils due to the longtime recalcitrant nature against microbial degradation. Currently, biochar is used as carrier sorbent for various microorganisms since they stimulate the in situ bioremediation of several hazardous polycyclic aromatic hydrocarbon (PAH) compounds and heavy metals, due to the large surface area and micropores; consequently pollutants are adsorbed on the surface. Biochar may work singly and along with manure compost and remediates many hazardous pollutants from contaminated soils. Keywords Biochar · Bioremediation · Soil pollutants · Microorganism · Agricultural waste

C. Singh · S. Tiwari · J. S. Singh (*) Department of Environmental Microbiology (DEM), School for Environmental Sciences (SES), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_19

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1 Introduction Soil is a large basin for agricultural production and associated biological activity on this planet, but explosive growth in human population is a challenging task to maintain all natural sources for feeding the food demands all over the world. Therefore, soil is the rudimentary and life supportive component of the environment that contains several minerals and biological forms. Microorganisms are the key components of the soil that play an important role to maintain soil fertility due to their involvement in several biogeochemical mechanisms (Beesley et al. 2011). But indiscriminate anthropogenic pollution is affecting environmental disparity such as industrialization, urbanization, and access application of the synthetic fertilizers, also polluting the agricultural and nonagricultural lands globally. A pollutant frequently comes from houses, cattle sheds, industries, agriculture, and other places that destroy soil or water ecosystems. Soil contamination is an excess amount of any element or chemical compound, through direct or indirect exposure, which causes a toxic response against soil biota and human being, resulting in various kinds of environmental health hazards (Adriano 2001; Abrahams 2002; Vangronsveld et al. 2009; Beesley et al. 2011). Mench et al. (2010) reported that the organic and inorganic pollutant occurs in contaminated soil worldwide. Now various environmentally acceptable techniques have been discovered for the sustainable waste management dealing with this problem. Currently, the group of experts is working in the direction of bioremediation of different contaminated soils in the cost-effective ways (Beesley et al. 2011). The in situ application of biochar in contaminated soils to bind pollutants and providing material conditions that promote plant growth and stimulate ecological diversity have become more popular (Adriano et al. 2004; Bernal et al. 2006; Vangronsveld et al. 2009). Carbon-rich amendments, such as activated carbons, have been deployed for soil and water pollutant detoxification purposes due to their ability to reduce contaminant bioavailability (Brändli et al. 2008; Cho et al. 2009) and risk. Activated carbon is a pollutant’s adsorbing material that is called biochar which is produced from the incomplete combustion of organic materials (e.g., rice husk, sugarcane bagasse, wheat straw, etc.) (Jonker and Koelmans 2002; Ghosh et al. 2003; Cornelissen and Gustafsson 2005; Brändli et al. 2008; Beesley et al. 2011). Biochar may provide an important sink for dissolved contaminants in soil where contaminants are run off and collected (Robinson et al. 2007; Beesley et al. 2011). Various anthropogenic sources (industries, agriculture, and household) release the petroleum hydrocarbons that can affect the large surface area of soil in the form of highly toxic pollutants (Singh et al. 2016a, b). Bioremediation of hydrocarbons and toxic metals through activated charcoal is an important aspect to reducing the toxic effect of various organic and inorganic pollutants (pentachlorophenol, pentachlorophenol biphenyls, benzene, dioxins, linden and arsenic, chromium, and mercury, respectively) because of relatively low costs and environmentally friendly properties in contrast to other value-adding technologies (Grace Liu et al. 2011;

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Beesley et al. 2011). Furthermore, this technique has been proved in the remediation of several petroleum-based pollutants in contaminated soils during laboratory and field study (Xu and Lu 2010; Mukherjee and Bordoloi 2011). The hydrophobic nature of most of the soil contaminants suppresses the rate of uptake and degradation of pollutants by microorganism (Semple et al. 2003). Biochar produced from biomass of organic material may sequester atmospheric CO2 in soils for a long time due to more recalcitrant properties, thus reducing the carbon sink of sorbentbased remediation in contaminated soils (Bushnaf et al. 2011). The use of biochar could be cheaper in a remediation of organic and inorganic pollutants because the waste source materials are essentially free and the production of biochar requires less energy and cost (Hale et al. 2011). However, biochar amendment to contaminated soil has also been shown to increase the pollutant’s availability for microbial breakdown by their own microbial community which is found inside biochar due to large surface area and microspores (Rhodes et al. 2008). In various current and past studies, the feasibility of the use of biochar in petroleum hydrocarboncontaminated soil remediation has been carried out (Beesley et al. 2011). In this chapter, we depict that the application of biochar in contaminated soil, affected by various organic and inorganic pollutants, may be a sustainable approach to remediate the concentration of soil pollutants.

2 What Is Biochar? Biochar is a carbon-rich by-product of different feedstocks, which is produced at higher temperature (450–600 approximately) in partial presence of oxygen that is termed as thermal degradation of biomass such as rice straw, grass, woodchip, crop residue, and sugarcane bagasse that enhances the agronomic variables and significantly remediates soil pollutants (Table 19.1: showing physicochemical properties of different feedstocks). The addition of biochar can significantly improve soil properties by decreasing methane emission and soil bulk density, enhancing soil pH and organic carbon, increasing available nutrients and the number of methanotrophs, etc. (Milla et al. 2013). The intense temperature is required for the production of high-quality biochar which can be differentiated from charcoal due to its use in soil amendments for various purposes likely fertility enhancement, crop yield, to combat with pollution, etc. (Johannes 2007; Peter 2007; Gaunt and Johannes 2008; Lehmann and Joseph 2009a, b). As stated earlier, biochar enhances the crops’ yield and mitigates environmental pollution, such as reduction of GHGs and soil pollutants (Peter 2007; Johannes 2007; Laird 2008; Ghoneim and Ebid 2013). Previously, it was reported that biochar increases the agriculture production and mitigates methane emanation. However, biochar could also remediate various hydrocarbons and toxic metals due to unique adsorption properties in soils.

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Table 19.1 Physicochemical properties of different feedstocks using biochar production

Components Soil C (%) Ash (%) pH EC (mS cm 1) CEC (c mol kg 1) C (%) N (%) S (%) Ca (ppm) K (ppm) Mg (ppm) Si mg kg 1 P (ppm)

Woodchip (Yargicoglu et al. 2015) 74.5 25.4 7.88 0.14

Grass (Jouiada et al. 2015; Mohammed et al. 2015) – 14.7 6.1 –

Poultry litter (Jindo et al. 2012) 71.47 28.53 23.596 3.0

Rice husk (Shackley et al. 2012) – 6.5 6.6 –

Sugarcane bagasse (Carriea et al. 2012) – 11.9–16.4 – –

Wheat straw (Mahinpey et al. 2009; Bruun et al. 2012) – 5.9 6.76 2770

–

–

–

45–110

–

–

51.9 0.4 – 0.56 0.21 0.04 – 0.06

42.5 1.9 5.3 4.3 4 64.80 2.3 4 7.44 2.31

38.6 1.37 – 1.85 0.99 0.19 – 0.35

41 1.4 0.1 250 2604 827 5.8 –

60.4–65.3 0.8–1.0 25.4–15.1 – – – – –

43.7 0.9 0.283 0.18 0.15 – 0.18 0.05

Adapted from Singh et al. (2017a, b, c)

3 Biochar Applications in Agro-environmental Development The biochar can play multidisciplinary role in agricultural and environmental sustainability development due the longtime persistent nature in soil. The higher concentration of stable organic matter with carbon contents was found of about 70–80% (Lehmann et al. 2002; Singh et al. 2017a) and mineral matter, including nutrients, in higher quantity. Pyrolyzed biomass of agricultural waste has high surface area, porosity, variable charge, and functional groups (Fig. 19.1) that can increase soil water-holding capacity, pH, cation exchange capacity (CEC), base saturation, bioremediation of hydrocarbons and heavy metals, and finally crop resistance to disease when added to soil in adequate amount (Glaser et al. 2002; Keech et al. 2005; Liang et al. 2006; Tang et al. 2013). These properties vary with the pyrolysis temperature and types of feedstocks (Gundale and DeLuca 2006; Bornermann et al. 2007; Chan and Xu 2009; Singh et al. 2010, 2017b). Biochar application to soil is widely advocated for a variety of reasons related to sustainability (Jeffery et al. 2015). The most often claimed benefits of biochar include carbon sequestration, soil fertility improvement, pollutant immobilization, pollutant remediation, and waste management. Furthermore, the additions of biochar to soil can increase bacterial abundance that involve in bioremediation of soil pollutants and reduce soil bulk density (Gundale and DeLuca 2006). Biochar strongly adsorbed

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Reduced methane emission

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Bioremediation of soil pollutants

Reduced nutrients leaching

Increased Soil Microbial biomass

Biochar Reduced N2O

Improved soil fertility

emission

Increased soil Carbon

Increased water holding capacity

Fig. 19.1 Applications of biochar in agricultural and environmental developments

salts and reduced the effect of salt on plants in agricultural, urban, and contaminated soils (Thomas et al. 2013; Singh et al. 2017c). Applications of biochar to soil have increased the accessibility of P and Zn and the total N concentrations in amended soil (Glaser et al. 2002; Lehmann et al. 2003). Biochar additions can increase crop yields without the use of synthetic fertilizers (Blackwell et al. 2010). Additionally, biochar can minimize the effect of soil erosion and nutrient leaching due to the strongly binding pattern to soil particles.

4 Biochar and Soil Interaction It is well known that biochar addition in soil can bind the soil particles and plant roots in which the interactions between them increase the phytoremediation as well as bioremediation. Biochar might play a significant role in in situ bioremediation of contaminated soil due to utmost binding properties to soil particles and longtime survival nature in soil; therefore, the process of bioremediation can occur by both biochar and microbial community which is present inside the biochar (inside porous region) (Singh 2011; Singh and Pandey 2013). Hence, biochar can be a better option to remediate soil contamination from affected land area due to various human

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activities. The cheapest cost of biochar is affordable to all types of application like agricultural purpose and bioremediation. Biochar supported the bacterial assisted bioremediation because the large surface area and porosity to attract higher community and surface area to adsorbed the pollutants which is present in contaminated soil. Higher shallow surface is a characteristic feature of biochar. So the size of shallow surface depends on the pyrolysis temperature and categories of feedstock. Furthermore, it has been reported that the size of surface area and pores depends on the volatile matter and tars which are filled in the feedstock; therefore, the intense temperature plays an important role in the removal of these volatile matters from the feedstock (Lehmann and Joseph 2009a, b). Joseph et al. (2010) reported that the interaction between biochar and soil not long time although it can involve all the mechanisms related to bioremediation in contaminated soil (Steiner et al. 2007; Bruun et al. 2008; Singh and Cowie 2008; Kuzyakov et al. 2009) suggested that the types and rates of interactions that take place in the soil depended on (1) feedstock composition, particularly mineral fraction; (2) pyrolysis conditions; (3) biochar particle size and delivery system; and (4) soil properties and local environmental conditions. Low-temperature biochar, which has a less-condensed “C” structure and higher nutrient content, is expected to have a greater reactivity in soils than highertemperature biochar and a better contribution to soil fertility (Steinbeiss et al. 2009; Nguyen and Lehmann 2009).

5 Remediation of Soil Pollutants Primarily, bioremediation is a process used to treat contaminated sources, such as water, soil, and other environmental materials, by stimulating growth of microorganism (Singh and Strong 2016; Singh et al. 2016a, b; Vimal et al. 2017) for the removal of contaminants by in situ and ex situ bioremediation practice (Bharagava et al. 2017a; Gautam et al. 2017; Saxena and Bharagava 2015, 2017). Additionally, some new reported approach is being used in the bioremediation such as agricultural waste-mediated biochar (Fig. 19.2) that can remediate the soil contamination through their adsorption properties due to unique surface and internal structure, etc. Biological treatment of hazardous waste is a similar approach used to treat wastes including wastewater, industrial waste, and solid waste. Industries discharged huge amount of wastewater which is released in the canal, river, and other bodies without adequate treatments of these solid and liquid wastes containing a variety of organic and inorganic pollutants that may cause health hazards to living beings (Saxena et al. 2016, 2019; Goutam et al. 2018; Bharagava et al. 2017b, c; Chandra et al. 2015). Moreover, some time due to the leakage of wastewater in agricultural fields, the soil health is affected. Wrong agricultural practices affect the soil health as well as the structure due to the greater application of various pesticides and insecticides that contain highly toxic compounds, and although household also releases huge amount of solid and liquid waste that directly enters in agricultural fields which is polluting productive land,

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Fig. 19.2 Imaginary pictorial demonstration for enhancing bioremediation of soil organic pollutants with immobilized microorganism technique and using biochar as a strong adsorption carrier. (Modified from Zhang et al. 2013)

some farmers also use this liquid household for irrigation purpose without knowing the harm of this wastewater on soil health (Singh 2013a, b, 2014, 2015a, b, c, d, 2016). So these are the major environmental degrading sources, but the use of some new and sustainable technology like biological- and chemical-based treatment could help to overcome the toxic effect of pollutants on soil productivity (Kelly et al. 2014). Remediation and rehabilitation of the contaminated soils and hazardous waste can be achieved by phytostabilization, a long-term and costeffective rehabilitation strategy, through promoting the revegetation to reduce the

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risk of pollutant transfer and ecological restoration (Fellet et al. 2011), although these are difficult without proper soil amendments (Reverchon et al. 2015). The biochar-mediated bioremediation is a sustainable approach to overcome the effect of air pollution due to direct burning of crop residue in the fields. Biochar could work as a soil fertility booster and in toxic pollutant adsorption in soil, so this practice may be beneficial on commercial and small scale (Fellet et al. 2014; Beesley et al. 2013).

6 Treatment of Inorganic and Organic Pollutants Through Biochar in Contaminated Soils The combined contamination of metals, metalloids, and organic pollutants is difficult to remediate together because each contamination requires different treatment processes. However, biochar can remediate combined form of pollutants (organic and inorganic); both contain several types of pollutants (Table 19.2). (Wang et al. 2010; Chen and Yuan 2011; Tang et al. 2013; Sneath et al. 2013) through electrostatic interaction and precipitation of heavy metals and the surface adsorption, partition, and sequestration of organic contaminants (Zhang et al. 2013). One new developed approach (consortium like amendments) of biochar with iron reported by Sneath et al. (2013) is that neither biochar treatment (1%, weight/weight) nor iron treatment could successfully reduce leaching of Cu and As, but single treatment of iron negatively impacted soil structure and plant mortality. However, the combined treatments of biochar and iron could reduce the Cu and As leaching and increased phenanthrene degradation and enabled plant growth, so it is reported that this is a useful practice for the treatment of co-contaminated mining sites (Sneath et al. 2013). Biochar was more effective than other composts at reducing Cd and Zn as well as the heavier metals, highly toxic likely PAHs (Beesley et al. 2010). Biochar could augment the bioremediation of PAHs in contaminated soil as microbial carriers of immobilized microorganism technique (Fig. 19.2). Nevertheless, it is important to select an appropriate category of biochar as an immobilized carrier of microorganism for the stimulation of biodegradation as well as bioremediation (Chen et al. 2012). But the degradation and remediation potential has been seen higher in two- to four-ring PAHs than fiveto six-ring in contaminated soil treated with biochar (Liu et al. 2015). According to Zhang et al. (2013), biochar addition to soil could stimulate PAH-metabolizing bacterial activity by enhancing the number of gene copies related to PAH degradation and changing the community structure of soil microbial community. However, it has been reported that the application of biochar in agricultural purpose can decrease the efficacy of pesticide, which directly indicates the negative effect on the industrial production of pesticides (Tang et al. 2013; Evangelou et al. 2015). So like this, many other controversial facts are accepted in published form; therefore, further research is needed.

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Table 19.2 Heavy metal removal by different types of biochar in contaminated soils Contaminants As and Cu

Biochar type Hardwood

Effects Mobilization due to enhanced pH and DOC

As, Cr, Cd, Cu, Ni, Pb, and Zn

Sewage sludge (500–550  C)

Cd and Zn

Hardwood

Immobilization of As, Cr, Ni, and Pb due to the rise in soil pH Mobilization of Cu, Zn, and Cd due to highly available concentrations in biochar Immobilization due to enhanced pH

Cd, Cu, and Pb

Chicken manure and green waste (550  C) Broiler litter (700  C)

Cu

Immobilization due to partitioning of metals from exchangeable phase to less bioavailable organic-bond fraction Cation exchange; electrostatic interaction; sorption on mineral ash content; complexation by surface functional groups Complexation with phosphorus and organic matter

Cu and Pb

Oakwood

Pb Pb

Dairy manure (450  C) Oakwood (400  C)

Immobilization by hydroxypyromorphite formation Immobilization by the rise in soil pH and adsorption on biochar

Pb

Rice straw

Nonelectrostatic adsorption

Pb, Cu, and Zn

Broiler litter (300 and 600  C)

Stabilization of Pd and Cu

Ni, Cu, Pb, and Cd

Cottonseed Hulls (200–800  C)

Surface functional groups of biocharcontrolled metal sequestration

References Beesley et al. (2010) Khan et al. (2004)

Beesley et al. (2010) Park et al. (2011) Uchimiya et al. (2011) Karami et al. (2011) Cao et al. (2011) Ahmad et al. (2012) Jianga et al. (2012) Uchimiya et al. (2012) Uchimiya et al. (2011)

Adapted from Hayyat et al.

7 Impact of Biochar on Arsenic Concentration and Mobility Biochar can reduce the concentration and mobility of arsenic (As) in amended soil in comparison with unamended soil. According to the previous study, an orchard residue-mediated biochar significantly increased As concentrations in pore water (500–2000 g L 1); however, in root and shoot, As concentrations in tomato (Solanum lycopersicum L.) plants were significantly reduced compared to the control (without biochar amendments) (Beesley et al. 2013). Moreover, Gregory et al. (2014) reported that in the increasing dose rates of biochar from willow feedstock

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(Salix sp.) application from 30 to 60 t ha 1, shoot tissue of ryegrass extracted significantly higher (P < 0:05) concentrations of As. The Al (III)-modified biochar, prepared from the rice straw, soybean residue, and peanut, had higher sorption capacity for As (V) under acidic conditions compared with unchanged biochar, increased with decreasing suspension pH, and could substitute Fe/Al oxides used for water purification under acidic conditions at pH > 4.0 (Qian et al. 2013).

8 Impact of Biochar on Degraded and Contaminated Land Land has vast majority of ecosystem on this planet, but the various anthropogenic activities are affecting the environmental balance; therefore, there is the need of some adequate policy related to environmental protection such as properly treated wastewater that should be discharged from different industries and proper management of daily discharge wastewater from household. In this concern, some biotechnological technique can be used in the management of environmental pollution like bioremediation and biodegradation as well as phytoremediation with the use of biochar as carrier to driving this process. The adequate restoration of degraded and contaminated land requires cooperation, integration, and assimilation of different biotechnological advances (Mani and Kumar 2014). Additionally, the recovery and restoration of degraded land could be by application of activated charcoal, pyrolyzed biomass, and organic fertilizers alone and combined amendments in affected land. Furthermore, it has been well supported that the use of biochar and fungi (arbuscular mycorrhizal) may facilitate grassland recovery in severely degraded habitats and the promotion of grassland ecosystem sustainability also (Ohsowski et al. 2012). Biochar amendment significantly increased plant roots’ length, enhanced root establishment in contaminated soils, and reduced Cu uptake to plants compared to the unamended soil (Brennan et al. 2014). Belyaeva and Haynes (2012) reported that biochar had a little or no stimulatory effect on the size of the soil microbial community composition, soil microbial biomass, N fertility, or plant growth during revegetation (Belyaeva and Haynes 2012). Biochar could adsorb soil Cd and Cu in industrial wastewater treatment but does not promote the growth of the wetland plant species (Juncus subsecundus) during early growth stage, maybe due to the negative interaction between biochar and waterlogged plants (Zhang et al. 2013). Therefore, further study is required to illuminate the proper mechanisms.

9 Influence of Biochar on Bioavailable PAH The addition of biochar in contamination soil could reduce the bioavailable fraction of all PAH groups, promoting the remediation in amended soils. Likewise, other carbon-rich materials (sorbents) like activated carbon have previously been shown to reduce the bioavailable fraction of PAHs. Brandli et al. (2008) found that the

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dissolved concentration of PAH was measured by using polyoxymethylene (POM) strips and observed lower concentration in activated carbon-amended contaminated soils. However, Yu et al. (2009) reported that the sorbing nature of biochar could reduce the bioavailability of various organic pollutants (pesticides) to plants. Though this reduction could be only in the bioavailable fraction, organic contaminants (pesticide) cannot persist for longer time in the presence of biochar or activated carbon (Table 19.3: depicting the impact of biochar on organic pollutants in contaminated soils). Because the fact related to the PAH concentration reduction related Table 19.3 Impact of biochar on adsorption of organic pollutants in contaminated soils Feedstock Eucalyptus wood

Production temperature 450  C and 850  C

Woodchip

500  C

Atrazine and acetochlor

Dairy manure

200  C and 350  C

Atrazine

Pine wood

350  C and 700  C

Terbuthylazine

Green wastes

450  C

Atrazine

Pine wood

350  C and 700  C

Phenanthrene

Eucalyptus woodchips

850  C

Diuron

Pine needles

100  C, 300  C, 400  C, and 700  C 250  C and 400  C

PAHs

Herbicides

Biochars showed high sorption ability for two herbicides, fluridone and norflurazon

Sun et al. (2012)

350  C and 700  C

Carbaryl

At low carbaryl concentrations, the sorption capacity of BC700 > BC350; similar sorption capacity at high carbaryl concentrations

Zhang et al. (2013)

Poultry litter, wheat straw, and swine manure Swine manure

Contaminant Diuron chlorpyrifos and carbofuran

Effect Higher pyrolysis temperature and higher rates of biochar applied to soils result in stronger adsorption and weaker desorption of pesticides Acetochlor adsorption increased by 1.5 times; atrazine adsorption also increased At 200  C, partitioning of atrazine is positively related to biochar carbon content Soil sorption increased 2.7- and 63-fold in the BC350 and BC700 treatments, respectively Biochar enhanced adsorption of pesticide Biochar produced at 700  C showed a greater ability at enhancing a soil’s sorption than that prepared at 350  C Pesticide absorption increases with the biochar contact time with soil and application rate Sorption capacity increased with pyrolysis temperature

References Yu et al. (2006)

Spokas et al. (2009) Cao et al. (2009) Wang et al. (2010) Zheng et al. (2010) Zheng et al. (2010) Yu et al. (2011) Chen and Yuan (2011)

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to biochar has been reported in some earlier study. Zhang et al. (2013) reported that the possibility of acetone/hexane extraction was not exhaustive enough to extract the PAHs that were more strongly bound to the added soil amendments (biochar), suggesting that these more strongly bound PAHs are not freely bioavailable and thus unlikely to pose an environmentally toxic risk. Otherwise, the application of biochar in contaminated soil could stimulate the biodegradation of PAH due to the enhancement of soil microbial activity in amended soils (Rhodes et al. 2008; Steinbeiss et al. 2009). According to Haritash and Kaushik (2009), the addition of compost alone or with the combined form of biochar could increase the degradation of PAH compounds as well as improve the soil texture and oxygen transfer and provide energy to the microbial population that can actively participate in this process. But Kaestner and Mahro (1996) found that the degradation process of PAH can be increased due to the biochar amendments, but sorption of organic matter in amended soil is unchanged. Biochar and compost (vermicompost and farmyard manure, FYM) both could reduce PAH concentrations by different mechanisms in the soil. Furthermore, when the blend of biochar and compost is added in combination, they generally worked less effectively than when added individually. Zhang et al. (2013) reported that the reduced PAH bioavailability caused by the addition of biochar may have an antagonistic relationship with the increased microbial activity caused by the addition of compost, by reducing the substrate available for the microbes.

10

Biochar and Immobilization of Contaminants in Soils

Biochar is an important solid sorbent that immobilizes organic and inorganic pollutants by various mechanisms in contaminated soil (Chen and Chen 2009; Ahmad et al. 2014). According to the previous study, electrostatic attraction, polar and nonpolar organic attraction to the carbonized stage of biochar, and partitioning to the non-carbonized phase enhanced the attachment of biochar with organic contaminants (Chen et al. 2008; Chen and Chen 2009; Huang and Chen 2010). Soil contaminants are attached on biochar surface by the proficient mechanism of ion exchange, anionic metal attraction, precipitation, and cationic metal attraction (Cao et al. 2009; Qian and Chen 2013; Ahmad et al. 2014, Xu and Chen 2015). Cao et al. (2009) reported that in some cases, biochar can be even more efficient than activated charcoal to immobilize heavy metals; in a study, six times more Pb absorption was detected in biochar-amended soil. Biochar could adsorb some herbicides (such as simazine) in micropores or on the surface and prevents their accessibility to microbial cells, extracellular enzymes, and plants as well as prevents their leaching into the groundwater (Jones et al. 2011). After the changes in biochar properties with the treatment of iron oxide obtained such magnetic biochar with an efficient sorption capability of organic and inorganic pollutants, which enables biochar a multifunctional application in agricultural and environmental fields (Chen and Yuan 2011; Qian et al. 2013), current research exposes the significant role of biochar

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in the reduction of aluminum (Al) toxicity. Biochar surface could be oxidized during aging process and obtain more heavy metal binding sites such as Cd and Al; therefore, the rivalry sorption between Cd and Al was seen in previous studies (Qian et al. 2015). Silicon plays a significant role in biochar and provides strength which reduced the Al phytotoxicity and protects the plant from Al absorption, so that Si particles can reduce the amount of soil exchangeable Al and that Si released from biochar can form Si-Al complex in the upper surface of plant tissue and consequently reduced the risk of exchangeable Al uptake in the inner tissue of plant (Qian et al. 2016). However, above application showing the multifunctional role of biochar in agricultural and environmental research rather than the complete underlying mechanism is further needed to investigate the biogeochemical application in agroecosystem.

11

Conclusions and Future Research Direction

Biochar is a strong solid sorbent that can be used to develop new viable technology in highly challenging research field bioremediation and can reduce the bioavailability of organic and inorganic pollutants in contaminated soils. Biochar generally produced at higher temperature (pyrolysis) in limited supply of oxygen contains severalfold higher carbon layers and diverse physicochemical properties which can affect the effectiveness in the remediation of pollutants in contaminated soils. The distinct studies have been reported that the application of biochar in contaminated soil not only can reduce the bioavailability of pollutants but also may significantly lead and enhance the rate of diffusion in soils. The auxiliary research requires to explore the experimental viability of biochar-assisted diffusion of several PAHs and heavy metals in affected sites. According to Zhang et al. (2013), the immobilized microorganism technique (IMT) with biochar as supportive and carrier medium shows imaginary experimental demonstration for the removal of soil contaminants in Fig. 19.2. Moreover, the small amendments of biochar in soil can significantly reduce the uptake efficacy of pesticides and other organic and inorganic pollutants in plant’s tissue (Zhang et al. 2013). Various other studies also support the multifunctional application of biochar in diverse areas of agriculture and environment such as the removal of several herbicides and hydrocarbons as well as improving soil microbial biomass (Singh and Gupta 2018) and agricultural production. The different feedstock-based biochar is a potential technology for remediation of several pollutants in agricultural contaminated soils. However, some new emerging approaches have to develop in the field of bioremediation for prevention of various environmental risks because every year the graphical bar of environmental pollution is growing in an upward direction due to the distinct indiscriminate anthropogenic activity such as industrialization, urbanization, direct discharging of households, and improper agricultural waste management practices that are serious concerns to all living beings.

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Acknowledgments The authors express their sincere thanks to the head of the Department of Environmental Microbiology for providing infrastructure facilities and the University Grant Commission (UGC), New Delhi, for the financial support provided.

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Singh JS (2015a) Biodiversity: current perspective. Clim Change Environ Sustain 3(1):71–72 Singh JS (2015b) Microbes: the chief ecological engineers in reinstating equilibrium in degraded ecosystems. Agric Ecosyst Environ 203:80–82 Singh JS (2015c) Biodiversity: current perspectives. Clim Change Environ Sustain 2:133–137 Singh JS (2015d) Plant-microbe interactions: a viable tool for agricultural sustainability. Appl Soil Ecol 92:45–46 Singh JS (2016) Microbes play major roles in ecosystem services. Clim Change Environ Sustain 3:163–167 Singh BP, Cowie AL (2008) A novel approach, using 13C natural abundance, for measuring decomposition of biochars in soil. In: Currie LD, Yates LJ (eds) Carbon and nutrient management in agriculture. Occasional report no. 21. Fertilizer and Lime Research Centre, Massey University, Palmerston North, p 549 Singh JS, Gupta VK (2018) Soil microbial biomass: a key soil driver in management of ecosystem functioning. Sci Total Environ 634:497–500 Singh JS, Pandey VC (2013) Fly ash application in nutrient poor agriculture soils: impact on methanotrophs population dynamics and paddy yields. Ecotoxicol Environ Saf 89:43–51 Singh JS, Strong PJ (2016) Biologically derived fertilizer: a multifaceted bio-tool in methane mitigation. Ecotoxicol Environ Saf 124:267–276 Singh B, Singh BP, Cowie AL (2010) Characterisation and evaluation of biochars for their application as a soil amendment. Aust J Soil Res 48:516–525 Singh JS, Abhilash PC, Gupta VK (2016a) Agriculturally important microbes in sustainable food production. Trends Biotechnol 34:773–775 Singh C, Chowdhary P, Singh JS, Chandra R (2016b) Department of pulp and paper mill wastewater and coliform as health hazards: a review. Microbiol Res Int 4:28–39 Singh C, Tiwari S, Singh JS (2017a) Impact of rice husk biochar on nitrogen mineralization and methanotrophs community dynamics in paddy soil. Int J Pure App Biosci 5:428–435 Singh C, Tiwari S, Singh JS (2017b) Application of biochar in soil fertility and environmental management: a review bull. Environ Pharmacol Life Sci 6:07–14 Singh C, Tiwari S, Boudh S, Singh JS (2017c) Biochar application in management of paddy crop production and methane mitigation. In: Singh JS, Seneviratne G (eds) Agro-environmental sustainability (Managing environmental pollution). Springer, Cham, pp 123–146 Sneath HE, Hutchings TR, de Leij FAAM (2013) Assessment of biochar and iron filing amendments for the remediation of a metal, arsenic and phenanthrene co-contaminated soil. Environ Pollut 178:361–366 Spokas KA, Koskinen WC, Baker JM, Reicosky DC (2009) Impacts of woodchip biochar additions on greenhouse gas production and sorption/degradation of two herbicides in a Minnesota soil. Chemosphere 77:574–581 Steinbeiss S, Gleixner G, Antonietti M (2009) Effect of biochar amendment on soil carbon balance and soil microbial activity. Soil Biol Biochem 41:1301–1310 Steiner C, Teixeira WG, Lehmann J, Nehls T, Luis Vasconcelos de Macêdo J, Blum WEH, Zech W (2007) Long term effects of manure, charcoal and mineral fertilization on crop production and fertility on a highly weathered Central Amazonian upland soil. Plant Soil 291:275–290 Sun K, Gao B, Ro KS, Novak JM, Wang ZY, Herbert S, Xing BS (2012) Assessment of herbicide sorption by biochars and organic matter associated with soil and sediment. Environ Pollut 163:167–173 Tang J, Zhu W, Kookana R, Katayama A (2013) Characteristics of biochar and its application in remediation of contaminated soil. J Biosci Bioeng 116:653–659 Thomas SC, Frye S, Gale N, Garmon M, Launchbury R, Machado N, Melamed S, Murray J, Petroff A, Winsborough C (2013) Biochar mitigates negative effects of salt additions on two herbaceous plant species. J Environ Manag 129:62–68 Uchimiya M, Wartelle LH, Klasson T, Fortier CA, Lima IM (2011) Influence of pyrolysis temperature on biochar property and function as a heavy metal sorbent in soil. J Agric Food Chem 59:2501–2510

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Chapter 20

Bioremediation of Melanoidins Containing Distillery Waste for Environmental Safety Vineet Kumar and Ram Chandra

Abstract Distillery waste (DW) is a major threat to the environment for its safe disposal due to their high content of various toxic inorganic and organic compounds. The degradation and detoxification of color-contributing compounds such as caramel, melanoidins, and plant phenolics and their metabolic products are essential prior to disposal of DW into the environment. Distilleries employ different forms of primary and secondary processes for effluent treatment; however, these treatment methods are highly energy intensive and hence quite expensive. Biological methods present an incredible alternate for decolorization and detoxification of DW due to their environmentally friendly, low cost, and publicly acceptable treatment processes. A wide variety of aerobic microorganisms including bacteria, fungi, actinomycetes, and cyanobacteria have the ability to decolorize melanoidins containing DW. But DW containing different types of pollutants are not easily degraded by the single-step treatment process. Up to the present, however, no suitable method for the promising treatment of huge amounts of DW has been developed. The present book chapter aims to provide a comprehensive overview of some of the promising bioremediation and phytoremediation approaches used for the management of DW. Further, the use of two-step sequential approach for the treatment of DW is highlighted. Furthermore, the challenges and future prospects of bioremediation of DW are also discussed. Keywords Distillery sludge · Heavy metals · Androgenic-mutagenic compounds · In situ phytoremediation · Vermicomposting

V. Kumar (*) · R. Chandra (*) Department of Environmental Microbiology (DEM), School for Environmental Sciences (SES), Babasaheb Bhimrao Ambedkar University (A Central University), Lucknow, Uttar Pradesh, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_20

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1 Introduction Industrial waste management is the major thrust area of research in the world, as these wastes contain a variety of organic and inorganic pollutants that cause serious environmental pollution and toxicity in living beings (Goutam et al. 2018; Bharagava et al. 2017a, b; Chandra and Kumar 2015a). The discharge of waste (solid and liquid) into the environment from distilleries is a threat to the ecosystem due to the big environmental hassle (Chandra and Kumar 2017a, b, Chowdhary et al. 2017). In India, distilleries are one of the major agro-based polluting industries; in addition, they are a high consumer of raw water, utilizing the molasses from cane sugar manufacturing (sugar mills) as the starting material for ethanol production. Molasses (a by-product of the sugar industry) contains roughly 50% fermentable sugars (glucose, fructose, and sucrose). In distilleries, during the ethanol production, a major portion (approximately 90%) of the fermented wash going to distillation column is discharged as wastewater named vinasses or spent wash (SW) or raw effluent (Chandra and Kumar 2017b). The characteristic of the SW varies and depends on the fermentation feedstock and fermentation and distillation processes adopted in distilleries. The molasses-based distilleries are mainly located in subtropical and tropical regions of the world, and these generate approximately 12–15 liters of SW per liter of ethanol production (Chandra et al. 2018a; Chandra and Kumar 2017b). The SW is characterized with unpleasant odor; dark-brown color; high level of total dissolved solids (TDS); chemical oxygen demand (COD); biological oxygen demand (BOD); total Kjeldahl nitrogen (TKN), sulfide (SO32
), sodium (Na+), and phosphate (PO43
); and the presence of various toxic heavy metals (HMs), viz., iron (Fe), nickel (Ni), copper (Cu), zinc (Zn), lead (Pb), and manganese (Mn), and nitrogen-containing complex organic pollutants and endocrine-disrupting chemicals (EDCs) with low pH (Kaushik and Thakur 2013; Chandra and Kumar 2017b). The origin of dark-brown color in SW is primarily associated with the low and high molecular weight (mw) melanoidin pigments, a class of compounds known as Maillard reaction products (MRPs) (Arimi et al. 2014). Currently, in India, there are 397 distilleries that are producing approximately 3.25  1010 liters of ethanol and generating approximately 40.90  1015 liters of SW per annum (Kumar and Chandra 2018). Due to high pollution nature of SW, the Ministry of Environment, Forest, and Climate Change, Govt. of India, has listed ethanol industries at the top among the “Red category” industries (Tewari et al. 2007). Regarding water and soil pollution, the Bureau of Indian Standard (BIS) provides guidelines to the central and state government authorities which would help to decide suitable restrictions on effluent disposal and to the industry for selecting a suitable site and suitable treatment technology for effluents prior to their disposal into the environment. Consequently, distillery effluent must be properly treated before it is discharged into the environment (Tewari et al. 2007). In industry, distillery SW treatment is being carried out commonly by four routes: (i) concentrate the SW followed by incineration, (ii) direct oxidation of SW by air (oxygen) at a high temperature followed by aerobic treatment, (iii) anaerobic treatment (methanogenesis) of SW with methane (biogas)

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recovery followed by aerobic polishing, and (iv) reverse osmosis. Out of the four routes, anaerobic treatment has been considered as the most attractive first-step treatment technique for the SW due to its reputation as an eco-friendly and low-cost technique besides its methane production potential (Oller et al. 2011). In addition, anaerobic treatment is preferred due to fact that a significant component of SW is biodegradable. The anaerobic treatment is reported to remove about 40–50% COD and 60–65% BOD from the SW (Arimi et al. 2015). The color of SW converts darker with higher TDS after anaerobic digestion due to complexation of organic and inorganic pollutants present in the SW. This means that SW after anaerobic treatment still contains high organic and inorganic load and it is not safe for discharge into the environment (Chandra and Kumar 2017a; Chandra et al. 2018a). The SW received after anaerobic treatment (methanogenesis) is called postmethanated distillery effluent (PMDE) or biomethanated distillery effluent (BMDE). PMDE is characterized by an extremely higher level of BOD, COD, TDS, TKN, SO32
, Na+, and PO43
, with alkaline pH and dark-brown color. In addition, PMDE retains a high amount of various toxic HMs, viz., Ni, Mn, Fe, Zn, Cu, Pb, and Cd along with melanoidins and various recalcitrant/refractory organic pollutants (ROPs) (Bharagava and Chandra 2010a; Chandra et al. 2012). Moreover, Indian distilleries produce approximately 1500 tons of distillery sludge per day during anaerobic treatment of SW (Kansal et al. 1998) which is characterized by high organic matter (OM), phenolics, SO32
, Na+, PO43
, and HMs along with melanoidins and various androgenic and mutagenic compounds (Chandra et al. 2018b; Chandra and Kumar 2017a). Satyawali and Balakrishnan (2008) reported that anaerobic treatment does not result in effective color removal from SW. When the untreated or partially treated melanoidins containing SW or PMDE are released into water bodies, they damage aquatic ecosystem by reducing the sunlight penetration power and finally the photosynthetic activities and dissolved oxygen content, whereas in agricultural soil it causes depletion of vegetation by reducing the soil alkalinity and Mn availability and inhibiting the seed germination (Chowdhary et al. 2017) (Fig. 20.1). Therefore, it is essential to study additional pre- or post-treatment methods to remove color and recalcitrant organic compounds from PMDE. Fig. 20.1 shows the environmental pollution caused by discharged SW and PMDE. The complexity of distillery effluent makes not easy for the development of efficient tertiary treatment technologies for removal of color and toxic androgenic and mutagenic compounds prior to its disposal into the environment. For instance, Indian distilleries have been directed by the Ministry of Environment, Forest, and Climate Change, Govt. of India, to achieve zero liquid discharge into the surface water bodies from December 2005. In many regions of India, distillery industries dilute the PMDE by mixing with raw water prior to discharge in the environment in order to meet the set effluent disposal standard. This dilution of PMDE even though accepted in some regions of India is of great environmental concern as it does not reduce the complete pollution load of the PMDE (Chaudhari et al. 2007). Various physicochemical methods such as membrane filtration (Kumar et al. 2008), flocculation (Liang et al. 2009a), coagulation (Liang et al. 2009b), adsorption (Onyango et al. 2011), ozone oxidation (Kim et al. 1985), and ultraviolet/hydrogen peroxide

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Fig. 20.1 Distillery wastes and their environmental impacts: (a) spent wash, (b) spent wash after anaerobic digestion, known as PMDE, (c) anaerobically digested distillery sludge, (d) leaching of PMDE in soil, (e) aquatic pollution (f) underground water pollution

(UV/H2O2) treatment (Dwyer and Lant 2008) have been used for elimination of melanoidins and decolorization of SW and/or PMDE, but these treatment techniques are not practicable at big scale due to the elevated cost, blockage of filtration column, and generation of huge amount of toxic sludge and other secondary pollutants (Liang et al. 2009b). Therefore, adequate treatment of SW and/or PMDE in an eco-friendly manner is necessary prior to its disposal into the environment (Chowdhary et al. 2017). Interest in the microbial degradation and detoxification of distillery waste has intensified in recent years as mankind strives to find sustainable ways to cleanup and restore polluted sites. Due to the relatively low cost and the variations of work progress, the biological methods have been most widely used all over the world. Bioremediation is an eco-friendly technique that employs many different microbes acting in parallel or sequentially to degrade and detoxify toxic contaminants. It is defined as the acceleration of the natural metabolic process whereby microorganisms (i.e., bacteria, actinomycetes, and fungi), green plant (termed phytoremediation), or their enzymes degrade or transform toxic contaminants to CO2 (carbon dioxide), H2O (water), microbial biomass, inorganic salts, and other by-products (metabolites) that may be less toxic than the parent compounds (USEPA 2006, 2012a). Thus, the use of microbes for degradation and detoxification of pollutants is now being increasingly applied as the technology of choice for cleanup or restores contaminated sites to environmentally sustainable (Megharaj et al. 2011). The aim of this book chapter is to provide a concise discussion of the processes associated with the use of biological agents to remediate hazardous complex distillery waste. In this chapter, we describe phytoextractions strategy of the plant with

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special emphasis on removal of heavy metals from EDCs reaches complex distillery waste. We also discuss the challenges for the elimination of inorganic and organic compounds from distillery waste as well as contaminated sites, and we provide some information on costs. Furthermore, the use of a two-step sequential treatment approach for the treatment of distillery effluent is highlighted.

2 Distillery Waste and Its Physicochemical Characteristics Currently, safe disposal of distillery waste into the environment is of paramount importance due to the existence of toxic complex inorganic and organic compounds (Chandra and Kumar 2017a, b). Distillery effluent has a very high organic load, high TDS, unpleasant odor, and dark-brown color. The main components of these complex effluent are carbohydrates, amino acids, lipids, proteins, fatty acids, melanoidins and sometimes lignin, and their fragmentary products, while the inorganic constituents include large concentrations of Na+, NO32
, S2
, chloride (Cl-), potassium (K), phosphorus (P), ammonium salts, and toxic HMs, viz., Zn, Fe, Mn, Cu, Ni, Pb, and Cd (Chandra et al. 2012; Chaturvedi et al. 2006). Distillery waste is broadly divided into two categories: (i) distillery effluent and (ii) distillery sludge.

2.1

Distillery Effluent

Distillery effluent is characterized by high level of BOD, COD, TDS, and phenolics, with strong odor, alkaline pH, and dark-brown color. It also contains a high amount of N, K, and P which can lead to eutrophication of aquatic ecosystem (Chowdhary et al. 2017; Sowmeyan and Swaminathan 2008). In addition, distillery effluent also retains a huge amount of Fe, Cu, Pb, Zn, Mn, Ni, and Cd (Bharagava and Chandra 2010a; Tiwari et al. 2013). The refractory nature of distillery effluent is due to the existence of various MW melanoidins, caramel, polyphenols, and a variety of sugar decomposition products such as anthocyanin and tannins and diverse ROPs including EDCs (Chandra and Kumar 2017b). The physicochemical characteristics of SW and PMDE are shown in Table 20.1.

2.2

Distillery Sludge

Distillery industry produces a huge quantity of sludge after anaerobic digestion, whose management and disposal are environmental problems due to their polluting characteristics. The generated sludge is enriched with various androgenic and mutagenic compounds and other potentially toxic compounds such as melanoidins, EDCs, resin acids, fatty acids, and HMs (Chandra and Kumar 2017a, Chandra et al.

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Table 20.1 Physicochemical characteristics of melanoidins containing distillery waste S. No. 1. 2. 3. 4. 5. 6. 7. 8. 9. 11. 12. 13. 14. a b c d e f g h

Parameter Color intensity pH EC BOD COD TS TDS Chloride Total nitrogen Phenol Sulfate Phosphate Trace Elements Mn Cr Zn Cu Fe Pb Cd Ni

Distillery waste Spent wash 150000 4.07 – 42000  1848 90000  3600 103084  5150 77776  3768 2200  105 2800  130 4.20  1.8 5760  260 5.36  0.168 4.55  0.159 1.05  0.031 2.48  0.094 0.33  0.016 163.94  6.52 BDL BDL 1.17  0.047

PMDE 80,000  2127 8.5  0.17 – 9800  160 18700  321 47422  336 17,612  284 4096  218 6893  147 2786  112 1625  108

Distillery sludge – 8.1  0.00 4.12  0.01 – – – – 1272.74  5.13 – 501.34  1.22 145.07  0.68 2268.83  1.70

43.63  3.41 – 1.24  0.20 0.75  0.21 57.50  3.90 0.23  0.04 0.13  0.10 0.31  0.03

238.47  0.83 – 43.47  1.31 847.46  1.00 5264.49  59.64 31.22  1.14 – 15.60  0.54

All values are mean (n ¼ 3)  SD in mg kg
1/mg l
1 except color intensity (Pt-Co unit); pH; EC (μs cm
1); BOD biological oxygen demand, COD chemical oxygen demand, TS total solids, TDS total dissolved solids, VS volatile solids, BDL below detection limit Chandra and Kumar (2017b), Chandra et al. (2012), Chandra et al. (2018b), Bharagava and Chandra (2010a, b)

2018b). Approximately 1500 tons of anaerobically digested distillery sludge is being wasted per day in India (Kansal et al. 1998). On disposal of distillery sludge in the aquatic and terrestrial ecosystem, it upsets natural food chains, immobilizes soil nutrients, depresses microbial activity, and affects microflora and macrofauna as well as human health. However, in nature, this toxic distillery sludge is slowly converted into a soil amendment (compost) by microflora and/or macrofauna; therefore, their use singly or in combination is popular nowadays for rapid recycling of nutrients of distillery waste. The sludge environments could be effective incubators of a wide variety of resistant and adapted bacterial communities and should offer an ideal source of potential ligninolytic enzyme producing potential bacteria suitable for related industrial applications (Kumar and Chandra 2018; Chandra and Kumar 2017a). Consequently, investigations of distillery sludge bacterial communities and its organic compounds have been reported.

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3 Organic Pollutants of Distillery Waste The origin of high COD, BOD, and color in distillery effluent is due to the existence of tannin- and lignin-like structural compounds, for example, melanoidins, crosslinked polyphenols, and EDCs which are resistant to physical, chemical, and biological degradation. These recalcitrant organic compounds present in distillery effluent are generated during the distillation process of fermented molasses slurry and the subsequent methanogenesis of the SW.

3.1

Melanoidins

Melanoidins, the major nitrogenous color imparting organic compounds in distillery effluent, formed through the Maillard reaction (MR) or nonenzymatic browning reaction occur between amino acids and carbohydrates at temperatures above 50  C and pH 4.0 to 7.0 (Hayase 2000). Melanoidins with an MW distribution between 5 and 40 kDa consist of polymeric, acidic, negatively charged, and highly dispersed colloids. The negative charge of melanoidins is due to the dissociation of carboxylic and phenolic groups (Moreira et al. 2012). The MR is a complex series of reactions; the first stage involves the initial condensation of the sugar with amino acid and the isomerization of the resulting products to form Heyns or Amadori products. Amadori product and its dicarbonyl derivatives can undergo concurrently retro-aldol reactions producing more reactive C2-C5 sugar fragments, such as diketones, glyceraldehydes, and hydroxyacetone derivatives. This chemical reaction is called as Strecker degradation, and it is characterized by the production of CO2 (Arimi et al. 2015). The aldehydes formed may be important as auxiliary flavor compounds, and they also contribute to melanoidin formation. The formation of melanoidins from sugar and amino acid reactions is illustrated in Fig. 20.2a. Due to the recalcitrant antioxidant nature of melanoidins, it inhibits microbial growth used in biological treatment processes of distillery effluent. Consequently, distillery effluent is a major source of aquatic and terrestrial pollution in the environment. Since, melanoidins are generally regarded to be heterogenous, negatively charged, and highly dispersed polymer with similar chemical properties to humic substances (Migo et al. 1993). It has been reported that due to their net negative charge, various HMs such as Cu2+, Cr3+, Fe3+, Zn2+, Pb2+, etc. bind with melanoidins to form an organometallic complex (Hatano et al. 2016; Migo et al. 1997). Consequently, the high HM-binding tendency of melanoidins also enhances the vulnerability of organometallic complex toward its toxicity in the ecosystem. In sugar processing, melanoidins are formed during purification and evaporation steps. The elemental and chemical structure of melanoidins depends heavily on the nature and molecular concentration of reactant and reaction conditions such as pH, heating time temperature, and solvent. Cammerer et al. (2002) have proposed the

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a

Amino Compounds

Aldose sugar

b

Glycosamine+ H2O

Amadori rearrangement products (ARPs) 1-amino-1-deoxy-2-ketose -2 H2O

-3 H2O Schiff base of hydroxymethyl furfural (HMF)

Reductones -2 H2O

-Amino compounds -2 H2O

Dehydroreductones

Polymerisation + Amino acids

HMF or furfural

Melanoidin

Polymerisation + Amino acids

Fig. 20.2 The formation of advanced MRPs (melanoidin): (a) the basic structure of melanoidin proposed by Cammerer et al. (2002)

fundamental structure of melanoidins pigment formed from Amadori reaction products and 3-deoxyhexosuloses as shown in Fig. 20.2b.

3.2

Polyphenols

Distilleries produce a huge volume of effluent along with sludge containing phenolic compounds. Polyphenols are categorized into three major classes: phenolic acid, flavonoids, and tannins. Phenolic acids have been detected in distillery effluent, including cinnamic acid and its derivatives (caffeic acid, p-coumaric acid, ferulic acid, and chlorogenic acid) and benzoic acid and its derivatives (e.g., gentisic acid, gallic acid), (Payet et al. 2005, 2006), which give a high antimicrobial activity to this effluent, thus slowing down the aerobic and anaerobic treatment process (Borja et al. 1993). A number of organic compounds containing these highly reactive phenolic groups are generated and extracted into the distillery effluent and responsible for its color.

3.3

Endocrine-Disrupting Chemicals (EDCs) and Other Refractory Organic Pollutants

The effluent discharged from distilleries can be a significant source of EDCs and refractory organic pollutants (ROPs) to the environment (Chandra and Kumar

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2017b). EDCs are class of chemicals which have exogenous and xenobiotic origins while inhibiting or mimicking the natural activities (secretion, synthesis, binding, and transport) of the endocrine system which play a significant role in sexual development, metabolism, hormone production, and their utilization in growth, gender behavior, reproduction processes, and stress response in humans and animals (Kabir et al. 2015; Somm et al. 2009). The existence of EDCs such as 2-hydroxysocaproic acid; butanedioic acid, bis(TMS)ester; and vanillylpropionic acid, bis(TMS), benzenepropanoic acid, α-[(TMS)oxy], TMS ester, and ROPs such as benzoic acid 3-methoxy-4-[(TMS)oxy], TMS ester; 2-furancarboxylic acid, 5-[[(TMS)oxy] methyl], TMS ester; and tricarballylic acid 3TMS in SW has been reported by Chandra and Kumar (2017b). For the characterization of EDCs and ROPs from SW, they employed several organic solvents (i.e., acetone, ethyl acetate, isopropanol, methanol, ethanol, and n-hexane) to extract the broad range of organic pollutants under acidic conditions (95%), phosphate (77%), and sulfate (35%).

6 Phytoremediation of Distillery Waste Phytoremediation is a long-lasting, cost-effective, and aesthetic solution for remediation of hazardous pollutants from complex distillery waste (Chandra et al. 2018c; Chandra and Kumar 2017c; 2018). Phytoremediation is based on the use of green plants to extract, sequester, and detoxify pollutants present in distillery effluent (Chandra et al. 2015; Chandra and Kumar 2018). Phytoextraction is the most commonly used technique of phytoremediation; it involves the utilization of hyperaccumulator and non-hyperaccumulator plants for accumulation, absorption, and degradation of pollutants from the soil, sediment, or water. Phytoremediation technology can be divided into two groups based on the physical location of the remedial action: (i) in situ bioremediation and (ii) ex situ bioremediation.

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In Situ Phytoremediation

In situ phytoremediation involves placement of green plants in contaminated soil or sediment or in soil or sediment that is in contact with contaminated groundwater for the purpose of remediation. In this approach, the contaminated material is not removed prior to phytoremediation. Several native plants that grown on distillery waste-contaminated sites under natural environment have indicated the phytoremediation potential as natural hyperaccumulators of HMs from complex organic wastes (Fig. 20.3), because these plants are naturally adapted in terms of growth, survival, and reproduction under the environmental stresses, compared to plants introduced from another environment. Chandra and Kumar (2017c) evaluated the phytoextraction potential of 15 potential native plants, namely, Datura stramonium, Achyranthes sp., Kalanchoe pinnata, Trichosanthes dioica, Parthenium hysterophorus, Cannabis sativa, Amaranthus spinosus L., Croton bonplandianum, Solanum nigrum, Ricinus communis, Saccharum munja, Basella alba, Setaria viridis, Chenopodium album, and Blumea lacera grown on stabilized distillery sludge. The HM accumulation from distillery sludge and the distribution of these metals in roots, shoots, and leaves followed variable patterns. They observed that C. bonplandianum, S. munja, D. stramonium, Achyranthes sp., A. spinosus L., T. dioica, B. alba, C. sativa, C. album, S. viridis, P. hysterophorus, and B. lacera have been noted as root accumulators for Fe, Zn, and Mn, while Achyranthes sp., K. pinnata, B. alba, B. lacera, D. stramonium, T. dioica, C. album, C. sativa, P. hysterophorus, and S. munja are shoot accumulators of Fe. In addition, A. spinosus L. was found to be a shoot accumulator for Mn and Zn. Similarly, all plants were determined to be leaf accumulators of Mn, Zn, and Fe, except for A. spinosus L. and R. communis. Further, the bioconcentration factor (BCF) of all native plants was found 1. This gives a strong evidence of hyperaccumulation for the tested metals from complex distillery waste. Furthermore, the transmission electron microscopic (TEM) observations of the root of R. communis, Solanum nigrum, C. sativa, and P. hysterophorus showed the formation of multi-vacuoles and multi-nucleolus and deposition of HM granules in cellular component of roots as a plant adaptation mechanism for phytoextraction of HM-rich polluted site. This study recommended that these native plants be used as tools for in situ phytoremediation and eco-restoration of industrial wastecontaminated sites. Another study conducted by Chandra et al. (2018b) showed that distillery sludge not only contains the mixture of complex organic pollutants but also retains high quantity of Pb (31.22), Ni (15.60), Fe (5264.49), Mn (238.47), Cu (847.46), and Zn (43.47 mg kg
1) which enhances the toxicity of sludge to the environment. The major identified organic compounds were benzoic acid, 3,4,5-tris (TMS oxy), TMS ester; stigmasterol TMS ether; hexanedioic acid, dioctyl ester; benzene, 1-ethyl-2-methyl; 5α-cholestane, 4-methylene; campesterol TMS; and β-sitosterol and lanosterol. These compounds are listed under the EDCs also as per USEPA (2012b). Authors collected nine representative native plant species (grasses and weeds), viz., B. alba, Calotropis procera, Tinospora cordifolia, Rumex

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Fig. 20.3 Native plants luxuriantly grown on stabilized post-methanated distillery sludge: (a–b) Ricinus communis, (c) Cannabis sativa, (d) Parthenium hysterophorus, (e) Achyranthes sp., (f) Rumex dentatus, (g) Saccharum munja, (h) Argemone mexicana, (i) Cynodon dactylon, (j) Blumea lacera, (k) Solanum nigrum, (l) Chenopodium album, (m) Achyranthes sp., (n) Croton bonplandianum, (o) collection of fresh distillery sludge discharged after biomethanation of spent wash

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dentatus, C. album, Pennisetum purpureum, Cynodon dactylon, S. munja, and Argemone mexicana, which were abundantly grown on disposed distillery sludge. Further, the phytoextraction potential of grown native grasses and weeds revealed the high accumulation and translocation of Pb, Ni, Mn, Cu, Zn, and Fe in their leaves and roots compared to shoot. The BCF and TF of B. alba, C. procera, T. cordifolia, R. dentatus, C. album, P. purpureum, C. dactylon, S. munja, and A. mexicana for Ni, Cu, and Zn were found >1, while the TF of most of the grown plant for Pb, Ni, Mn, Cu, and Zn was noted >1. All the grown plants have been found for hyperaccumulation properties of HMs from organometallic polluted site mixed with androgenic and mutagenic compounds. The observation of root tissues of all plant species through TEM analysis showed the formation of multi-vacuoles, multinucleolus, multi-nucleus, and mitochondria and opaque deposition of HMs granules in the cellular organelle of the plant cell which indicated the HMs tolerance mechanism of the plant at an elevated concentration of HMs and other complex co-pollutants. At an elevated level of HMs, the production of a larger number of vacuoles and nucleolus increases the production of mRNA and ribosome, which ultimately increase the production of new proteins being involved in the HM tolerance in the plant. Moreover, the larger number of mitochondria formation indicated the production of the higher amount of energy in the form of ATP inside the cell, which is required in order to combat the HM toxicity. The authors recommended that these native plants can be used for in situ phytoextraction of HMs from industrial waste polluted sites.

6.2

Ex Situ Phytoremediation

For polluted sites where the pollutants are not bioavailable to plants, such as pollutants present in deep aquifers, an alternative method of remediation applying ex situ phytoremediation is possible. In this approach, the contaminated media are removed from the actual site using mechanical means and then transferred to a temporary treatment area where they can be exposed to plants selected for phytoremediation. After treatment (phytoremediation), the cleaned soil or water can be transferred to the original site, and the plant may be harvested for disposal if necessary. Phytoremediation of melanoidins and HMs containing distillery waste is an eco-friendly approach, and to date, only a restricted number of plant species have been explored for the degradation of melanoidins, for example, Phragmites karka, P. australis, P. communis, and Typha angustifolia (Chandra and Yadav 2010, Chandra et al. 2012; Chandra et al. 2018d; Chaturvedi et al. 2006). For the removal of distillery effluent contaminants, there is some significant work that has been done by Billore et al. (2001) for a horizontal flow gravel bed constructed wetland (CW) to treat conventional secondary treatment distillery effluent. After secondary treatment, the concentrations of COD and BOD5 in distillery effluent amounted to 2540 and 13,866 mg L
1, respectively, and therefore, additional treatment was essential. The CW treatment system achieved BOD5, COD, total P (TP), and total Kjeldahl

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nitrogen (TKN) reductions up to 84%, 64%, 79%, and 59%. This study recommended that CW may be a sustainable tertiary treatment technique for the remediation of contaminants present in distillery effluent. In another study, Trivedy and Nakate (2000) used wetland plant T. latifolia for treatment of distillery effluent in a CW treatment system. This treatment system resulted in 47% and 78% decrease in BOD and COD, respectively, in an incubation of 10 days. Increasing concentration of distillery effluent significantly reduced the biomass of growing plant with the highest accumulation of Fe being recorded in plants growing in 100% concentration of effluent. Aquatic macrophyte Potamogeton pectinatus was used to accumulate Mn, Zn, Cu, and Fe and efficiently clear out the distillery effluent (Singh et al. 2005). Chaturvedi et al. (2006) reported the phytoremediation potential of P. australis grown on distillery effluent contaminated site. She also characterized the diverse bacterial species from the rhizospheric zone of P. australis. The culturable bacterial species were helpful for the degradation and detoxification of noxious pollutants that exist in the distillery effluent. They observed 75.5% reduction of color by the same bacterial species along with a concomitant reduction in BOD, COD, sulfate, phenol, and HM values. Bharagava et al. (2008) study the HM accumulation efficiency and its physiological effects in Brassica nigra L. (Indian mustard) plants grown in soil irrigated with different concentrations (25%, 50%, 75%, 100%, v/v) of PMDE after 30, 60, and 90 days treatment. This study concluded that B. nigra L. accumulate elevated concentration of Zn, Ni, Mn, Fe, Cu, and Cd due to increased amount of cysteine and ascorbic acid (work as antioxidants) in leaves, shoot, and root of B. nigra L. at all the concentration and exposure periods of PMDE except at 90 days period, where a decrease was observed at 100% PMDE concentration as compared to their respective control. Chandra and Yadav (2010) conducted a pot culture experiment to evaluate the accumulation pattern of Cu, Pb, Ni, Fe, Mn, and Zn in T. angustifolia grown in Zn-, Mn-, Fe-, Ni-, Pb-, and Cu-rich aqueous solutions of phenols and melanoidins. They concluded that T. angustifolia could be an efficient phytoremediator for HMs from melanoidin-, phenol-, and metal-containing industrial effluent at optimized conditions. Recently, Hatano et al. (2016) observed the chelating property of melanoidin like the product (MLP) and to assess the facilitatory influence on the phytoextraction potential of Raphanus sativus var. longipinnatus (Japanese radish). They reported that MLP binds with all the tested HM ions and the metal ions binding capability of MLP toward Cu2+ found to be the maximum among them. The metal detoxification by MLP followed the order of Pb2 + > Zn2+ > Ni2+ > Cu2+ > Fe2+ > Cd2+ > Co2+. In another study, Hatano and Yamatsu (2018) evaluated the facilitatory effect of MLP on phytoextraction potential of three Brassica species grown in a medium containing Pb or Cd. They reported that biomass and Pb2+ uptake in the nutrient medium with 1 mM Pb nitrate were significantly increased by the addition of MLP and all the Pb2+ from the medium was accumulated in the root tissues. They concluded that MLP was able to detoxify Pb2+ and to improve their bioavailability in Brassica species.

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7 Mechanism of Microbial Degradation of Melanoidins The mechanism involved in melanoidins degradation through the action of enzymes is not still completely revealed, but it seems to be connected with that of the fungal ligninolytic pathway. The most studied enzymes, which are capable of breaking a large number of different chemical bonds present in phenolic and non-phenolic recalcitrant compounds present in distillery effluent, include MnP, LiP, and Lac (Gonzalez et al. 2008; Miyata et al. 1998a, b; Chandra et al. 2017; Pant and Adholeya 2009b). The decolorization of distillery effluent and the decrease in the MW of melanoidins are due to cleavage of C¼O, C¼C, and CN bonds in their molecules by microbial treatment, since the role of enzymes other than Lac or MnP in the degradation and decolorization of melanoidins by Coriolus (Trametes) strains was reported during the 1980s. Several scientific studies claimed that intracellular sugar-oxidase-type enzymes (glucose oxidase or sorbose oxidase) had melanoidindecolorizing activities. It was suggested that melanoidins were degraded and decolorized by the active oxygen species (O2; H2O2) formed by the reaction with glucose oxidase and/or sorbose oxidase. In addition, melanoidin-decolorizing activity in Coriolus versicolor Ps4a was attributed to intracellular enzymes composed of two types: sugar-independent and sugar-dependent activities. Hayase et al. (1984) reported that melanoidins were predominantly decolorized by glucose oxidase, but when glucose as a substrate of glucose oxidase was not used, melanoidins were not decolorized. Decolorization of melanoidins without supplementation of the carbon source is not possible (Ghosh et al. 2004). Several authors evaluated the decolorization of melanoidins in the presence of different carbon sources such as glucose, fructose, galactose, sucrose, maltose, etc., and they observed that melanoidin decolorization was achieved maximum with supplementation of glucose. They concluded that degradation and decolorization of melanoidin polymer were the results of co-metabolism of microorganism in the existence of an adequate concentration of glucose. This indicated that carbon bounds with melanoidins are not bioavailable to growing microorganisms. Therefore, decolorization of melanoidins without supplementation of the carbon source is not observed. Carbon source provided energy in form of ATP for growth and survival of microorganisms, which are necessary for the cleavage of conjugated C¼O, C¼C, and C  N bonds of melanoidins by the active oxygen such as H2O2 which is secondarily produced by the enzymatic oxidation reaction from glucose. Generally, H2O2 reacts with the OH- to give mainly perhydroxyl anion (HOO-) which has a strong nucleophilic activity. The HOO- is considered to nucleophilically attack the carbonyl groups of melanoidins. The finding suggested that both decolorization and degradation of melanoidins occur markedly in the alkaline medium due to the abundant presence of the HOO anion. The presence of HM ions such as Fe, Cu, etc. can induce an autodegradation of H2O2 to form the free radicals, viz., OH, OOH, and OO which may positively react with melanoidins (Miyata et al. 1998a, b, 2000; Kalavathi et al. 2001). Miyata and co-workers in 1998a, b studied the role of enzymes in melanoidin degradation and decolorization. Miyata et al. (1998a, b) demonstrated that synthetic melanoidin is

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decolorized by the participation of MnP and MIP of Coriolus hirsutus pellets and the H2O2 produced by glucose oxidase along with the participation of Lac enzyme. The presence of melanoidins and other similar compounds in the medium can induce the expression of MnP and Lac genes, both described as having synergistically action in the degradation and depolymerization of melanoidins and other industrial pollutants (Miyata et al. 2000; González et al. 2008). Gonzalez et al. (2008) reported the induction of Lac by molasses wastewaters and molasses-melanoidins (MM) in Trametes sp. I-62. The expression of Lac genes (lcc1 and lcc2) enhanced as a result of the addition of complete molasses wastewater and its high MW fraction to fungal cultures. Tapia-Tussell et al. (2015) also reported the expression of Lac genes in T. hirsuta strain Bm-2, in the presence of phenolic compounds, as well as its efficacy in removing colorants from vinasse. They found that Lac-encoding mRNA level was increased in the occurrence of phenolic compounds compared to control sample. The decolorization of vinasse was concomitant with the enhanced in laccase activity. The highest (69.2%) decolorization of 10% (v/v) vinasse was obtained in the presence of highest enzyme activity (2543.7 U mL
1). However, ligninolytic enzymes have failed to bring about complete mineralization of melanoidins and other organic pollutants present in distillery effluent.

8 Two-Step Treatment Approach for Bioremediation of PMDE In the current scenario, considering the advantages and disadvantages of different effluent treatment technologies, no single technology can be used for the complete treatment of SW and PMDE. Hence, there is a need to establish a comprehensive effluent treatment approach involving sequential treatment technologies. Therefore, treatment of complex effluent by a novel two-step treatment/phase separation method might be a novel and more promising approach for bioremediation of distillery effluent. For instance, Ghosh et al. (2002) investigated the treatment of anaerobically treated distillery effluent (PMDE) by a two-stage bioreactor by using P. putida followed by Aeromonas sp. strain EMa in a two-stage bioreactor. In the first stage bioreactor treatment, P. putida reduced the color and COD of distillery effluent up to 60% and 44.4, respectively, while in the second stage bioreactor treatment, Aeromonas sp. strain Ema reduced the effluent COD up to about 44.4%. Kaushik et al. (2010) investigated the treatment of SW in a three-stage bioreactor by using fungus followed by bacterial species. Treatment was first carried out by two fungus species Emericella nidulans var. lata and Neurospora intermedia followed by one bacterium Bacillus sp. The treated spent wash showed a significant reduction in COD (up to 93%) and color (up to 82%) after incubation of 30 h. However, a combination of CW plant treatment technology after bacterial degradation offers an excellent system for removal of color from industrial effluent and further reduction of BOD, COD, TDS, and HMs for safe disposal. Kumar and

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Chandra (2004) successfully treated PMDE in two-stage treatment process involving the transformation of recalcitrant coloring components of the PMDE by aerobic bacterium B. thuringiensis followed by a subsequent decline of a remaining load of pollutants by a macrophyte Spirodela polyrrhiza Schleiden. A similar biphasic (two steps) treatment of the PMDE was carried out in a CW with B. thuringiensis followed by T. angustata L. by Chandra et al. (2008) which resulted in 98–99% COD, BOD, and color reduction after 7 days. The results suggested that the bacterial pre-treatment of PMDE integrated with CW will improve the treatment process of PMDE and promote safe disposal of hazardous distillery waste. Chandra et al. (2012) reported 96.0% and 94.5% reduction in COD and BOD values in two-step sequential treatment of PMDE by bacteria and wetland plant Phragmites communis. He also characterized rhizosphere bacterial communities of P. communis and metabolic products generated during the sequential treatment of PMDE in CW plant treatment system. A two-step sequential treatment for sugarcane molasses-based anaerobically treated distillery effluent was reported by Pant and Adholeya (2009a). In the first step, distillery effluent was treated in a hydroponic-based system using two plant species, namely, Vetiveria zizanioides and P. karka to reduce (up to 84%) the high nitrogen content of the effluent. This first-step hydroponically treated distillery effluent was subjected for treatment by two fungus species in a bioreactor. Decolorization of effluent up to 86.33% was obtained with Pleurotus florida Eger EM1303 followed by Aspergillus flavus TERIDB9 (74.67%) with a significant reduction in COD as well. This study recommended that distillery effluent treatment without the need of high dilutions and addition of supplementary carbon sources. The efficacy of the two-step treatment approach has been demonstrated under pilot scale. For organic in-contaminants present in wastewater, this approach found to be effective in the field scale, and it is likely during the next 5–10 years, its use will become widespread.

9 Challenges and Future Prospects The major challenges of PMDE are the removal of TDS and degradation of colorcontributing compounds like melanoidin, polyphenols, caramel, an alkaline degradation product of hexose (ADPH), and their metabolic products as well as inorganic compounds before its safe disposal into the environment. Different amino acids, polysaccharide, and other organic compounds react at various extents; the polymerization most likely occurs in complex ways and extends to diverse levels. Accordingly, it is not easy to characterize those pigments responsible for color in distillery effluent. It should also be highlighted that several colorants instead of one type of pigment alone may result in the undesirable color, which will then enhance the difficulties of decolorization and degradation of organic pollutant present in distillery effluent. It is reported that conventional biological treatment processes are capable of accomplishing the reduction of ADPH, caramels, and melanoidins up to 70% and 67%, respectively (Mohana et al. 2009). However, the low MW

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pigmented organic compounds probably repolymerize during conventional biological treatment processes. However, melanoidins and polyphenols exhibit antioxidant and antimicrobial properties which make difficult the treatment of distillery effluent (Arimi et al. 2014). Thus, advanced techniques are strongly required to eliminate the OM and coloring compounds from the distillery effluent, which has been increasingly recognized as a tough challenge (Tsioptsias et al. 2015; Prajapati and Chaudhari 2015). Due to anionic nature of melanoidins in distillery effluent, it binds with a mixture of various HMs and some other organic compounds which makes more difficult its biological degradation. Melanoidins show variable absorption range in the UV region which makes more difficult to understand the mechanism of melanoidins degradation and decolorization and characterization of its metabolic products (Chandra et al. 2018a). But most of the decolorization and degradation studies of MM are reported at 475 nm only based on purified melanoidin absorption maxima with specific MW, while the MM contains a mixture of MRPs (i.e., initial, intermediate, and advanced stages with variable MW) (Kumar and Chandra 2018). Therefore, prior to attempting the decolorization and degradation of distillery effluent, the degradation of model melanoidins with a mixture of complex MRPs should be evaluated for its degradability. However, due to the presence of a mixture of different MW MRPs in distiller effluent, it shows variable absorption spectrum at a different wavelength. But absorption and elimination patterns of different peaks in melanoidins are not reported so far. Therefore, it is vital to reveal the decline pattern of the specific peak of MM in degradation and generation of its metabolites formed during the decolorization process. In addition, the details of the enzyme and its mode of action also need to be revealed. Moreover, the effect of sulfides and metals along with phenolic compounds required systematic studies to establish the mechanism of biological decolorization prior to its scope to develop an industrial-scale decolorization technique for safe disposal. Further, sulfate and heavy metals present in distillery effluent are reduced into black-colored precipitate of metal sulfides which act as a competitive inhibitor for sulfate-reducing bacteria and non-sulfate-reducing bacteria leading to inhibition of methanogenesis or sulfate reduction and giving toxicity to PMDE. Recently, the applications of bacteria and wetland plants in two-step/sequential treatment have been reported to be very promising technology for detoxification, but this has to be optimized yet with detailed physiology wetland plants, microbiology of plant rhizospheres, and detoxification mechanism. The bacterial communities grown in such an adverse environment are very much important to understand the microbiology of distillery waste. Acknowledgments The financial support from Department of Biotechnology (DBT), Govt. of India, to Professor Ram Chandra (Grant No. BT/PR13922/BCE/8/1129) and Rajiv Gandhi National Fellowship (RGNF) from University Grants Commission (UGC), New Delhi, to Mr. Vineet Kumar, Ph.D. student, is highly acknowledged.

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References Agnihotri S (2015) Decolorization study on synthetic colorants by using spore inoculum of Aspergillus oryzae JSA-1. Int J Curr Microbiol App Sci 4(10):12e17 Arimi MM, Zhang Y, Goetz G, Kiriamiti K, Geissen SU (2014) Antimicrobial colorants in molasses distillery wastewater and their removal technologies. Int Biodeterior Biodegrad 87:34–43 Arimi MM, Zhang Y, Götz G, Geisen S-U (2015) Treatment of melanoidin wastewater by anaerobic digestion and coagulation. Environ. Technol. 36(19): 2410–2418. Bharagava RN, Chandra R (2010a) Effect of bacteria treated and untreated post-methanated distillery effluent (PMDE) on seed germination, seedling growth and amylase activity in Phaseolus mungo L. J Hazard Mater 180:730–734 Bharagava RN, Chandra R (2010b) Biodegradation of the major color containing compounds in distillery wastewater by an aerobic bacterial culture and characterization of their metabolites. Biodegradation 21:703–711 Bharagava RN, Chandra R, Rai V (2008) Phytoextraction of trace elements and physiological changes in Indian mustard plants (Brassica nigra L.) grown in post methanated distillery effluent (PMDE) irrigated soil. Bioresour Technol 99:8316–8324 Bharagava RN, Chandra R, Rai V (2009) Isolation and characterization of aerobic bacteria capable of the degradation of synthetic and natural melanoidins from distillery effluent. World J Microbiol Biotechnol 25:737e744 Bharagava RN, Saxena G, Mulla SI, Patel DK (2017a) Characterization and identification of recalcitrant organic pollutants (ROPs) in tannery wastewater and its phytotoxicity evaluation for environmental safety. Arch Environ Contam Toxicol 75(2):259–272. https://doi.org/10. 1007/s00244-017-0490-x Bharagava RN, Saxena G, Chowdhary P (2017b) Constructed wetlands: An emerging phytotechnology for degradation and detoxification of industrial wastewaters. In: Bharagava RN (ed) Environmental pollutants and their bioremediation approaches, 1st edn. CRC Press, Taylor & Francis Group, Boca Raton, pp 397–426. https://doi.org/10.1201/9781315173351-15 Billore SK, Singh N, Ram HK, Sharma JK, Singh VP, Nelson RM, Dass P (2001) Treatment of a molasses based distillery effluent in a constructed wetland in central India. Water Sci Technol 44:441–448 Borja R, Martin A, Maestro R, Luque M, Durán J (1993) Enhancement of the anaerobic digestion of wine distillery wastewater by the removal of phenolic inhibitors. Bioresour Technol 45:99–104 Cammerer B, Jaluschkov V, Kroh LW (2002) Carbohydrates structures as part of the melanoidins skeleton. Int Congr Ser 1245:269 Chandra R, Kumar V (2017a) Detection of androgenic-mutagenic compounds and potential autochthonous bacterial communities during in situ bioremediation of post methanated distillery sludge. Front Microbiol 8:887 Chandra R, Kumar V (2017b) Detection of Bacillus and Stenotrophomonas species growing in an organic acid and endocrine-disrupting chemicals rich environment of distillery spent wash and its phytotoxicity. Environ Monit Assess 189:26 Chandra R, Kumar V (2017c) Phytoextraction of heavy metals by potential native plants and their microscopic observation of root growing on stabilized distillery sludge as a prospective tool for in-situ phytoremediation of industrial waste. Environ Sci Pollut Res 24:2605–2619 Chandra R, Kumar V (2018) Phytoremediation: A green sustainable technology for industrial waste management. In: Chandra R, Dubey NK, Kumar V (eds) Phytoremediation of environmental pollutants. CRC Press, Boca Raton Chandra R, Yadav S (2010) Potential of Typha angustifolia for phytoremediation of heavy metals from aqueous solution of phenol and melanoidin. Ecol Eng 36:1277–1284 Chandra R, Yadav S, Bharagava RN, Murthy RC (2008) Bacterial pretreatment enhances removal of heavy metals during treatment of post-methanated distillery effluent by Typha angustata L. J Environ Manag 88:1016–1024

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Chandra R, Bharagava RN, Kapley A, Purohit HJ (2012) Characterization of Phragmites cummunis rhizosphere bacterial communities and metabolic products during the two stage sequential treatment of post methanated distillery effluent by bacteria and wetland plants. Bioresour Technol 103:78–86 Chandra R, Saxena G, Kumar V (2015) Phytoremediation of environmental pollutants: An eco-sustainable green technology to environmental management. In: Chandra R (ed) Advances in biodegradation and bioremediation of industrial waste. CRC Press, Boca Raton, pp 1–29 Chandra R, Kumar V, Yadav S (2017) Extremophilic ligninolytic enzymes. In: Sani R, Krishnaraj R (eds) Extremophilic enzymatic processing of lignocellulosic feedstocks to bioenergy. Springer, Cham Chandra R, Kumar V, Tripathi S (2018a) Evaluation of molasses-melanoidin decolourisation by potential bacterial consortium discharged in distillery effluent. 3 Biotech 8:187 Chandra R, Kumar V, Tripathi S, Sharma P (2018b) Heavy metal phytoextraction potential of native weeds and grasses from endocrine-disrupting chemicals rich complex distillery sludge and their histological observations during in situ phytoremediation. Ecol Eng 111:143–156 Chandra R, Kumar V, Tripathi S, Sharma P (2018c) Phytoremediation of industrial pollutants and life cycle assessment. In: Chandra R, Dubey NK, Kumar V (eds) Phytoremediation of environmental pollutants. CRC Press, Boca Raton Chandra R, Kumar V, Singh K (2018d) Hyperaccumulator versus nonhyperaccumulator plants for environmental waste management. In: Chandra R, Dubey NK, Kumar V (eds) Phytoremediation of environmental pollutants. CRC Press, Boca Raton Chaturvedi S, Chandra R, Rai V (2006) Isolation and characterization of Phragmites australis (L.) rhizosphere bacteria from contaminated site for bioremediation of colored distillery effluent. Ecol Eng 27:202–207 Chaudhari PK, Mishra IM, Chand S (2007) Decolourization and removal of chemical oxygen demand (COD) with energy recovery: treatment of biodigester effluent of a molasses-based alcohol distillery using inorganic coagulants. Colloids Surf A Physicochem Eng Asp 296 (1–3):238–247 Chowdhary P, Raj A, Bharagava RN (2017) Environmental pollution and health hazards from distillery wastewater and treatment approaches to combat the environmental threats: A review. Chemosphere 194:229. https://doi.org/10.1016/j.chemosphere.2017.11.163 Dwyer J, Lant P (2008) Biodegradability of DOC and DON for UV/H2O2 pre-treated melanoidin based wastewater. Biochem Eng J 42:47–54 Fahya V, FitzGibbona FJ, McMullana G, Singhb D, Marchanta R (1997) Decolourisation of molasses spent wash by Phanerochaete Chrysosporium. Biotechnol Lett 19(1):97–99 Ghosh M, Ganguli A, Tripathi AK (2002) Treatment of anaerobically digested distillery spentwash in a two-stage bioreactor using Pseudomonas putida and Aeromonas sp. Process Biochem 37:857–862 Ghosh M, Verma SC, Mengoni A, Tripathi AK (2004) Enrichment and identification of bacteria capable of reducing chemical oxygen demand of anaerobically treated molasses spent wash. J Appl Microbiol 96:1278–1286 Ghosh M, Ganguli A, Tripathi AK (2009) Decolorization of anaerobically digested molasses spent wash by Pseudomonas putida. Appl Biochem Microbiol 45:68e73 Gonzalez T, Terron MC, Yague S, Zapico E, Galletti GC, Gonzalez AE (2000) Pyrolysis/gas chromatography/mass spectrometry monitoring of fungal-biotreated distillery wastewater using Trametes sp I-62 (CECT 20197). Rapid Commun Mass Spectrom 14:1417–1424 González T, Terrón MC, Yagüe S, Junca H, Carbajo JM, Zapico EJ, Silva R, Arana-Cuenca A, Téllez A, González AE (2008) Melanoidin- containing wastewaters induce selective laccase gene expression in the white-rot fungus Trametes sp. I-62. Res Microbiol 159(2):103–109 Goutam SP, Saxena G, Singh V, Yadav AK, Bharagava RN (2018) Green synthesis of TiO2 nanoparticles using leaf extract of Jatropha curcas L. for photocatalytic degradation of tannery wastewater. Chem Eng J 336:386–396. https://doi.org/10.1016/j.cej.2017.12.029

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Hatano K, Kanazawa K, Tomura H, Yamatsu T, Tsunoda K, Kubota K (2016) Molasses melanoidin promotes copper uptake for radish sprouts: the potential for an accelerator of phytoextraction. Environ Sci Pollut Res 23:17656–17663 Hatano K, Yamatsu T (2018) Molasses melanoidin-like products enhance phytoextraction of lead through three Brassica species. Int J Phytoremediation 20:552. https://doi.org/10.1080/ 15226514.2017.1393397 Hatano K, Kikuchi S, Miyakawa T, Tanokura M, Kubota K (2008) Separation and characterization of the colored material from sugarcane molasses. Chemosphere 71:1730–1737 Hayase F, Kim SB, Kato H (1984) Decolorization and degradation products of the melanoidins by hydrogen peroxide. Agric Biol Chem 48(11):2711–2717 Hayase F (2000) Recent development of 3-deoxyosone related Maillard reaction products. Food Sci Technol Res 6(2):79–86 Jiranuntipon S, Chareonpornwattana S, Damronglerd S, Albasi C, Delia ML (2008) Decolorization of synthetic melanoidins-containing wastewater by a bacterial consortium. J Ind Microbiol Biotechnol 35:1313–1321 Kabir ER, Rahman MS, Rahman I (2015) A Review on endocrine disruptors and their possible impacts on human health. Environ Toxicol Pharmacol. https://doi.org/10.1016/j.etap.2015.06. 009 Kalavathi DF, Uma L, Subramanian G (2001) Degradation and metabolization of the pigment— melanoidin in distillery effluent by the marine cyanobacterium Oscillatoria boryana BDU 92181. Enzym Microb Technol 29:246–251 Kansal A, Rajeshwari KV, Balakrishnan M, Lata K, Kishore VVN (1998) Anaerobic digestion technology for energy recovery from industrial wastewater: a study in Indian context, TERI inform. Monitor Environ Sci 3:67–75 Kaushik G, Thakur IS (2009) Isolation of fungi and optimization of process parameters for decolorization of distillery mill effluent. World J Microbiol Biotechnol 25:955 Kaushik G, Thakur IS (2013) Adsorption of colored pollutants from distillery spent wash by native and treated fungus: Neurospora intermedia. Environ Sci Pollut Res 20(2):1070–1078 Kaushik G, Gopal M, Thakur IS (2010) Evaluation of performance and community dynamics of microorganisms during treatment of distillery spent wash in a three stage bioreactor. Bioresour Technol 101:4296–4305 Kim SJ, Shoda M (1999) Batch decolorization of molasses by suspended and immobilized fungus of Geotrichum candidum. J Biosci Bioeng 88(5):586–589 Kim SB, Hayase F, Kato H (1985) Decolourisation and degradation products of melanoidin on ozonolysis. Agric Biol Chem 49:785–792 Krishnamoorthy S, Premalatha M, Vijayasekaran M (2017) Characterization of distillery wastewater–An approach to retrofit existing effluent treatment plant operation with phycoremediation. J Clean Prod 148:735–750 Krzywonos M (2012) Decolorization of sugar beet distillery effluent using mixed cultures of bacteria of the genus Bacillus. Afr J Biotechnol 11(14):3464e3475 Kumar P, Chandra R (2004) Detoxification of distillery effluent through Bacillus thuringiensis (MTCC 4714) enhanced phytoremediation potential of Spirodela polyrrhiza (L.) Schliden. Bull Environ Contam Toxicol 73:903–910 Kumar P, Chandra R (2006) Decolourisation and detoxification of synthetic molasses melanoidins by individual and mixed cultures of Bacillus spp. Bioresour Technol 97:2096–2102 Kumar V, Chandra R (2018) Characterisation of MnP and laccase producing bacteria capable for degradation of sucrose glutamic acid-maillard products at different nutritional and environmental conditions. World J Microbiol Biotechnol 34:32 Kumar U, Muthukrishnan RM, Guha BK (2008) Tertiary treatment of distillery wastewater by nanofiltration. Desalination 230(1–3):70–78 Liang Z, Wang Y, Zhou Y, Liu H, Wu Z (2009a) Variables affecting melanoidins removal from molasses wastewater by coagulation/ flocculation. Sep Purif Technol 68(3):382–389

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Chapter 21

Progresses in Bioremediation Technologies for Industrial Waste Treatment and Management: Challenges and Future Prospects Ram Naresh Bharagava and Gaurav Saxena

Abstract Industrial wastewater treatment and management is a major challenge of the twenty-first century and essential to safeguard the environment and public health. Industrial wastewaters are considered as one of the major sources of environmental contamination because these carry a variety of environmental contaminants that may cause serious health hazards in living beings. To protect the environment and public health from the adverse effects of such industrial pollutants, several methods are currently being applied to manage such industrial wastes. These methods include physicochemical techniques, which are not eco-friendly in nature as these use chemicals for environmental cleanup and thus cause secondary pollution and are costly. However, bioremediation technologies are one of the self-driven eco-friendly methods as these rely on the activity of microbes and plants that remove an array of pollutants from polluted wastewaters. This chapter reviews the progresses made in the bioremediation technologies for industrial waste treatment and management with reference to tannery wastewater and focuses on challenges and future directions in the field. Keywords Industrial wastewater · Treatment · Management · Progresses · Challenges · Prospects

1 Introduction Industries are responsible for the economic growth of the developing countries as these generate revenue and provide employment opportunities to locals and thus positively affect the well-being of humankind. However, these are also the major R. N. Bharagava (*) · G. Saxena Laboratory for Bioremediation and Metagenomics Research (LBMR), Department of Microbiology (DM), Babasaheb Bhimrao Ambedkar (Central) University, Lucknow, Uttar Pradesh, India © Springer Nature Singapore Pte Ltd. 2020 R. N. Bharagava, G. Saxena (eds.), Bioremediation of Industrial Waste for Environmental Safety, https://doi.org/10.1007/978-981-13-3426-9_21

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polluters of the environment as these discharge huge amount of highly toxic and hazardous wastewaters that cause serious soil/water pollution. Industrial effluents mainly have a characteristic of high biochemical oxygen demand (BOD), chemical oxygen demand (COD), total dissolved solids (TDS), total suspended solids (TSS), and a variety of potentially toxic organic and inorganic pollutants which may cause serious toxicity in human beings upon exposure (Kishor et al. 2018; Goutam et al. 2018; Gautam et al. 2017; Bharagava et al. 2017a, b; Saxena et al. 2016; Saxena and Bharagava 2017, 2015). Currently, physicochemical methods are applied to treat and manage such highstrength wastewater to protect the environment and public health. However, these methods are environmentally destructive as these cause secondary pollution, while environmental cleanup is costly and disturbs the site to be cleaned. Being a cleangreen approach, bioremediation provides a sustainable solution to effectively treat and manage industrial effluents to achieve environmental safety and security. Bioremediation is an eco-friendly technique wherein microbes and plants or their enzymes are used to degrade and detoxify the organic and inorganic pollutants in contaminated matrix and thus restore the contaminated sites (Bharagava et al. 2017c, 2018; Saxena et al. 2019; Saxena and Bharagava 2016; Chandra et al. 2015). To date, a lot of advances have been made in the bioremediation technologies to treat and manage such industrial waste. Therefore, this chapter reviews the progresses, challenges, and future research prospects in the bioremediation technologies used for the treatment and management of industrial waste with special reference to tannery wastewater.

2 Progresses, Challenges, and Future Prospects in Bioremediation Technologies for Waste Treatment and Management Currently, the conventional wastewater treatment plants (WWTPs) are considered as the best examples of bioremediation technologies. Conventional common effluent treatment plants (CETPs) are viewed as the most feasible option for the effective treatment and detoxification of wastewaters (Moharikar et al. 2005; Kapley et al. 2007; Moura et al. 2009) and mostly relied on activated sludge treatment process. Microorganisms (such as bacteria/fungi/microalgae) are considered as the eco-friendly tools for the degradation and detoxification of organic and inorganic pollutants in industrial wastewaters (Bharagava et al. 2017a). CETPs primarily relied on the activated sludge process (a biological treatment process) and are used for the secondary treatment process. In CETPs, TWW received from nearby tannery industries is aerated with potential microbial populations, which results in the degradation and detoxification of organic and inorganic pollutants to achieve the high quality of treated wastewater (Munz et al. 2008). The diversity of microbial populations growing in the CETP primarily depends on the nature and characteristics of

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pollutants; geography and shift in the diversity of microbes may compromise the whole process of wastewater treatment (Chandra et al. 2011). However, there is an increasing ecological issue concerning the discharge of recalcitrant organic pollutants (ROPs) in industrial wastewaters because they are resistant to biodegradation during the secondary (biological) treatment at CETP and thus represent serious challenges to the sustainability of aquatic resources. For instance, Bharagava et al. (2017b) have reported the presence of a variety of ROPs such as benzyl chloride, butyl octyl phthalate (BOP), 2,6-dihydroxybenzoic acid 3TMS, dibutyl phthalate, benzyl alcohol, benzyl butyl phthalate (BBP), 4-chloro-3methyl phenol, phthalic acid, 20 60 -dihydroxyacetophenone, diisobutyl phthalate, 4-biphenyltrimethylsiloxane, di-(2ethy hexyl)phthalate (DEHP), 1,2-benzenedicarboxylic acid, dibenzyl phthalate (DBP), and nonylphenol (NP) in TWW after secondary treatment process using gas chromatography-mass spectroscopy data analysis. The US Environmental Protection Agency (USEPA) listed these chemical compounds as “priority pollutants” and banned their application in leather production processes because they have been reported to cause endocrine disruption and toxicity to aquatic organisms (Bharagava et al. 2017a). It has been reported that the CETP is not working properly and wastewater discharge after secondary treatment does not meet the permissible limits as set by the Environment Protection Agency. Hence, there is an urgent need to upgrade the existing CETP as per requirement to adequately treat and detoxify the untreated TWW received from tannery industries to prevent the pollution as well as safeguard the public health. Microbes are oftenly used in the degradation and detoxification of TWW. For instance, Kim et al. (2014) reported the 98.3% of COD and 88.5% of Cr removal from TWW. Noorjahan (2014) reported 90% of COD, 90% of BOD, and 63.8% of Cr removal from TWW using E. coli whereas 95.4% of COD and BOD and 73.5% of Cr removal from TWW using Bacillus sp. Yusuf et al. (2013) reported 87.6% of COD from TWW using B. subtilis whereas 85.2% of COD from TWW using P. fragi. El-Bestawy et al. (2013) reported 79.16 of COD, 94.14 of BOD, and 93.66 of Cr from TWW using an optimized bacterial consortium containing Escherichia coli O7:K1 CE10, Providencia vermicola W9B-11, Bacillus sp. 58, Pseudomonas stutzeri M15–10-3, Bacillus amyloliquefaciens T004, and Bacillus sp. PL47. Sivaprakasam et al. (2008) also reported 80% of COD removal from TWW using a bacterial consortium (B. flexus, P. aeruginosa, S. aureus, and E. homiense). Further, the highly saline nature of TWW may create obstacles in the biotreatment of TWW, which include as follows (Saxena et al. 2016; Sundarapandiyan et al. 2010; Sivaprakasam et al. 2008): (a) a high concentration of salts (>3–5% w/v) makes the conventional bacterial culture less adaptive during the biotreatment of TWW and thus a less effective treatment; (b) if adapted to high salt concentration, it may get easily lost when conventional bacterial culture is grown in salt-free growth medium, and (c) cellular disruption in the acclimatized bacterial culture due to even slight changes in the salt concentration (ranging from 0.5 to 2% w/v) thus leads to system failure. Further, the presence of high concentration of potentially toxic metals such as Cr and less biodegradable tannins in TWW also inhibit its biotreatment

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(Schrank et al. 2004). Cr(VI) has been reported to cause inhibition of growth of the nitrifying/denitrifying bacteria as well as other heterotrophs (Stasinakis et al. 2002; Farabegoli et al. 2004). However, being of low-cost and flexible operation, a sequencing batch reactor (SBR) may provide a sustainable solution to effective biological treatment as it is feasible for the removal of nitrogen from TWW in the presence of inhibitors due to selection and enrichment of a particular microbial species (Farabegoli et al. 2004; Durai and Rajasimmam 2011; Rameshraja and Suresh 2011; Lofrano et al. 2013). Sulfide present in TWW primarily inhibits the methanogenesis during its anaerobic treatment, and this might be due to its toxic effect on remediating bacteria, competition for substrate among the methanogenic and sulfate-reducing bacteria, and trace element precipitation (Midha and Dey 2008; Rameshraja and Suresh 2011; Mannucci et al. 2014). Nevertheless, how sulfide creates toxicity and its mechanisms are still unknown and need to be understood. Further, constructed wetlands (CWs) are also representing the most promising wastewater treatment solution (Bharagava et al. 2017c). The phytoremediation of Cr from TWW uses Brachiaria decumbens, Phragmites australis, and Pennisetum purpureum CW (Mant et al. 2004). The potential of Typha latifolia, Canna indica, Stenotaphrum secundatum, Phragmites australis, and Iris pseudacorus for the phytoremediation of TWW in CW has been also reported, and Typha latifolia and Phragmites australis were the two dominant plant species (Calheiros et al. 2007). Plant species, Sarcocornia fruticosa and Arundo donax, during the phytoremediation of secondary treated TWW in the horizontal subsurface flow CW effectively reduced the COD and BOD up to 80 and 90%, respectively (Calheiros et al. 2012). Further, Dunn et al. (2013) have also suggested TWW as a suitable medium for the growth and cultivation of Arthrospira (Spirulina). Moreover, the chromium salt can be retained in wetlands with nonspecialized supporting media (Dotro et al. 2012). However, future research should be focused on (a) understanding of the dynamics of microbiological processes and correlation between the biotic and abiotic processes in CWs; (b) knowledge of the dynamics of element cycle which might help us to know the basic mechanism of greenhouse gas emission in CWs; and (c) understanding of microbial community and plant-microbe interactions to know the underlying mechanism of pollutant removal in CWs (Carvalho et al. 2017). Anammox is an emerging technology for wastewater treatment and used to remove ammonia from TWW during its anaerobic treatment. It is a low-cost and less energy-intensive biological wastewater treatment process wherein oxidation of ammonia takes place in the absence of oxygen with nitrite as a favored electron acceptor. On contrary to conventional nitrification/denitrification process, it consumes less amount of O2 (up to 50%), has no organic chemicals, and saves high cost of operation (up to 90%) for sludge disposal (Anjali and Sabumon 2014). Hence, it is more suitable for the industrial effluents with huge amount of ammonia. On the other hand, the long time required for its start-up and its reductive nature in the presence of organic chemicals and NH4-N limit its commercial applications (Anjali and Sabumon 2014). Hence, there is a need to construct a potential consortium able to do anammox in the presence of organic chemicals. Furthermore, the construction of

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a potential bacterial consortium containing anammox-oxidizing bacteria, anammoxdenitrifying bacteria, and anammox bacteria will be efficient for the treatment of industrial effluents having high concentration of ammonia and organic chemicals. Further, phytoremediation of TWW is an excellent strategy for its effective treatment and management. For instance, Kassaye et al. (2017) reported the phytoremediation potential of swamp smartweed (Polygonum coccineum), para grass (Brachiaria mutica), and papyrus (Cyprus papyrus) for Cr-containing TWW. They reported that all the three plants exhibited a significant transfer of Cr from wastewater to roots and shoots, but removal efficiency of Cr for swamp smartweed was relatively low as compared to para grass and papyrus and further suggested the use of para grass and papyrus for effective phytoremediation of TWW. Gupta et al. (2018) studied the microbe-assisted phytoremediation of tannery effluent contaminated agricultural soils. They isolated a Cr6+-resistant plant growth-promoting Pseudomonas sp. (strain CPSB21) from the tannery effluent-contaminated agricultural soils and evaluated for the various plant growth-promoting activities, oxidative stress tolerance, and Cr6+ bioremediation. Further, they applied the isolated strain for microbe-assisted phytoremediation and reported that the inoculation of strain CPSB21 alleviated the Cr6+ toxicity and enhanced the plant growth parameters and nutrient uptake in sunflower plant during pot experiment. However, the selection of suitable plants for phytoremediation of TWW is a challenging task as it should primarily be of high salt and chromium tolerant in nature. In addition, electrobioremediation is an emerging technology for waste treatment and management. For instance, Kanagasabi et al. (2012) reported the electrobioremediation of TWW in electrooxidation reactor containing immobilized bacterial biomass. They reported a maximum reduction in COD up to 73.1% at 1.5 A dm2 from raw TWW using electrooxidation technique and up to 91.5% reduction in COD from diluted TWW using immobilized Bacillus Strain B. However, the combined application of electrooxidation and biooxidation approach was proven good for the degradation of TWW samples and achieved 66.2% and 76.6% COD degradation, respectively. However, the optimization parameters related to electrical operation such as intensity of electrical field, mode of current application, the distance between electrodes, and period of stimulation and their effect on the microbes are the associated key challenges that must be taken into consideration. From the above discussion, it is clear that every remediation technology has its own merits and demerits and thus wisely chosen. However, further research efforts are required to expand the scope and efficacy of the remediation methods used to effectively treat and manage TWW to safeguard the environment and public health.

3 Conclusions and Recommendations (a) TWW is ranked as one of the highly toxic wastewaters among all the industrial wastewater types.

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(b) Constructed treatment wetlands are viewed as the emerging and eco-friendly treatment solution for TWW but still have certain drawbacks which should be addressed for full-scale application. (c) Combinatorial treatment approaches (physical + chemical + biological treatment) are considered as effective in the treatment of TWW and can reduce pollutant concentration to an acceptable level. (d) Search for novel microbes with high remediation potential is warranted to effectively treat and manage TWW. (e) Electrobioremediation is an innovative hybrid bioremediation technology for TWW treatment, but some associated challenges need to be addressed. (f) There is a need to identify the potential bacteria and construct a consortium able to do anammox in the presence of organic chemicals to effectively treat the TWW containing both ammonia and organic carbon. Acknowledgments The financial support such as “Major Research Projects” (Grant No.: EEQ/2017/000407) from “Science and Engineering Research Board” (SERB), Department of Science and Technology (DST), Government of India (GOI), New Delhi, India, and University Grant Commission (UGC) Fellowship received by Mr. Gaurav Saxena for doctoral studies is duly acknowledged.

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