Methods for Bioremediation of Water and Wastewater Pollution [1st ed.] 9783030489847, 9783030489854

Table of contents :
Front Matter ....Pages i-xv
Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals (Anca Giorgiana Grigoras)....Pages 1-26
Hardware and Software Remediation Technologies for Water Resources Pollution (B. Abadi, M. Shahvali)....Pages 27-60
Anaerobic Biotechnology for the Treatment of Pharmaceutical Compounds and Hospital Wastewaters (Ali Khadir, Afsaneh Mollahosseini, Mehrdad Negarestani, Ali Mardy)....Pages 61-84
Bacterial Metabolites for Removal of Toxic Dyes and Heavy Metals (Sriparna Datta, Dipanjan Sengupta, Ishika Saha)....Pages 85-116
Bacterial Biofilms in Bioremediation of Metal-Contaminated Aquatic Environments (Rafig Gurbanov, Feride Severcan)....Pages 117-135
Laccase-Mediated Bioremediation of Dye-Based Hazardous Pollutants (Muhammad Bilal, Syed Salman Ashraf, Hafiz M. N. Iqbal)....Pages 137-160
Remediation of Freshwaters Contaminated by Cyanobacteria (Sana Saqrane, Brahim Oudra, Moulay Abderrahim El Mhammedi)....Pages 161-180
Biochemical Methods for Water Purification (Gulzar Muhammad, Adeel Mehmood, Munazza Shahid, Raja Shahid Ashraf, Muhammad Altaf, Muhammad Ajaz Hussain et al.)....Pages 181-212
Biosorptive Removal of Toxic Pollutants from Contaminated Water (A. Saravanan, P. Senthil Kumar)....Pages 213-224
Microbial Exopolymeric Substances for Metal Removal (Caleb Cheah, Adeline Su Yien Ting)....Pages 225-251
Bioremediation of Bisphenols and Phthalates from Industrial Effluents: A Review (Meghana Ganta, Anuradha Shilli, Soukhya Channapatana Adishesh, Bhanu Revathi Kurella, Shinomol George Kunnel)....Pages 253-265
Phytoextraction of Heavy Metals (A. N. Anoopkumar, Sharrel Rebello, Elsa Devassy, K. Kavya Raj, Sreedev Puthur, Embalil Mathachan Aneesh et al.)....Pages 267-276
Tree Barks for Bioremediation of Heavy Metals from Polluted Waters (Puneet P. Jain, Zufeshan Nahar Ali, Srishti J. Sisodiya, Shinomol George Kunnel)....Pages 277-288
Environmental Effects and Microbial Detoxification of Textile Dyes (Zahid Maqbool, Habibullah Nadeem, Faisal Mahmood, Muhammad Hussnain Siddique, Tanvir Shahzad, Farrukh Azeem et al.)....Pages 289-326
Natural Remediation Techniques for Water Quality Protection and Restoration (George Pavlidis, Helen Karasali)....Pages 327-340
Phytoextraction of Heavy Metals from Complex Industrial Waste Disposal Sites (Babatunde Oladipo, Aramide M. Akintunde, Sheriff O. Ajala, Samuel O. Olatunji, Olayomi A. Falowo, Eriola Betiku)....Pages 341-371
Biosorption of Nickel (II) and Cadmium (II) (Rajeswari M. Kulkarni, K. Vidya Shetty, G. Srinikethan)....Pages 373-391
Biological Strategies for Heavy Metal Remediation (Memory Tekere)....Pages 393-413
Back Matter ....Pages 415-424

Citation preview

Environmental Chemistry for a Sustainable World 51

Inamuddin Mohd Imran Ahamed Eric Lichtfouse Abdullah M. Asiri   Editors

Methods for Bioremediation of Water and Wastewater Pollution

Environmental Chemistry for a Sustainable World Volume 51

Series Editors Eric Lichtfouse, Aix-Marseille University, CNRS, IRD, INRAE, Coll France, CEREGE, Aix-en-Provence, France Jan Schwarzbauer, RWTH Aachen University, Aachen, Germany Didier Robert, CNRS, European Laboratory for Catalysis and Surface Sciences, Saint-Avold, France

Other Publications by the Editors

Books Environmental Chemistry http://www.springer.com/978-3-540-22860-8 Organic Contaminants in Riverine and Groundwater Systems http://www.springer.com/978-3-540-31169-0 Sustainable Agriculture Volume 1: http://www.springer.com/978-90-481-2665-1 Volume 2: http://www.springer.com/978-94-007-0393-3 Book series Environmental Chemistry for a Sustainable World http://www.springer.com/series/11480 Sustainable Agriculture Reviews http://www.springer.com/series/8380 Journals Environmental Chemistry Letters http://www.springer.com/10311

More information about this series at http://www.springer.com/series/11480

Inamuddin • Mohd Imran Ahamed Eric Lichtfouse • Abdullah M. Asiri Editors

Methods for Bioremediation of Water and Wastewater Pollution

Editors Inamuddin Department of Chemistry King Abdulaziz University Jeddah, Saudi Arabia Department of Applied Chemistry Aligarh Muslim University Aligarh, Uttar Pradesh, India Eric Lichtfouse Aix-Marseille University, CNRS, IRD, INRAE, Coll France, CEREGE Aix-en-Provence, France

Mohd Imran Ahamed Department of Chemistry Aligarh Muslim University Aligarh, Uttar Pradesh, India Abdullah M. Asiri Department of Chemistry King Abdulaziz University Jeddah, Saudi Arabia

ISSN 2213-7114 ISSN 2213-7122 (electronic) Environmental Chemistry for a Sustainable World ISBN 978-3-030-48984-7 ISBN 978-3-030-48985-4 (eBook) https://doi.org/10.1007/978-3-030-48985-4 © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 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, expressed 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 Switzerland AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Preface

The rapid growth of population and industrialization has resulted in the disposal of harmful chemicals into the environment, thus inducing health issues for life. The major sources of contamination include industries, agrochemistry, mining activities and waste disposal. Remediation of environmental media using biological methods is actually a fast-developing research field because bioremediation is considered as a cheap, sustainable, green and socially acceptable way of cleaning. Scientists have designed many remediation mechanisms such as detoxification, immobilization, degradation, concentration, disposal and recycling by the action of microorganisms or enzymes. Bioremediation strategies include natural attenuation, biostimulation by encouraging natural processes of biodegradation and bioaugmentation by adding beneficial microorganisms. This book presents strategies, concepts and methods for bioremediation of metals, dyes and organic pollutants. The structure, classification, properties, ecotoxicology and bioremediation of various pollutants are discussed. This book is a good reference guide for faculty, postgraduates, researchers and industrial professionals who are linked to environmental science, analytical chemistry, biotechnology, nutrition, photochemistry and toxicology.

v

Preface

Contamination through soil and water

Microbial remediation

Bioremediation Mycoremediation

Pharmaceutical Compounds Toxic Dyes Cyanobacteria Bisphenols Phthalates

Phytoremediation

Industrial Effluents

vi

Chapter 1 discusses the bioremediation of samples contaminated by lead, cadmium or chromium using Pseudomonas-based biosorbents. Additionally, the modelling of the biosorption process of heavy metals, the efficiency of new biosorbents and toxicological limits are highlighted. Chapter 2 discusses world scientific paradigms on water pollution remediation, as the author believes that environmental catastrophes are a consequence of how the scientific society looks at nature. Chapter 3 introduces anaerobic processes for the removal of pollutants. Chapter 4 presents an alternative approach for remediating dyes and metals by sorption using bacterial strains and metabolites. Chapter 5 focuses on bacterial biofilm-based strategies for metal sequestration, with emphasis on biofilms, quorum sensing and functional bacterial items. Chapter 6 reviews laccase for immobilization of dyes. Chapter 7 describes economical, eco-friendly and efficient biochemical water purification methods, with focus on removal efficiencies of various microorganisms, and on the effect of temperature, pH and initial dye concentration. Wastewater bioremediation of metals, dyes and pigments by plants, bacteria, fungi and algae are presented in Chap. 8. Chapter 9 details the application of microorganisms to remove metals from wastewater, with focus on biosorption and mechanisms. Sorption of metals on exopolymeric substances from bacteria is then discussed in Chap. 10. Chapter 11 presents the bioremediation of bisphenols and phthalates by chelation with nanoparticles from soil microbiota, plants and fungi (Figure). Metal phytoextraction is discussed in Chaps. 12 and 16. Chapter 13 reviews the use of bark and extracts to decrease metal contamination. Chapter 14 discusses dye history, classification,

Preface

vii

properties, remediation and environmental impact. Natural remediation techniques such as vegetative filter strips, phytoremediation and constructed wetlands are presented in Chap. 15. Chapter 17 focuses on nickel and cadmium removal from aqueous solutions using microorganisms. Biological strategies for the removal of metals using microorganisms and plants are presented in Chap. 18. Jeddah, Saudi Arabia Aligarh, India Aligarh, India Aix-en-Provence, France Jeddah, Saudi Arabia

Inamuddin Mohd Imran Ahamed Eric Lichtfouse Abdullah M. Asiri

Contents

1

2

3

4

Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Anca Giorgiana Grigoras

1

Hardware and Software Remediation Technologies for Water Resources Pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Abadi and M. Shahvali

27

Anaerobic Biotechnology for the Treatment of Pharmaceutical Compounds and Hospital Wastewaters . . . . . . . . . . . . . . . . . . . . . . Ali Khadir, Afsaneh Mollahosseini, Mehrdad Negarestani, and Ali Mardy Bacterial Metabolites for Removal of Toxic Dyes and Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sriparna Datta, Dipanjan Sengupta, and Ishika Saha

61

85

5

Bacterial Biofilms in Bioremediation of Metal-Contaminated Aquatic Environments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117 Rafig Gurbanov and Feride Severcan

6

Laccase-Mediated Bioremediation of Dye-Based Hazardous Pollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137 Muhammad Bilal, Syed Salman Ashraf, and Hafiz M. N. Iqbal

7

Remediation of Freshwaters Contaminated by Cyanobacteria . . . . 161 Sana Saqrane, Brahim Oudra, and Moulay Abderrahim El Mhammedi

8

Biochemical Methods for Water Purification . . . . . . . . . . . . . . . . . . 181 Gulzar Muhammad, Adeel Mehmood, Munazza Shahid, Raja Shahid Ashraf, Muhammad Altaf, Muhammad Ajaz Hussain, and Muhammad Arshad Raza

ix

x

Contents

9

Biosorptive Removal of Toxic Pollutants from Contaminated Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213 A. Saravanan and P. Senthil Kumar

10

Microbial Exopolymeric Substances for Metal Removal . . . . . . . . . 225 Caleb Cheah and Adeline Su Yien Ting

11

Bioremediation of Bisphenols and Phthalates from Industrial Effluents: A Review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 253 Meghana Ganta, Anuradha Shilli, Soukhya Channapatana Adishesh, Bhanu Revathi Kurella, and Shinomol George Kunnel

12

Phytoextraction of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . 267 A. N. Anoopkumar, Sharrel Rebello, Elsa Devassy, K. Kavya Raj, Sreedev Puthur, Embalil Mathachan Aneesh, Raveendran Sindhu, Parameswaran Binod, and Ashok Pandey

13

Tree Barks for Bioremediation of Heavy Metals from Polluted Waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 277 Puneet P. Jain, Zufeshan Nahar Ali, Srishti J. Sisodiya, and K. Shinomol George

14

Environmental Effects and Microbial Detoxification of Textile Dyes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 289 Zahid Maqbool, Habibullah Nadeem, Faisal Mahmood, Muhammad Hussnain Siddique, Tanvir Shahzad, Farrukh Azeem, Muhammad Shahid, Saima Muzammil, and Sabir Hussain

15

Natural Remediation Techniques for Water Quality Protection and Restoration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 327 George Pavlidis and Helen Karasali

16

Phytoextraction of Heavy Metals from Complex Industrial Waste Disposal Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 341 Babatunde Oladipo, Aramide M. Akintunde, Sheriff O. Ajala, Samuel O. Olatunji, Olayomi A. Falowo, and Eriola Betiku

17

Biosorption of Nickel (II) and Cadmium (II) . . . . . . . . . . . . . . . . . . 373 Rajeswari M. Kulkarni, K. Vidya Shetty, and G. Srinikethan

18

Biological Strategies for Heavy Metal Remediation . . . . . . . . . . . . . 393 Memory Tekere

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 415

Contributors

B. Abadi Department of Biosystem Mechanics Engineering, Faculty of Agriculture, University of Maragheh, Maragheh, Iran Soukhya Channapatana Adishesh Department of Biotechnology, Dayananda Sagar College of Engineering, Bengaluru, Karnataka, India Sheriff O. Ajala Frank H. Dotterweich College of Engineering-Sustainable Energy, Texas A&M University-Kingsville, Kingsville, TX, USA Aramide M. Akintunde Department of Environmental Engineering, Texas A&M University-Kingsville, Kingsville, TX, USA Zufeshan Nahar Ali Department of Biotechnology, Dayanandasagar College of Engineering, Bengaluru, Karnataka, India Muhammad Altaf Department of Chemistry, GC University Lahore, Lahore, Pakistan Embalil Mathachan Aneesh Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India A. N. Anoopkumar Department of Zoology, Christ College, University of Calicut, Irinjalakuda, Kerala, India Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India Raja Shahid Ashraf Department of Chemistry, GC University Lahore, Lahore, Pakistan Syed Salman Ashraf Department of Chemistry, College of Arts and Sciences, Khalifa University, Abu Dhabi, United Arab Emirates Farrukh Azeem Department of Bioinformatics and Biotechnology, Government College University, Faisalabad, Pakistan

xi

xii

Contributors

Eriola Betiku Department of Chemical Engineering, Obafemi Awolowo University, Ile-Ife, Osun State, Nigeria Muhammad Bilal School of Life Science and Food Engineering, Huaiyin Institute of Technology, Huaian, China Parameswaran Binod Microbial Processes and Technology Division, CSIRNational Institute of Interdisciplinary Science and Technology (CSIR-NIIST), Trivandrum, Kerala, India Caleb Cheah School of Science, Monash University Malaysia, Bandar Sunway, Selangor Darul Ehsan, Malaysia Sriparna Datta Department of Chemical Technology, Rajabazar Science College, University of Calcutta, Kolkata, India Elsa Devassy Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India Moulay Abderrahim El Mhammedi Polydisciplinary Faculty, Laboratory of Chemistry, Modeling and Environmental Sciences, Sultan Moulay Slimane University, Khouribga, Morocco Olayomi A. Falowo Department of Chemical Engineering, Obafemi Awolowo University, Ile-Ife, Osun State, Nigeria Meghana Ganta Department of Biotechnology, Dayananda Sagar College of Engineering, Bengaluru, Karnataka, India Anca Giorgiana Grigoras “Petru Poni” Institute of Macromolecular Chemistry, Iasi, Romania Rafig Gurbanov Department of Molecular Biology and Genetics, Bilecik Şeyh Edebali University, Bilecik, Turkey Biotechnology Application and Research Center, Bilecik Şeyh Edebali University, Bilecik, Turkey Muhammad Ajaz Hussain Department of Chemistry, University of Sargodha, Sargodha, Pakistan Sabir Hussain Department of Environmental Sciences and Engineering, Government College University, Faisalabad, Pakistan Hafiz M. N. Iqbal Tecnologico de Monterrey, School of Engineering and Sciences, Campus Monterrey, Monterrey, Mexico Puneet P. Jain Department of Biotechnology, Dayanandasagar College of Engineering, Bengaluru, Karnataka, India Helen Karasali Laboratory of Chemical Control of Pesticides, Department of Pesticides Control and Phytopharmacy, Benaki Phytopathological Institute, Athens, Greece

Contributors

xiii

Ali Khadir Young Researcher and Elite Club, Yadegar-e-Imam Khomeini (RAH) Shahre Rey Branch, Islamic Azad University, Tehran, Iran Rajeswari M. Kulkarni Department of Chemical Engineering, M. S. Ramaiah Institute of Technology, Bangalore, Karnataka, India Department of Chemical Engineering, National Institute of Technology, Surathkal, Karnataka, India P. Senthil Kumar Department of Chemical Engineering, SSN College of Engineering, Chennai, Tamil Nadu, India SSN-Centre for Radiation, Environmental Science and Technology (SSN-CREST), SSN College of Engineering, Chennai, Tamil Nadu, India Shinomol George Kunnel Department of Biotechnology, Dayananda Sagar College of Engineering, Bengaluru, Karnataka, India Bhanu Revathi Kurella Department of Biotechnology, Dayananda Sagar College of Engineering, Bengaluru, Karnataka, India Faisal Mahmood Department of Environmental Sciences and Engineering, Government College University, Faisalabad, Pakistan Zahid Maqbool Department of Environmental Sciences and Engineering, Government College University, Faisalabad, Pakistan Ali Mardy Faculty of Civil Engineering, K. N. Toosi University of Technology, Tehran, Iran Adeel Mehmood Department of Chemistry, GC University Lahore, Lahore, Pakistan Afsaneh Mollahosseini Research Laboratory of Spectroscopy & Micro and Nano Extraction, Department of Chemistry, Department of Chemistry, Iran University of Science and Technology, Tehran, Iran Gulzar Muhammad Department of Chemistry, GC University Lahore, Lahore, Pakistan Saima Muzammil Department of Microbiology, Government College University, Faisalabad, Pakistan Habibullah Nadeem Department of Bioinformatics and Biotechnology, Government College University, Faisalabad, Pakistan Mehrdad Negarestani Department of Civil and Environmental Engineering, Iran University of Science and Technology, Tehran, Iran Babatunde Oladipo Department of Chemical Engineering, Obafemi Awolowo University, Ile-Ife, Osun State, Nigeria Samuel O. Olatunji Department of Environmental Engineering, Texas A&M University-Kingsville, Kingsville, TX, USA

xiv

Contributors

Brahim Oudra Faculty of Sciences Semlalia, Laboratory of Biology and Biotechnology of Microorganisms, Environmental Microbiology and Toxicology Unit, Cadi Ayyad University, Marrakech, Morocco Ashok Pandey CSIR-Indian Institute for Toxicology Research (CSIR-IITR), Lucknow, Uttar Pradesh, India George Pavlidis Laboratory of Chemical Control of Pesticides, Department of Pesticides Control and Phytopharmacy, Benaki Phytopathological Institute, Athens, Greece Sreedev Puthur Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India K. Kavya Raj Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India Muhammad Arshad Raza Department of Chemistry, GC University Lahore, Lahore, Pakistan Sharrel Rebello Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India Ishika Saha Department of Chemical Technology, Rajabazar Science College, University of Calcutta, Kolkata, India Sana Saqrane Polydisciplinary Faculty, Laboratory of Chemistry, Modeling and Environmental Sciences, Sultan Moulay Slimane University, Khouribga, Morocco A. Saravanan Department of Biotechnology, Rajalakshmi Engineering College, Chennai, Tamil Nadu, India Dipanjan Sengupta Department of Chemical Technology, Rajabazar Science College, University of Calcutta, Kolkata, India Feride Severcan Department of Biological Sciences, Middle East Technical University, Ankara, Turkey Department of Biophysics, Faculty of Medicine, Altinbaş University, Istanbul, Turkey Biomedical Sciences Graduate Programme, Altinbaş University, Istanbul, Turkey Muhammad Shahid Department of Bioinformatics and Biotechnology, Government College University, Faisalabad, Pakistan Munazza Shahid Department of Chemistry (SSC), University of Management and Technology, Lahore, Lahore, Pakistan M. Shahvali Department of Agricultural Extension and Education, Agricultural College, University of Shiraz, Shiraz, Iran Tanvir Shahzad Department of Environmental Sciences and Engineering, Government College University, Faisalabad, Pakistan

Contributors

xv

K. Vidya Shetty Department of Chemical Engineering, National Institute of Technology, Surathkal, Karnataka, India Anuradha Shilli Department of Biotechnology, Dayananda Sagar College of Engineering, Bengaluru, Karnataka, India Muhammad Hussnain Siddique Department of Bioinformatics and Biotechnology, Government College University, Faisalabad, Pakistan Raveendran Sindhu Microbial Processes and Technology Division, CSIRNational Institute of Interdisciplinary Science and Technology (CSIR-NIIST), Trivandrum, Kerala, India Srishti J. Sisodiya Department of Biotechnology, Dayanandasagar College of Engineering, Bengaluru, Karnataka, India G. Srinikethan Department of Chemical Engineering, National Institute of Technology, Surathkal, Karnataka, India Memory Tekere Environmental Science Department, University of South Africa, Johannesburg, South Africa Adeline Su Yien Ting School of Science, Monash University Malaysia, Bandar Sunway, Selangor Darul Ehsan, Malaysia

Chapter 1

Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals Anca Giorgiana Grigoras

Contents 1.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2 Species of Pseudomonas with Ecological Potential . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.1 Sources of Pseudomonas-Type Sorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2.2 Chemical Compounds of Pseudomonas Responsible for Interaction with Toxic Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3 Biosorption or Bioaccumulation? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4 Kinetics of the Biosorption Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5 Isolated or Synergic Action of Pseudomonas Species? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.6 Toxicological Limits of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.7 Conclusions and Perspective Remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2 2 4 5 8 8 15 20 22 22

Abstract Since many industries are generating toxic heavy metal ions, the environment is continuously polluted involuntarily or voluntarily, in particular in order to increase the degree of comfort of people. Besides classical methods of decontamination, it should also be used the microorganisms in the process of environment greening. This chapter brings into discussion the studies regarding bioremediation of samples contaminated with only one type of heavy metal ions like lead, cadmium, or chromium and modeling of the biosorption process of heavy metals with the help of kinetic models and isotherm models. Also, it concerns about the synergic action of Pseudomonas species in the presence of other microorganisms or materials. The efficiency of new biosorbents designed with the help of Pseudomonas was evaluated in accordance with the toxicological limits regulated by the internationally authorized control bodies.

A. G. Grigoras (*) “Petru Poni” Institute of Macromolecular Chemistry, Iasi, Romania e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 1 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_1

2

A. G. Grigoras

Keywords Toxic heavy metals · Lead · Cadmium · Chromium · Bioremediation · Pseudomonas · Biosorption kinetic models · Isotherm models of absorption · Synergism · Toxicology

1.1

Introduction

The pesticides, heavy metals, and other organic compounds represent contaminants with known and unknown repercussions on life in general. Beside conventional physical and chemical methods, the biotechnology is an economical alternative solution to eliminate the toxic heavy metals from contaminated media. The biosorption process uses various materials with biological origin to efficiently treat high volume of waste charged with low but danger concentration of contaminants. Thus, in some attempts to treat wastewater and contaminated soils, the researchers have been used eggshells, flour of bones, carrot peels (Wang and Chen 2009), peat seaweed (Bulgariu and Bulgariu 2012), fungi, or yeast to remove the toxic heavy metals. The bacteria represent another low cost and large available biological material intensively used for the environment bioremediation. The intracellular or extracellular constituents and metabolites of microorganisms could act as metal chelators by physicochemical binding (Naik and Dubey 2013). In this paper, two aspects were considered: isolated or synergic action of Pseudomonas species as well as the studies of synthetic or real samples contaminated with a single type of heavy toxic metal like lead, cadmium, or chromium. After reviewing the main species, sources of Pseudomonas with ecological potential, the main chemical compounds responsible for interaction with toxic heavy metals by the biosorption or bioaccumulation mechanisms were inventoried. Also, the kinetics of biosorption process was discussed. Thus, equilibrium isotherm data resulted from batch biosorption experiments were analyzed using Langmuir, Freundlich, Dubinin– Radushkevich, Redlich–Peterson, and Temkin isotherm models. In addition, the experimental kinetic data were applied to the Lagergren pseudo-first-order, Elovich, pseudo-second-order, modified Ritchie’s second-order, and intraparticle diffusion kinetic models. All these data will be helpful to design new Pseudomonas-based biosorbents such that the toxicological limits of heavy metals to meet international standards.

1.2

Species of Pseudomonas with Ecological Potential

Because the environments in which they could be find are so diverse and new identification techniques like amplification of genetic material sequences and proteomic were developed, the taxonomy of Pseudomonas genus was continuously

1 Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals

3

revised over almost 50 years in accordance with physiological, phenotypic, or genotypic characteristics such that since 2009 more than 70 new species have been classified as Pseudomonas. The pace of discovery and classification of new species remains high, and it has been found that over the past 3 years there have been recorded ten new species per year (Peix et al. 2009, 2018). They are present in air, water, or soil and could be isolated from algae, fungi, plants, animals, compost, or sediments, being spread even in extreme environments like desert or the polar zone. Species of Pseudomonas belong to Pseudomonas genus, Pseudomonadaceae family, Pseudomonadales order, Gammaproteobacteria class, Proteobacteria phylum, and Bacteria kingdom. Pseudomonas species are aerobe Gram-negative and nonspore-forming bacteria with rod shape and bearing one or more polar flagella. According to the biochemical reactions used in their identification, most species are oxidase positive and catalase positive. Certain Pseudomonas species secrete pigments like pyoverdine (a fluorescent yellow–green siderophore), pyocyanin (blue pigment, a siderophore), thioquinolobactin (a yellow siderophore which become dark green in the presence of iron), pyorubrin (a reddish pigment), and pyomelanin (brown pigment). They are unpretentious, use a wide range of nutritional sources, and can grow on simple media like King Agar composed from agar, magnesium chloride, potassium sulfate, and protein hydrolysate. Selective Pseudomonas media containing inhibitors such as cetrimide, nalidixic acid, cephaloridine, penicillin G, pimaricin, or malachite green also can be used for isolation and presumptive identification (https://www.gov.uk/uk-standards-for-microbiology-investigationssmi-quality-and-consistency-in-clinical-laboratories; https://www.sigmaaldrich. com/technical-documents/articles/analytix/pseudomonas-media.html). Pseudomonas aeruginosa is an encapsulated, opportunistic, secreting piocyanin, multidrug-resistant pathogen which commonly affects immunocompromised patients. It is a strict aerobic, mucoid bacterium which grows in the temperature range of 5–42
C and has a characteristic grape-like smell produced by aminoacetophenone. Pseudomonas fluorescens and Pseudomonas putida belong to the fluorescent group of Pseudomonas unable to grow at 42
C. Pseudomonas fluorescens strains secret quinolobactin. Pseudomonas putida is a saprotrophic soil bacterium which not liquefy gelatin, used for bioremediation of naphthalene-contaminated soils or for conversion of styrene oil into biodegradable polyester-type plastics. Pseudomonas pseudoalcaligenes is a non-pigmented, mesophilic bacterium, with no distinct colony morphology, rarely recorded in clinical specimens, and biochemically inert compared with other Pseudomonas species. Using cyanide as a nitrogen source, it is a good candidate for bioremediation. Pseudomonas veronii is a fluorescent bacterium isolated from natural mineral waters, and it is also proper for bioremediation of contaminated soils. Environments contaminated with toxic heavy metals could be bioremediated with the help of Pseudomonas. Beside already presented species of Pseudomonas, others relatively newly discovered and recorded in National Centre for Biotechnology Information GenBank sequence database were the subject of the environment

4

A. G. Grigoras

cleaning of toxic heavy metals: Pseudomonas plecoglossicida (Kumari et al. 2015) and Pseudomonas taiwanensis (Majumdera et al. 2016), too.

1.2.1

Sources of Pseudomonas-Type Sorbents

Diverse soils, waters, and sediments are ordinary sources of Pseudomonas-type sorbents. These microorganisms, in them alive or non-living forms, were isolated and evaluated for potential applications in lead, cadmium, copper, chromium, mercury, or multiple toxic heavy metals bioremediation. Pseudomonas aeruginosa N6P6 (GenBank accession number KC771235, Culture submission no: BCCM/LMG 28185) has marine origin (Kumari et al. 2017a) and was useful for lead biosorption. Pseudomonas aeruginosa (LC102202), isolated from soil contaminated with electroplating industrial wastewaters, removed cadmium from the environment (Muneer et al. 2016). Strain 9# of Pseudomonas aeruginosa was isolated from rice fields contaminated with cadmium in Hangzhou, China, and applied for cadmium bioremediation (Lin et al. 2016). Pseudomonas aeruginosa PA1, a mercury-resistant strain, came from spiked drinking, river and lake water samples, and represented a source of carboxylesterase E2 for absorption of mercury ions (Yin et al. 2016). Non-living Pseudomonas fluorescens strains, isolated from collected sewage samples in Western Alexandria in Egypt, were a part from ternary nanoparticles used for biosorption of cadmium (Shaker 2015). Pseudomonas fluorescens SC106 strain collected from soil samples of a tannery industrial area at Misr Al-kadima (Al-Gayarah suburb) south Cairo, Egypt, or from the main drain for a mixture of both domestic sewage and tannery effluents was capable as monocultures and in consortia with another bacteria to remedy hexavalent chromium ions (Mawgoud 2015). Pseudomonas fluorescens Ps448 was a plant growth-promoting bacteria collected from soil samples of University College of Agriculture and Natural Resources, University of Tehran, located at Dolat Abad, Karaj, used for Eucalyptus cultivation and cadmium phytoremediation (Motesharezadeh et al. 2017). Pseudomonas putida I3 (GenBank accession No.KU990893), isolated from permafrost soil of Mohe wetland in Northeast China, was evaluated for its biosorption capacity of heavy metals from artificial wastewater (Wang et al. 2017). In another study, the living biomass of the same psychrotrophic strain I3 and non-living one were studied as biosorbents for Pb2+ removal under 15
C (Li et al. 2017). As a part of a biocomposite obtained through genetic recombination techniques, wild-type Pseudomonas putida Migula CCTCC (China Center for Type Culture Collection) AB92012 was evaluated for Cu2+ biosorption and the flocculation in aqueous solution in the presence of such ions (Hu et al. 2017). Pseudomonas pseudoalcaligenes NP103 bacterium isolated from marine water sample (Paradeep port, Odisha coast of the Bay of Bengal, India) formed a biofilm capable of interacting with Pb(II) ions (Kumari et al. 2017b).

1 Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals

5

Strain 2E of autochthonous bacterium Pseudomonas veronii, isolated from sediments associated to the Reconquista River basin (Buenos Aires Metropolitan Area), secreted some extracellular polymeric substances with the role of cadmium ligands for the bio-treatment of electroplating wastes (Ferreira et al. 2017). Pseudomonas plecoglossicida bacterium originated from acidic copper mine tailings of Xinjiang, China, while Pseudomonas taiwanensis was collected from the municipal sewage treatment plant of Birla Institute of Technology & Science, Pilani Campus, India. Both strains were tested for biosorption of multiple heavy metals from contaminated environments (Kumari et al. 2015; Majumdera et al. 2016).

1.2.2

Chemical Compounds of Pseudomonas Responsible for Interaction with Toxic Heavy Metals

Some of the chemical compounds responsible for interaction of Gram-negative bacteria with heavy toxic metals come from the structure of the cell wall of these microorganisms, and others are secreted by them and form a biofilm. Characteristic to Gram-negative bacteria are the following: the existence of outer membrane covered by a layer of lipopolysaccharides; location of a thin layer of murein in periplasmic space between outer and inner membranes; appearance of phosphatidylethanolamine and phosphatidylglycerol as predominant phospholipids in bacterial membrane. This heterogenic structure possesses amine, carboxyl, or phosphoryl functional groups placed in several discrete binding sites that ensure bacterial adhesion with various materials (Swoboda et al. 2010; Grigoras 2016). Bacteria also biosynthesize multiple glycoconjugates like lipopolysaccharides, peptidoglycan, glycoproteins, glycolipoproteins, capsules, teichoic acids, and exopolysaccharides such that the macromolecular composition of biofilm is heterogeneous, too (Hug and Feldman 2011). Generally, bacterial culture conditions, mainly the type of carbon source, will influence the composition and amount of exopolysaccharides. More specifically, nutrient deprivation, a lower temperature than the optimal one for growth, and a high carbon/nitrogen ratio in culture medium are key factors favorable for bacterial exopolysaccharide production (del Pilar Anzola-Rojas 2015). Peptidoglycan is a structural polymer, also found in the cell wall of Pseudomonas species, but in a thinner layer compared with the Grampositive bacteria, being structured in the form of a crystalline network of Nacetylglucosamine and N-acetylmuramic acid linear chains connected by β-(1,4)glycosidic bonds (Shaker 2015). Lipopolysaccharides dictate the surface charge or hydrophobicity of the cell. Architecture of a bacterial lipopolysaccharide consists from an O-specific polymer side chain, attached to the hydrophobic lipid A via a core oligosaccharide. Usually, lipid A is based on the 1,6-β-linked disaccharide of D-glucosamine carrying O- and N-fatty acyl substituents, and the inner core is composed from L-glycero-D-manno-

6

A. G. Grigoras

heptose and 3-deoxy-D-manno-pctulosonic acid. The outer region of the core oligosaccharides includes D-glucose, L-rhamnose, D-galactosamine, and possibly L-alanine (Wilkinson 1983). The group of self-generated extracellular polymeric substances, released into the growth medium, comprises exopolysaccharides, proteins, and extracellular nucleic acids able to improve the cell adhesion via dipole interactions, covalent or ionic bonding, steric interactions, and hydrophobic association (Grigoras 2016). Numerous researchers have been preoccupied with the composition of the slime produced by Pseudomonas species grown on different culture media (Bartell 1983; Jahn et al. 1999; Steinberger and Holden 2004). Commonly, as response of the environmental conditions, all Pseudomonas species produce slime with relatively the same chemical composition, but variable proportions of components. The polysaccharide fraction represents approximately 50–60% of the total slime and consists predominantly of glucose and small amounts of mannose. DNA and RNA form the nucleic acid material and are about 20% of total weight. In addition, there are small quantities of hyaluronic acid (5%) and minor components like protein (less than 5% from total), rhamnose, and glucosamine. Protein components of the biofilm matrix of Pseudomonas species include proteins with carbohydrate-binding capacity like CdrA protein, lectins, or amyloid-like protein. Rhamnolipids are surface-active glycolipids consisted from L-rhamnose and β-hydroxydecanoic acid moieties linked by an O-glycoside. Even that the specific biological role of these biosurfactants is still unknown, rhamnolipids are toxic to neutrophils and by their amphipathic-related characteristics facilitate the adaptation to the environmental stress through biofilm matrix formation and stabilization (Mann and Wozniak 2012). Microbial exopolysaccharides produced by Pseudomonas species include gellan (Banik et al. 2000), marginalan (Fett et al. 1995), capsular polysaccharides like alginates and levans, and aggregative polysaccharides like cellulose, Psl, and Pel. These compounds act like a protective shield which surrounds the bacterial cells. The spatial distribution of exopolysaccharides in the bacterial structure or in the medium surrounding them is related with the stage of growth and development of bacterial cultures. The role of each one in biofilm development and bacterial survival is determined by different properties, composition, and localization. Marginalan is an acidic galactoglucan with a succinyl group attached to a Dglucose residue, while levan is β-(2,6)-polyfructan-extensive branched through β(2,1)-linkages, synthesized from sucrose by a process catalyzed by levansucrase (Laue et al. 2006). Plant levan named phelins has recorded much lower molecular weight than bacterial levan (Rhee et al. 2002). Gellan gum is a water-soluble anionic polysaccharide with a tetrasaccharide repeating unit of two D-glucose, one L-rhamnose, and one D-glucuronic acid unit. The repeating unit of native polysaccharide is substituted with acyl groups like glyceryl and acetyl as the O-glycosidically linked esters. This high molecular weight polysaccharide produced by Pseudomonas cultures has gelling, emulsifying, and thickening properties with multiple applications in food, pharmaceuticals, and cosmetics industry (Jansson et al. 1983).

1 Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals

7

Mucoid and non-mucoid strains of Pseudomonas with clinical significance usually produce alginate. It is the best studied of the exopolysaccharides, a molecule resistant to degradation during phagocytosis or to depolymerization induced by free radicals produced by neutrophils and macrophages (Simpson et al. 1993; Anastassiou et al. 1987). Structurally, alginate is a negatively charged branched copolymer composed from D-mannuronic and L-guluronic acids linked by β-1,4 bonds in an irregular, nonrepetitive, and rigid pattern (Mann and Wozniak 2012). Usually, in the structure of a molecule, β-1,4 and β-1,3 linkages give an enhanced rigidity compared with β-1,2 and β-1,6 linkages. D-mannuronic acid is typically O-acetylated at C2 and/or C3. Alginate has multiple roles: like host immune system, it protects bacteria within biofilm, retains water and limits desiccation of biofilm, and increases bacterial resistance to antimicrobials by an overproduction of alginate (Chang et al. 2007). Being considered a capsular polysaccharide, it has a protective role rather than a structural function. The glucose-rich Pel polysaccharide and the mannose-rich Psl polysaccharide are encoded by Pel and Psl gene clusters, respectively. Psl, named as polysaccharide synthesis locus, and Pel are two loci or alternative polysaccharide-encoding genes involved in biofilm and pellicle formation of some Pseudomonas strains which contribute to adherence and aggregation. Studies related to viscoelastic properties of Psl and Pel exopolysaccharides revealed that Psl is alike an elastic material, being stiff or rigid, while Pel is more viscous (Jackson et al. 2004; Harmsen et al. 2010; Chew et al. 2014). Psl exopolysaccharide has a chemical structure still unknown, but it is supposed that it is a neutral exopolysaccharide composed from D-mannose, L-rhamnose, and D-glucose subunits (Byrd et al. 2009). Other authors state that Psl is rich in mannose and galactose. Psl works as a scaffold with relevant role in cell–cell signaling and cell–surface interactions (Ryder et al. 2007; Mann and Wozniak 2012). Bacterial species producing Psl have an increased ability to attach to glass, mucin-coated or epithelial cell surfaces, being involved not only in the initiation step of biofilm formation but even in maintenance and maturation of its structure. In the case of Pseudomonas aeruginosa, Psl has the capacity to bind with the protein part of the biofilm matrix represented by Cdr A, a protein with rod shape, β-helical tertiary structure, and carbohydrate-binding capacity, facilitating thus the specific interactions between Psl molecules and with the cells from biofilm matrix (Reichhardt et al. 2018). The chemical structure and signification of Pel exopolysaccharide are still unclear by some authors, but this newly discovered molecule plays a secondary role in initiation of biofilm formation. It has an important impact on cell adhesion to a surface and on ability to form a biofilm named pellicle at air–liquid interface. Pel locus encodes a glucose-rich matrix polysaccharide polymer, distinct from cellulose. In the presence of Pel, some Pseudomonas strains develop a three-dimensional structure and associate or aggregate in microcolonies (Ma et al. 2012). Using carbohydrate chemical analysis, some researchers discovered that Pel is a positively charged exopolysaccharide consisted from partially acetylated 1!4 glycosidic linkages of N-acetylgalactosamine and N-acetylglucosamine (Jennings et al. 2015).

8

1.3

A. G. Grigoras

Biosorption or Bioaccumulation?

As the adsorption and absorption are two distinct phenomena which occur in bacterial biofilm formation, the biosorption and bioaccumulation of metal ions are different processes, too, involved in interaction of Pseudomonas species with toxic heavy metals. The main difference consists in the way the contaminants are “sequestered.” If adhesion is an irreversible process, which follows Marshall model, implies electrostatic and van der Waals interactions, and follows the Derjaguin–Landau– Verwey–Overbeek approach, the adsorption supposes the existence of some strong binding forces (Bruscher and Van Der Mei 1997; Marshall 1992). Scientists work with living microorganisms or non-living biomass. Usually, the living biomass is cultivated on special microbiological media, e.g., Luria–Bertani medium. Instead, the non-living biomass is inactivated by autoclaving followed by drying and grinding. Exploring the heavy metal bonding capacities of both type of biomass, it was concluded that in case of dead biomass, only a biosorption process is involved, while in case of living biomass, it is carry out a process in two steps: spontaneously biosorption followed by bioaccumulation or intracellular accumulation. The bioaccumulation of toxic heavy metals can be highlighted by methods like scanning electron microscopy or fluorescent microscopy.

1.4

Kinetics of the Biosorption Process

Generally, the researchers realize the biosorption profile of a specific biosorbent using a batch experiment in order to analyze the environmental factors that influence the biosorption capacity of biosorbent such as pH, temperature range, biomass dose, contact or incubation time, and initial metal concentration. These factors modify the surface charge of biosorbents and thus affect the activity of functional groups of biomacromolecules which are an integral part of the biosorbent material. Information about the heavy metals adsorption rate and hydrodynamic parameters, useful for design of future effective biosorption process, could be extracted from adsorption kinetics. Usually, studying the graph of adsorption kinetic of any biosorbent, it is observed that biosorption capacity records few stages: in the first period of experimental time, the biosorption capacity dramatically increases and then gradually slows down until it reaches equilibrium. Also, during the process of biosorption, the number of available binding sites decreases with the increase of mass transfer resistance such that, finally, the biosorption capacity will be keep constant because all the biding sites present on the biosorbent are saturated. Kinetic mechanism of heavy metal biosorption could be investigated by linear fitting of experimental data with one of theoretical kinetic models: • Lagergren pseudo-first-order kinetic model • Elovich kinetic model • Pseudo-second-order kinetic model

1 Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals

9

• Modified Ritchie’s-second-order kinetic model • Intraparticle diffusion kinetic model Pseudo-first-order kinetic model appeals to the kinetic rate equation: ln ðqe ‐qt Þ ¼ ln ðqe Þ‐k 1 t

ð1:1Þ

where k1 (min1) is rate constant of pseudo-first-order adsorption and qt and qe (mg1) are biosorption capacity expressed as the amount of metal ions adsorbed at time t and equilibrium, respectively (Wang et al. 2017; Xu et al. 2017a; Shaker 2015). In case of the failure of pseudo-first-order kinetic model, it is used Elovich kinetic law of chemisorption and the equation with the same name: qt ¼

1 1 ln ðαβÞ þ ln ðt Þ β β

ð1:2Þ

with Elovich initial biosorption rate α (mg g1 h1) and activation energy of the chemisorptions β (g mg1) (Wang et al. 2017). Pseudo-second-order kinetic model presumes that the chemisorption is a rate limiting step. The linear form of the equation for this model is the following: t 1 1 ¼ þ t qt k2 q2e qe

ð1:3Þ

where k2 is pseudo-second-order rate constant of adsorption at equilibrium. Using Eq. 1.4, the initial adsorbent rate h (mg g1 min1) can be determined: h ¼ k 2 qe 2

ð1:4Þ

where k2 and qe have the same definitions as in Eqs. 1.4 and 1.1. The constant k2 is also involved in studying the effect of temperature on biosorption kinetic and allows the calculation of the value of the activation energy Ea from the regression of ln k2 and 1/T. Based on this value, it can be decided whether the biosorption process is physical or chemical one. For Ea lower than 28 kJ/ mol takes place a rapid and easily reversible physical biosorption based on weak forces, while stronger forces Ea that exceed the value of 43 kJ/mol will determine a chemical biosorption (Wang et al. 2017; Xu et al. 2017a; Shaker 2015). Modified Ritchie’s-second-order kinetic model can be expressed by the equation: 1 1 1 ¼ þ qt k R qe t qe

ð1:5Þ

and rate constant of the modified Ritchie’s-second-order kinetic model kR (min1) could be analyzed (Shaker 2015).

10

A. G. Grigoras

Intraparticle diffusion model is used as a supplementary option in the case that the pseudo-second-order kinetic model does not manage to predict the diffusion mechanism during the biosorption process. This model is given by the following equation which investigates if the diffusion plays the role of a controlling step in the biosorption process: qt ¼ kid t 0:5 þ Cid

ð1:6Þ

where kid (mg g1 h-0.5) is intraparticle diffusion rate constant and Cid (mg g1) is intercept of intraparticle diffusion at different stages (Wang et al. 2017; Xu et al. 2017a; Shaker 2015). In Table 1.1 are found the constants and correlation coefficients of kinetic models for biosorption of various toxic heavy metals using different living and non-living Pseudomonas sp.-type biosorbents. Comparing the values of correlative coefficients R2 of each kinetic model, the biosorption behavior of any sorbent is the best predicted or fitted by that model with the highest values. For instance, if R2 closed to value 1 corresponds to PSO kinetic model, then the chemical interactions between functional groups located to the surface of biosorbent and the metal ions are limiting control factor involved in rate of biosorption mechanism. Also, low values of pseudo-second-order rate constant k2 and activation energy of the chemisorptions β, but large values of Elovich initial biosorption rate α, indicate a higher affinity of heavy metal ions to the active sites of sorbents and a quicker and more suitable biosorption process of these ions. In addition, if the graphical representation of the equation characteristic to the intraparticle diffusion kinetic model is composed from multiple lines, each one with a different slope, then multiple rate constants kid values are recorded and assigned, for instance, to external surface adsorption and mesopore or micropore diffusion through the biosorbent material. When kid value for a region of the plot is much higher than kid values specific for other regions, but Cid values record an opposite trend, it is estimated that an external surface biosorption stage represents the rate controlling step. If some kid values are lower, but still positive ones, the intraparticle diffusion stage will perform a secondary and insignificant role in the adsorption process. Because experimentally it was observed that the adsorption capacity of a biosorbent increases with temperature increase, the adsorption can be considered an isothermal process that can be modeled by means of models. Thus, the isotherm models of absorption are based on the linear fitting of equilibrium experimental data of heavy metal biosorption on different sorbents and classified as the following: • • • • •

Langmuir isotherm model Freundlich isotherm model Dubinin-Radushkevich isotherm model Redlich–Peterson isotherm model Temkin isotherm model

Combination of 40% Pseudomonas putida I3 strain and xanthate-modified thiourea chitosan sponge for Pb2+ biosorption Biomass of psychrotrophilic Pseudomonas putida I3 grown on LB medium supplemented with 30 g/L K2SO4 for Pb2+ biosorption Ternary nanoparticles based on chitosan, gelatin, and non-living biomass of Pseudomonas sp. for Cd (II) biosorption Different experimental conditions (2 temperatures): 293 K and 323 K

Biosorbent with or without the synergic co-partner

30.30 64.75

189.479

qe, exp (mg g1)

13.26 14.71

–

0.330 0.378

–

0.81 0.74

0.1024

Pseudo-first-order (PFO) qe (mg g1) k1 (h21) R2 61.015 0.380 0.974

Elovich α (mg g1 h1) 3.26106 β (g mg1) 0.079 R2 0.989

–

31.03 66.75

63.69 (mg L1)

0.0756 0.126

4.92

Pseudo-second-order (PSO) qe k2 v0 (g mg1 h1) (mg g1) (g mg1 h1) 776.63 192.308 0.021

0.99 0.99

Shaker (2015)

Xu et al. (2017a)

Wang et al. (2017)

Ref.

(continued)

0.9997

R2 1

Table 1.1 Constants and correlation coefficients of kinetic models for biosorption of toxic heavy metals using living and non-living Pseudomonas-type biosorbents

Combination of 40% Pseudomonas putida I3 strain and xanthate-modified thiourea chitosan sponge for Pb2+ biosorption Biomass of psychrotrophilic Pseudomonas putida I3 grown on LB medium supplemented with 30 g/L K2SO4 for Pb2+ biosorption Ternary nanoparticles based on chitosan, gelatin, and non-living biomass of Pseudomonas sp. for Cd (II) biosorption Different experimental conditions (2 temperatures): 293 K and 323 K

Table 1.1 (continued)

1.0

–

39.79 69.75

25.4643

58.8 74.4

0.81 0.68

R12 0.974

Cid,1 (mg g1) –

kid,1 (mg g1 h0.5) 241.181

Intraparticle diffusion

2.2759

kid,2 (mg g1 h0.5) 18.2237

Cid,2 (mg g1) 137.594

0.9606

R22 0.960

0.1150

kid,3 (mg g1 h0.5) 3.059

Cid,3 (mg g1) 177.984

0.5247

R32 0.741

Shaker (2015)

Xu et al. (2017a)

Wang et al. (2017)

As useful tools for prediction of distribution of heavy metals ions in biosorbent material and metal ions spreaded in liquid phase, the equilibrium isotherms are indirectly related with the metal. Isotherm model is applied to a monolayer adsorption phenomenon which takes place onto a homogeneous biosorbent surface possessing a finite number of specific active sites endowed with identical energy. Equations related to the Langmuir theory are given as follows: Ce 1 C ¼ þ e qe K L qmax qmax RL ¼

1 1 þ K L C0

ð1:7Þ ð1:8Þ

with C0 (L mg 1) initial concentration, Ce (L mg 1) equilibrium or final concentration of residual pollutant ions, qe (mg g1) equilibrium amount of pollutant ions adsorbed onto 1 g of biosorbent surface, qmax (mg g1) maximum biosorption capacity, KL Langmuir constant or conditional biosorption equilibrium (affinity or saturation) constant, and RL separation factor. Generally, the equilibrium is established between metal ions adsorbed on biosorbent material and metal ions unabsorbed, but existing in solution. The inverse value of Langmuir constant KL is represented by the intercept of the plot Ce qe1 versus Ce. There are a number of favorable or unfavorable circumstances to the biosorption process which is depicted by the RL values. If RL is higher than 1, RL is equal with 1, RL lays between 0 and 1, or RL is 0, then the biosorption of heavy metals is unfavorable, linear, favorable, or irreversible, respectively (Wang et al. 2017; Li et al. 2017; Kumari et al. 2017b; Xu et al. 2017a, b; Shaker 2015). Freundlich isotherm model presumes that the uneven distribution of the absorption energy comes from a heterogeneous surface so that the adsorption process takes place in multiple layers. The linear form of Freundlich isotherm models a nonideal adsorption, and it is represented by the following equation: 1 log qe ¼ logK F þ logC e n

ð1:9Þ

where KF (mg g1) is adsorption capacity coefficient and n is Freundlich constant or adsorption intensity of the biosorbent. The Freundlich constants KF and n represent the intercept and slope of logarithmic plot between qe and Ce parameters. Also, their magnitude is related with adsorption efficiency. Usually, the values between 1 and 10 for Freundlich constant n indicate a favorable adsorption (Wang et al. 2017; Li et al. 2017; Kumari et al. 2017b; Xu et al. 2017a, b; Shaker 2015). Dubinin–Radushkevich isotherm model takes in account the difference between chemical and physical adsorption, and it is used to estimate the average free energy of biosorption E (Eq. 1.11) starting from the linear form of Eq. 1.10:

14

A. G. Grigoras

ln qe ¼ ln qDR ‐β ε2

ð1:10Þ

1 E ¼ pffiffiffiffiffi 2β

ð1:11Þ

ε ¼ RT ln ð1 þ 1=C e Þ

ð1:12Þ

where qDR (mg g1) is Dubinin–Radushkevich adsorption capacity coefficient, β (mol2 kJ2) activity coefficient related to mean biosorption energy E, and ε Polanyi adsorption potential. The values of E offer a preliminary idea about the dominant type of adsorption mechanism: chemisorption, namely, chemical ion exchange process (E between 8 and 16 kJ mol1), physicosorption (E lower than 8 kJ mol1), or particle diffusion (E higher than 16 kJ mol1) (Wang et al. 2017; Shaker 2015). Redlich–Peterson isotherm model is described by the nonlinear or linear forms in the following equations: K RP C e 1 þ αRP Cβe   K RP C e  1 ¼ β lnC e þ ln αRP ln qe qe ¼

ð1:13Þ ð1:14Þ

with KRP (L mg1) ¼ Redlich–Peterson isotherm constant, αRP (L mg1)β ¼ Redlich– Peterson isotherm constant, and β ¼ exponent that lies between extreme values 0 (becoming Henry’s law for infinite dilution) and 1 (becoming a Langmuir isotherm). Sometimes these three isotherm constants KRP, αRP, and β are unknowns because it is difficult to obtain a linear plot of ln (KRP Ce qe1)  1 versus ln Ce in Origin lab software. To overcome this situation, it could be applied a maximization procedure of the correlation coefficients R2 using the solver add-in function of Microsoft Excel (Shaker 2015). Temkin model postulates that the interactions between adsorbed entities, generically called adsorbate, and adsorbent material are the key factor responsible for linear decrease of adsorption heat with the coverage such that the binding energies evenly distributed will reach an energetic maximum. In this case, isothermal absorption is expressed by the equation: qe ¼

RT RT lnaT þ lnC e bT bT

ð1:15Þ

where bT (J mol1) is related to the heat of adsorption and aT (L mg1) is equilibrium binding constant which correspond to the maximum binding energy (Wang et al. 2017).

1 Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals

15

All these isotherm models are useful to elucidate the biosorption mechanisms and to describe the biosorption behaviors of toxic heavy metal ions onto various materials. Table 1.2 brings together the values for isotherm constants and correlation coefficients recorded for different Pseudomonas-type biosorbents with adsorption potential for toxic heavy metals. By comparing R2 values for each isothermal model, it is noticeable that the model with the highest values for R2 will best fit the experimental data. For instance, if the highest values R2 are recorded for Langmuir isotherm model, then the biosorption of heavy metals is considered a monolayer adsorption. There are situations where the experimental data of an absorption process is perfect fitted with two distinct isothermal models, for example, Freundlich and Langmuir isotherms. This situation suggests that the biosorbent surface is composed of both heterogeneous and homogeneous biosorption patches. In other cases, Langmuir kinetic takes place in two stages because the biosorbent material shows two distinct binding site types. In addition, the predicted values for maximum biosorption capacity qmax (mg g1) are pretty consistent with the values qe of the same parameter at equilibrium. Also, small values for equilibrium adsorption constant KL, recorded at the experimental studied conditions, imply strong binding of metal ions to the active sites from biomass.

1.5

Isolated or Synergic Action of Pseudomonas Species?

It could be taking in account the synergism of bacteria with other microorganism (bacteria or fungi) or even with other materials like natural polymers used by researchers to create new types of biosorbents. All these strategies were applied to improve the ability of biosorbents to remove the toxic heavy metals. The efficiency of synergistic or non-synergistic use of live microorganisms or biomass can be analyzed for each system. A solution of 40% Pseudomonas putida I3 strain was synergistic combined with xanthate-modified thiourea chitosan sponge, and the resulted biosorbent was tested on two samples: industrial effluent from the smelter and acid battery wastewater. Both samples contain in variable proportions different contaminants like Pb, Cu, Cd, Ni, Cr, Zn, Mn, Fe, Na, K, Ca, Mg, Cl, NO3, or SO42 (Wang et al. 2017). Kumari and the team (2017a) have succeeded to combine calcium alginate beads with immobilized form of extracellular polymeric substances from Pseudomonas aeruginosa N6P6 or with Pseudomonas aeruginosa biomass. They optimized the laboratory conditions such that to obtain a maximum removal of Pb(II) by designed biosorbents. The enhanced biosorption capacity of alginate beads modified bacterial extracellular polymeric substances compared with alginate biomass beads was due to the complexation of Pb(II) with matrix much rich in carboxyl, amide, and sulfhydryl groups than with the bacterial biomass poorer in reactive functional groups.

Combination of 40% Pseudomonas putida I3 strain and xanthate-modified thiourea chitosan sponge for Pb2+ biosorption Ternary nanoparticles based on chitosan, gelatin, and non-living biomass of Pseudomonas sp. for Cd (II) biosorption Different experimental conditions (2 temperatures): 293 K and 323 K

Biosorbent with or without the synergic co-partner

232.025

qe, exp (mg g1)

31.7 41.8

2102 3.9102

Dubinin–Radushkevich qDR β (mg g1) (mol2 kJ2) 215.506 1.92106

5 7.4

E (kJ mol1) 510.310

0.84 0.76

R2 0.591

0.99 0.98

0.09 0.08

Redlich–Peterson αRP KRP (L mg1) (L mg1)

0.89 0.86

R2

Temkin aT (L mg1) 1.09106 bT (J mol1) 219.34 R2 0.950

Shaker (2015)

Wang et al. (2017)

Ref.

Table 1.2 Constants and correlation coefficients of adsorption isotherm models for biosorption of toxic heavy metals using living and non-living Pseudomonas type biosorbents

Combination of 40% Pseudomonas putida I3 strain and xanthate-modified thiourea chitosan sponge for Pb2+ biosorption Living Pseudomonas sp. I3 for Pb2+ biosorption Non-living Pseudomonas sp. I3 for Pb2 + biosorption Extracellular polymeric substances from Pseudomonas aeruginosa N6P6 + alginate beads for Pb2+ biosorption Biomass of Pseudomonas aeruginosa N6P6 + alginate beads for Pb2+ biosorption Biomass of psychrotrophilic Pseudomonas putida I3

232.025

qe, exp (mg g1)

0.9990 0.9986

0.99 0.99

0.9873

0.1–0.2 0.1–0.2

– –

–

0.40 0.38

0.00077 0.00034

0.208

49.48 42.37

416.67 232.55

64.59

R2 0.9980

RL 0.007–0.404

Langmuir qmax KL (mg g1) (L mg1) 230.415 0.295

26.66

2.88 0.53

0.01441 0.00941

Freundlich KF (mg g>1) 155.221

5.571

0.88 0.76

3.66 3.01

n 15.873

0.8367

0.99 0.99

0.7420 0.8522

R2 0.979

(continued)

Xu et al. (2017a)

Kumari et al. (2017b)

Li et al. (2017)

Wang et al. (2017)

Ref.

Grown on LB medium supplemented with 30 g/L K2SO4 for Pb2+ biosorption Ternary nanoparticles based on chitosan, gelatin, and non-living biomass of Pseudomonas sp. for Cd (II) biosorption Different experimental conditions (2 temperatures, and two sites of reactions): (293 K; site A) (293 K; site B) (323 K; site A) (323 K; site B)

Table 1.2 (continued)

17.4 45.5 17.3 45.9

0.140 0.014 0.463 0.068

0.09 0.13 0.60 0.62

0.98 0.98 0.99 0.99

9.4 (293 K) 18.1 (323 K)

2.4 (293 K) 3.9 (323 K)

0.99 (293 K) 0.96 (323 K)

Shaker (2015)

1 Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals

19

Shaker (2015) prepared a biosorbent for removal of cadmium from aqueous solutions. Resulted biosorbent was composed from ternary nanoparticles based on non-living biomass of Pseudomonas fluorescens, chitosan as starting material, gelatin, glutaraldehyde as crosslinking agent, and silicon oil as oil phase. Due to the hydroxyl and amino groups of chitosan and to the bacterial peptidoglycan, the authors presume that ternary nanoparticles showed two sites of reaction with toxic heavy metal. This supposition was confirmed by the Langmuir isotherm of adsorption. Using these ternary nanoparticles, the cadmium removal capacity exceeded 90%, and thus the synergistic effect favor to this purpose was proved. Xu and collaborators (2017a) introduced a chemical additive, namely, 30 g L1 K2SO4, in the growth medium of psychrotrophic Pseudomonas putida I3 in order to modify the functional groups on the surface of bacterial cell. Initially, bacterial surface is rich in hydroxyl, carbonyl, amide, sulfonate, and phosphate groups. Using K2SO4 as additive, the potassium- and sulfur-containing groups increased on the surface of new biosorbent and favored the adsorption of Pb(II) mainly by chemical complexation and ion exchange, mechanisms in which the C¼O, O¼C–O, C–O/–OH, N–H/N–C, C¼N, and sulfur in sulfide and –SO3 functional groups were involved. Mawgoud (2015) exploited the ability of Pseudomonas fluorescens SC 106 and Bacillus subtilis SC106 strains to reduce the toxic hexavalent chromium from soil contaminated with tanneries effluent to less toxic derivate, namely, trivalent chromium. As Table 1.3 shows, the bioremediation efficiency in case of using the immobilized consortia of isolates was greater than for consortia or bacterial cells, and in this case, the synergism has played a positive role. In other cases, Pseudomonas has a secondary role in heavy metal remediation. The strains of Pseudomonas Ps36 or Ps448 were mixed with mycorrhizal fungus Glomus mosseae and then were inoculated in the growth medium for Eucalyptus Camaldulensis seedlings. In this way it was recorded an incremental effect on cadmium uptake by the plants compared to the non-inoculated (control) samples, Table 1.3 Remediation of chromium contaminated soil samples by individual bacterial cells, consortia, and immobilized cells of Pseudomonas fluorescens SC 106 and Bacillus subtilis SC106 (Adapted from Mawgoud 2015) Tested organisms Individual cells of Bacillus subtilis Individual cells of Pseudomonas fluorescens Consortia of Bacillus subtilis with Pseudomonas fluorescens Immobilized cells of Bacillus subtilis Immobilized cells of Pseudomonas fluorescens Immobilized cells of Bacillus subtilis and Pseudomonas fluorescens

Concentration of chromium after remediation (mg L1) 5.4 6.3 2.8

Efficiency (%) 99.05 98.89 99.5

3.9 4.6

99.31 99.19

2.1

99.63

Initial concentration of chromium ¼ 570 mg L1

20

A. G. Grigoras

and the complex system became a biosorbent with real potential in phytoremediation (Motesharezadeh et al. 2017). Lin and collaborators have investigated the effect of bacterial strains like Pseudomonas aeruginosa (9#), Stenotrophomonas acidaminiphila (2#), and Delftia tsuruhatensis (12#) on the growth of Oryza sativa L. in a soil sample contaminated with cadmium, in order to explore the synergistic or non-synergistic effect of microbes in the remediation of cadmium pollution. Thus, contrary to expectations, it was observed that Pseudomonas aeruginosa strain 9# was more effective in diminishing the accumulation of cadmium in rice grains compared with a mixture with the other strains such that Pseudomonas aeruginosa showed a non-synergistic effect (Lin et al. 2016).

1.6

Toxicological Limits of Heavy Metals

When a living organism comes into contact with heavy metals and subsequently cannot metabolize or excrete them, it is stipulated that, for each heavy metal, there is a limit or specific concentration of ions over which value the metal exhibits toxicity. Government bodies empowered to regulate the levels of toxicity considered normal and continuously update data for heavy metals, too, in collaboration with public health services. Thus, the permissible levels for heavy metals in different media are quantified in accordance with standards. According to the experts from the Center for Disease Control and Prevention in the United States, any exceeding of the reference level of lead in blood (5 μg dL1) as result of lead exposure through water, soil, food, paint dust, autoemissions, domestic mines, scrapped Pb-acid batteries, or air determines a serious injury of human health. Any exceeding of the allowed level of lead in blood of fetuses and young children causes intellectual impairment because this heavy metal is neurotoxic (Canfield et al. 2003). Since even for a lead level below 10 μg/dL recorded in adult blood, cardiovascular, and renal effects were recorded, the provisional tolerable weekly intake of 25 μg Pb per kg body weight per week stipulated by the World Health Organization in 2011 has been reset to a lower value (Rosen et al. 2017; WHO 2011). Being considered the most hazardous heavy metal, the World Health Organization and United States Environmental Protection Agency set the level allowed for lead in tap or drinking water which are 10 μg L1 and 50 μg L1, respectively (Wang et al. 2017). The lowest limits of detections achieved by different analytical methods in environmental media and diverse specimens are presented in Table 1.4 based on information for toxicological profile of lead provided by the Agency for Toxic Substances and Disease Registry, the Public Health Service, and the United States Department of Health and Human Services (ATSDR 2019). Cadmium is a pollutant from various industries that accumulates in the environment because it cannot be degraded. Over time, the researchers explore new methods

1 Pseudomonas Species for Environmental Cleaning of Toxic Heavy Metals

21

Table 1.4 Lead lowest limit of detection based on standards (adapted from ATSDR 2019) Media Air

Detection limit 1.5 ng cm2

Analytical method XRF

Drinking water

1.1 μg L1

ICP-AES

Surface water and groundwater Soil or sediment Whole blood or urine or tissues (postmortem)

0.07 μg L1 0.15 μg g1 0.05 μg Pb per g of blood 0.05 μg Pb per mL of urine 0.1 μg per 100 g blood 0.2 μg per g of tissue 0.1 μg g1

ICP-MS ICP-MS –

Whole blood or tissues (postmortem)

Animal tissue

References EPA (1999), Method IO-3.3 EPA (2003) Method 200.5 EPA (1997) NOAA (1998) NIOSH (1994b), Method 8003

–

NIOSH (1994a), Method 8005

ICP-MS or GFAA

NOAA (1998)

XRF x-ray fluorescence, ICP inductively coupled plasma, AES atomic emission spectroscopy, MS mass spectrometry, GFAA graphite furnace atomic absorption, EPA Environmental Protection Agency, WHO World Health Organization, NIOSH National Institute for Occupational Safety and Health, NOAA National Oceanic and Atmospheric Administration, United States Department of Commerce

to treat the industrial wastewaters contaminated with cadmium using inorganic or organic materials like zeolite, activated carbon, peanut shell, or bacterial biomass. Having a high toxic and carcinogenic potential, government bodies competent in health surveillance have imposed a maximum admissible limit of 0.001–0.1 mg L1 for cadmium concentration (Xu et al. 2017b). Chromium is a transition metal contaminant widespread in nature in two forms: soluble highly toxic anions Cr(VI) and less toxic and less soluble species Cr(III). The soils, waters, and sediments contaminated with traces of chromium usually resulted from industries such as pulp processing and wood preservation, leather tanning, electroplating, or steel manufacturing. Its mutagenic and carcinogenic characteristics necessitated the adoption of the maximum permissible levels in drinking water at 3 μg/dm3 for Cr(VI) and 100 μg dm3 for Cr(III) according to United States Environmental Protection Agency, while the World Health Organization standards imposed a maximum tolerance of Cr(VI) for public water suppliers of 0.05 mg/L (Mawgoud 2015).

22

1.7

A. G. Grigoras

Conclusions and Perspective Remarks

Dense matrix containing extracellular polymeric substances, secreted by Pseudomonas, is rich in biomacromolecules like lipids, proteins, and mainly in polysaccharides. Along with other structural polymers, all these biomacromolecules possess extremely reactive functional groups directly participating in the complexation with heavy metals ions. Biosorption process could be influenced by mechanisms like chemical reactions, diffusion control, mass transfer, and particle diffusion. The physical and chemical adsorptions are governed by forces that involve activation energies smaller than 28 kJ mol1 and larger than 43 kJ mol1, respectively. To better understand the kinetic mechanism of toxic heavy metals biosorption, the researchers involved different kinetic models and analyze the correlative or regression coefficients R2, rate constants k, the predicted qe values, and normalized standard deviations. The experimental data are consonant with the Lagergren pseudo-first-order, Elovich, pseudo-second-order, modified Ritchie’s second-order, and intraparticle diffusion kinetic models, and best fit to the experimental data was monitored. Langmuir, Freundlich, Dubinin–Radushkevich, Redlich–Peterson, and Temkin equilibrium models were used for computing the heavy metals binding efficiency of some biosorbents. Most adsorption isotherms recorded a good fit to the equilibrium experimental data and validate the enhanced potential and feasibility of different Pseudomonas-based biosorbents in sequestration of toxic heavy metals. The experimental equilibrium data for the toxic heavy metal biosorption at an optimum pH and specified temperature were applied to the one of the adsorption isotherm models to develop a mathematical relation which accurately represents the experimental data. The method of least squares is used for supplying the parameters of the models, while linear regression reveals the most fitted adsorption model. Researchers believe that the bioremediation of toxic heavy metals is more economical and safer if they use extracellular polymeric substances coming from non-living microorganisms because the energy consumption necessary for the biosorption of heavy metals ions is lower and the spreading of potential infections is limited.

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Chapter 2

Hardware and Software Remediation Technologies for Water Resources Pollution B. Abadi

and M. Shahvali

Contents 2.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1.1 Research Objectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1.2 Problem Formulation: Water Pollution of Dams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1.3 Water Contaminants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2 Management of Water and Sewage in Iran and the World . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.1 Indicators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.2 Wastewater Treatment Indicator: Environmental Performance Index . . . . . . . . . . . . . . . 2.3 Water Resources Pollution Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.1 Hardware System: Theoretical Basis of Technology-Based Pollution Management 2.3.2 Software Management of Pollution: Change in Scientific Intellectual Paradigms . . 2.4 Integrated Agenda for Management of Water Resources Pollution . . . . . . . . . . . . . . . . . . . . . . . 2.4.1 Ontology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.2 Epistemology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.3 Methodology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.4 Axiology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.5 Conclusion and Remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract The contamination of water resources in dam reservoirs risks the health of citizens and wildlife. In this chapter, it was intended to address multiple polluted dams in Iran, sources of water pollution, quantitative indicators of sewage development, wastewater treatment measures as a benchmark of environmental performance index worldwide, the review of the technological determinism and the social logic

B. Abadi (*) Department of Biosystem Mechanics Engineering, Faculty of Agriculture, University of Maragheh, Maragheh, Iran e-mail: [email protected] M. Shahvali Department of Agricultural Extension and Education, Agricultural College, University of Shiraz, Shiraz, Iran e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 27 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_2

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for technology, as well as need to mingle hardware and software remediation technologies concerned with water pollution. For more sustainable water resource, it is convinced that a paradigm shift should take place in the manner by which the respective stakeholders think of nature. Consequently, software technologies are defined in the context of scientific paradigms that are composed of the components of ontology, epistemology, methodology, and axiology. We cast a brief look at the viewpoint of each of the scientific paradigms to nature. Accordingly, an integrated agenda is delivered for water pollution management, which benefits from a fundamental scientific point of views. If paradigm shift to older styles of thinking about water resources come into notice, the remediation of pollution is likely to be possible, as would be the case in the study. This contribution would be a remarkable platform for experts, politicians, scientists, managers, and practitioners, as well as the public to disseminate hard and soft technologies in the field of sewage management and water pollution. Keywords Technology · Remediation · Hardware · Software · Water · Pollution · Resources · Agenda · Management · Paradigm

2.1

Introduction

The increasing population of urban and rural areas and the rise in living standards, which stem from the implementation of development schemes in many developing countries, has increased the demand for water resources, as consumed in agricultural, industrial, and household sectors (Arnell 1999; Vairavamoorthy et al. 2008; Dawadi and Ahmad 2013). Associated with the decisive dependency of humans and wildlife to water resources, the protection of these resources against contamination calls for remarkable actions necessarily taken by a wide variety of stakeholders. The presence of some soluble materials in water is essential for human health, but excessive levels of risk health. Nevertheless, heavy metals, such as lead, mercury, zinc, nickel, chromium, and cadmium, due to urban and industrial development, are flown into the soil by means of the inaccurate and non-sanitary disposal of urban and industrial wastewater, which can have long-term risks for ecosystems (Duruibe et al. 2007; Verma and Dwivedi 2013; Shakeri et al. 2015; Sarhangi et al. 2015). As an illustration, the water quality of the reservoir of dams is influenced by the features of the upper hands and water basin (AvazPoor et al. 2015). The entry of wastewater into reservoirs causes the contamination of water with various pollutants and also can lead to the proliferation of micro animals in the lake behind the dams. Rural wastewater is one of the sources of water pollution in dams. The importance of giving attention to this issue is that providing safe drinking water and controlling pollution for citizens is one of the most basic requirements. The lack of development

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of wastewater systems leads to the spread of environmental pollution (Azmi and Motiee Langroudi 2011).

2.1.1

Research Objectives

In this chapter, we initially exemplify multiple cases of dams at danger in Iran, the sources of pollutants of water resources, and quantitative indicators of sewage development in Iran and focus on wastewater treatment index as a benchmark of environmental performance index throughout the world. Subsequently, we concentrate on the need to regard hardware and software pollution remediation technologies. More specifically, we develop an integrated agenda of the management of water pollution applicable in practice. Therefore, the following questions would be addressed: 1. What are the quantitative indicators of water and sewage management in Iran and worldwide? 2. What are the sociological theories of technology and hardware systems? 3. What are paradigmatic viewpoints about the crisis of water pollution? To formulate an integrated agenda for managing the pollution of water resources, the following questions are outlined: 1. What is the nature of the realities about the occurrence of pollution, reasons, the context of the formation, and novel knowledge areas contributing to challenging with the crisis? 2. What is the physical and perceptual relationship of outsiders and insiders in the polluted environment generally and water pollution of dams particularly? 3. What are the deficiencies of past methodologies and epistemologies that cause the formation of the water pollution crisis, and how do complementary methodologies help cope with the crisis? 4. What role has past axiology played in shaping water pollution crisis and how alternative ones contribute to coping with the crisis?

2.1.2

Problem Formulation: Water Pollution of Dams

Sewage (i.e., gray water) from the rural houses together with excessive water from the irrigation of farmland, which contains pollutants, enter into the rivers flowing into dams (Mettetal 2019) as testimonies exemplify multiple endangered cases, the Siazakh Dam in Divandarreh, Kurdistan Province, Western Iran; Ekbatan Dam in Hamedan; Kamal Saleh Dam in Arak City, Gheshlag Dam in Sanandaj; Sardasht Dam in West Azerbaijan Province; Alavian Dam in Maragheh City, East Azerbaijan Province; Cham Gardalan Dam in Ilam province; and so on. The entry of sewage and wastewater from villages to Yalfan River leads to the water pollution of Ekbatan

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dam (Tasnim News Agency 2015). Drained agricultural pesticides and chemicals from Hamadan-Bahar Plain enter into the Hamdan watershed of Ekbatan Dam (Iranian Students’ News Agency 2015). In like manner, Arak City is another case, where more than 2400 cubic meters of wastewater is discharged in Kamal Saleh Dam (Tabnak 2016). Most often, the public expects authorities and government officials are responsible for launching environmental and health development projects through which, for example, rural sewage facilities and infrastructures are introduced in rural areas where collection and wastewater treatment facilities are lacking (Razmkhah et al. 2007). Despite the fact that government officials are obliged to do so, the effectiveness of the respective schemes emerge as local communities become involved in address and solving environmental challenges, which is called government to governance (Lockwood et al. 2010; Yazdanpanah et al. 2013). Unquestionably, this does not mean that the responsibility of governments is wholly shifted to local communities; shared policy initiatives are preferred. People may take unhealthy and non-environmental actions and believe that advanced technologies would modify the unpromising consequences of their actions. Although governments have the primary responsibility for supplying water and wastewater disposal (Stoutenborough and Vedlitz 2014), the role of the general public is crucial for the better management of these resources. Research shows that perceptions and attitudes of local people, for example, farmers, are very influential in water resources management (Abadi 2019; Abadi et al. 2018). In the area of wastewater treatment systems, for example, as the trash thrown by the public into pipelines linked to sewage systems, these actions would create huge economic costs while wastewater is treated. As the case may be, such persons think the respective technologies remedy the results of their actions. Of course, this is due to the scientific point of view dominated by universities and science centers because the scientific belief is that, if an ecological problem is remedied using technological means, the technologies also may lead to new challenges. In such a situation, the newer technologies treat the challenges created by earlier technologies (Bryant 2011).

2.1.3

Water Contaminants

There are a wide variety of sources that contaminate water resources (Harrison 2015; Shariat Panahi et al. 2008). The sources include septic wells, waste, agricultural drainage, geological formations, and mines. We describe the sources as follows.

Septic Wells Water contamination may occur through septic wells as the rapid transition of pollutant materials takes place owing to the steep slope of the ground, the low

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thickness of the alluvial layers, and the coarse grains in aquiferous systems. Sewage is discharged to the septic wells and then to rivers flowing into the dams. Meanwhile, biological contaminations from domestic wastewaters, heavy amounts of detergent, soot oil, and petroleum resources are introduced into surface water resources of the basin.

Waste The waste produced in the settlements and production centers located on the margin of the rivers are considered as threatening factors due to the lack of removal system. These materials are moved to rivers; rainfall and slopes of the basins give rise to rapid transition. The collected waste also creates an unfavorable situation due to improper and uncontrolled management.

Agricultural Drainage Agricultural waste contains large quantities of materials, mineral, organic matter, pesticide residuals, and chemicals. The drained water from farmlands also consists of the polluted materials. These materials are moved to the rivers, as the excessive water, that entails these materials, is drained from farmlands.

Geological Formations The geological formations in the catchment of dams are diverse. Formations with intrusive and outflow igneous rocks do not have any adverse effect on surface water and underground water resources in terms of chemical composition. But formations with coarse layers and possible sulfur, phosphate deposits, and gypsum are regarded as the most important sources of contamination.

Contamination from Mines Mines damage the environment in two stages. Firstly is at the exploration stage that explosive materials are used. Secondly is at the extraction stage; the aggregate of mineral materials sometimes on the margin of the river causes the introduction of these substances into the water.

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2.2 2.2.1

B. Abadi and M. Shahvali

Management of Water and Sewage in Iran and the World Indicators

According to a nationwide survey of 2003, the population of 69.9 million people in Iran, about 31% live in 68,000 villages. Of which, about 5000 villages (35.7%) have a population more than 200 households. About 0.6% of the villages in the country are covered by sewage collection and disposal services. According to the fourth development plan and the 20-year vision of the country, decision-makers have promised that the percentage of the rural population who will benefit from sewage treatment amounts to 30% in 2025 (Fahiminia et al. 2007). Water and wastewater management in Iran is undertaken by the municipal water and sewage companies, which are supervised by the Water and Sewage Engineering Company. This sector has defined a 20-year horizon (i.e., 2015–2035), the most important of which is the availability of 100% urban and rural population of the country to drinking water. Furthermore, it is predicted that sewage collection and treatment systems will be available for 60% of the urban population along with 30% of the rural population (Fahiminia et al. 2011). Table 2.1 shows the position of sewage disposal of Iranian cities in relation to 872 urban areas. The statistics of the measures and indicators of water and wastewater management sector at the end of 2015 in Iran (Information Technology Office of Power Ministry 2015) are displayed in Table 2.2. These measures are based on quantitative statistical frameworks. The framework includes the construction and equipping of infrastructures, including quantitative measurement, evaluation methods, surveys, and sampling.

2.2.2

Wastewater Treatment Indicator: Environmental Performance Index

Malik et al. (2015) deliver a global picture of the wastewater treatment indicator in detail, for example, percentage of wastewater treated normalized by connection rate for EPI in 2014, (i.e., the Environmental Performance Index) (Fig. 2.1). As reported, both the northern and southern hemispheres give an indication of disparate scores Table 2.1 Status of sewage disposal of Iranian cities Situation Sewage enters the river Sewage does not enter the river Reference: Fahiminia et al. (2011)

Number of city 312 560 872

Percent (%) 35.78 64.22 100

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Table 2.2 Measures and indicators of water and wastewater management in Iran at the end of 2015

Indices Number of dams (in operation by the Ministry of Energy) The volume of reservoirs of dams (million cubic meters) The adjustable annual volume of water The volume of storage water (total volume of reservoirs in operation circuit) (million cubic meters) The volume of water consumed (volume of sales of water) (million cubic meters) Total number of subscribers (million people) Percentage of drinking water consumption growth compared to the previous year (%) The length of drinking water distribution networks (km) The length of drinking water pipelines (km) Length of channels and drainage pipelines (km) Length of sewage collection network (km) Length of sewage transfer network (km) Number of water treatment plants in operation (unit) Number of sewage treatment plants in operation (unit) Population covered (%) Rainfall in the last year of the year (precipitation in 2014 (mm))

Until the end of 2015, Unit 361

Until the end of 2013, Unit –

Until the end of 2014, Unit 350

–

–

– 13.96

– 14.13

4236

4329

14,386,295 –

14.96 2.2

489,412

144,082

145,330

– – 489,412 27,519 124

27,671 39,099 51,147 2703 125

28,620 104,200 52,240 2850 128

–

157

160

99.1 20,592

99.1 217

99.1 184

48,373

14.29

4410 15.29 1.87

Reference: Information Technology Office of Power Ministry (2015)

among countries, although patterns appear at the regional level. The regions with the highest average score are situated in Europe (66.14
4.97), North America (50.42
17.44), Middle East and North Africa (36.45
6.33), East Asia and Pacific (27.06
6.91), Eastern Europe and Central Asia (18.34
5.40), Latin America and the Caribbean (11.37
2.51), sub-Saharan Africa (3.96
1.50), and South Asia (2.33
1.34). As shown, Iran has a position of 20–40%. Figure 2.2 also indicates the average percentage of wastewater treatment plants and connections to areas, which in Europe is higher than in other areas.

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Fig. 2.1 The level of wastewater treatment normalized by connection rate is used to calculate a proximity-to-target indicator for the 2014 Environmental Performance Index (Malik et al. 2015)

2.3 2.3.1

Water Resources Pollution Management Hardware System: Theoretical Basis of Technology-Based Pollution Management

Technological Determinism Technological determinism was introduced due to significant social and historical changes at the macro level, as well as profound psychological and social influences at the micro-social level referring to the systematic use of specific types of tools (Chandler 1996 cited in Gabberty and Vambery 2008). Kline (2001) takes account of two meanings for this term: (1) a technical and internal logic that determines the design of technical works and systems and (2) the development of the techniques and technological systems that determine widespread social change. He adds that these two meanings are in common with the claim that an autonomous technology forms social relations. The doctrine of technological determinism gives thought to the association of technology with society in a cause-effect, which contributes to creating the social systems and determining the limits of choices and changes, thus neglecting the contribution of any social factor (Zabet and Cheshmeh Sohrabi 2009). Jacques Ellul, the author of the book “Technological Society in the 21st Century,” and Marshall McLuhan developed this vision. Therefore, technological development is also done with the same internal logic and in accordance with the linear pattern of

0

25

50

75

connection

indicator

region East Asia and the Pacific Eastern Europe and Central Asia Europe Latin America and Caribbean Middle East and North Africa North America South Asia Sub–Saharan Africa

Fig. 2.2 Average percentages of wastewater treatment and connection rate in different regions, as bars reveal standard error (Malik et al. 2015)

Percent

100

treatment

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dissemination (Mehdiazdeh and Tavakol 2007). Given the utilization of hardware systems in pollution management, for instance, water and wastewater treatment technologies are used in the planning, coordination, and control of sewage management processes. It is also inclusive of the establishment of water resources laboratory networks with the aim of monitoring the quality of water resources, equipment, such as domestic water filtration, sampling, and measurement of the concentration of the main ions (HCO3, Na+, Ca2+, SO4(2-), NO3), total dissolved solids, PH, as well as the measurement of heavy metal ions of Cd, Cr, Ni, and Pb. As hardware technology, septic systems are widely used in rural areas and small communities (Struger et al. 2015). In Fig. 2.3, two examples of technological measures in sewage and wastewater treatment are presented.

Fig. 2.3 Pollution remediation technologies: septic systems in rural areas, upper photo (https:// slideplayer.com/slide/7746801/), and water treatment system, lower photo (https://slideplayer.com/ slide/10175993/)

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Social Technology Paradigm Contrary to the paradigm of technological determinism, which passes judgment on technology as the factor of forming social relations and structures, as also considers an internal logic of technology and economic requirement, the new perspective, called social technology paradigm, provides a view for the development of studies around the effects and consequences of technology (Mehdiazdeh and Tavakol 2007). Since the 1980s, scholars debated why users disapprove technologies and why it is difficult to use some technologies. Scholars elaborated the social logic of technology. Hence, the intellectual power of “powerful authority,” which is able to change the human world alone, gradually fell away, replaced with the newer one that believes in understanding the social destiny of any new technology; it is indispensable to study humans who produce and apply the technologies (Zabet and Cheshmeh Sohrabi 2009). Nevertheless, users have excessive authority, and therefore the requirements and constraints that users have with technologies are not regarded. Opposition to the theory of technological determinism appears in the works associated with the social construction of technology. Researchers in this area believe that the path to creativity (i.e., creation of new technologies) and the social consequences followed by technology are heavily influenced by society itself and factors, such as culture, politics, economic arrangements, and regulatory mechanisms. Accordingly, what matters is not the technology itself; more importantly is a social and economic system associated with the technology (Mehdiazdeh and Tavakol 2007).

2.3.2

Software Management of Pollution: Change in Scientific Intellectual Paradigms

Definition of Paradigm and Scientific Paradigm By paradigm, it is defined as an accepted model or pattern (Hedaa and Ritter 2005). In the context of the philosophy of science, the paradigm is a grand intellectual framework that governs a science or a theory. Truly, the scientific paradigm, as the universally recognized scientific achievements, provides a scientific community with a model for expressing a problem and solution in a given time period and shows the subsumed system of qualities (Zelichenko et al. 2016; Tuijnman 1989; Mina 2002). In the public definition, the paradigm can be attributed to the way individuals look at their internal and external environment, illustrated as an allegory in Figs. 2.4 and 2.5, irrespective of whether a person has the will and authority or whether the brain is the only data processing unit of the environment.

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Figs. 2.4 and 2.5 Individuals look into their surroundings differently, affecting their interpretation and giving meaning. Upper photo (https://www.janefriedman. com/distinguishingbetween-straight-up-adviceand-paradigm-shift/) and lower photo (https://www. pinterest.dk/pin/ 403283341618937003/)

Paradigm Shift The alteration of paradigm, implicitly called paradigm shift, is defined as a gradual change in the collective style of thinking, which involves changes in assumptions, values, goals, beliefs, expectations, theories, and knowledge, when the commonly used manner of thinking changes entirely (Terzidis 2004; Lee and Lee 2019; Simerly 1997). In compliance with water pollution, the way that scholars regard pollution is

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Fig. 2.6 Metaphor of puzzle of nine points, four lines, on the ineffectiveness of past thinking styles in solving current problems

gradually crystalized as a fixed reality due to having root in scientific intellectual foundations; when are ill-suited, the prior worldviews require a paradigm shift. Indeed, the newer paradigm, or a modified one, creates a new lens by which it is possible to look at wastewater management and the crisis of water pollution somehow different from older ones. A metaphor of paradigm shift is displayed in Fig. 2.6 (i.e., nine points, four lines). If one wants to draw four lines across nine points without removing pencil, the older styles of thinking to solve the problem are ill-advised, a paradigm shift is required to solve the puzzle, and the starting point is situated outside the framework. Similarly, to challenge with the pollution problem, newer methods of thinking are needed.

Paradigm Shift and Adoption The process of disseminating and adopting innovation is time-consuming; the different segments of the rural community do not behave in the same manner. Rogers has well documented the contribution of each of rural segments in accepting the changes. Bell-shaped curve, in blue, focuses on successive groups on innovation acceptance (i.e., innovators, early adopters, early majority, late majority, and laggards), and S-curve, in yellow, pertains to the diffusion of innovations among populations in various fields of science and knowledge; as the use of innovations increases (i.e., share of markets over time), the saturation point appears as laggards also adopt the respective innovations and changes (Nevalainen 2015) (see Fig. 2.7).

Types of Scientific Paradigms: The Intellectual Base of Scientists and Policymakers of Water and Wastewater Management Iman (2011) is convinced that for the expansion of scientific knowledge, “observation” and “thinking” are of two essential elements that influence the outcome of scientific knowledge. Scientific recognition is a knowledge, which examines how our events of life, both in the natural and human world, are created and the change and control of these events stem from the results of this scientific knowledge. He adds there are three types of knowledge of reality:

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Fig. 2.7 The diffusion of innovations with the five groups of people adopting an innovation (in blue), the increasing market share of the innovation (in yellow) (Nevalainen 2015)

1. Philosophical knowledge: the ontological and epistemological cognition of reality, derived from intuitive and intellectual knowledge 2. Paradigmatic knowledge: understanding the set of philosophical presumptions and principles that outline how the world works, resulted from intuitive, intellectual, and experimental knowledge 3. Scientific knowledge: a methodological process, which is involved with methodology (i.e., the boundary that sets separate philosophical from paradigmatic knowledge). Three types of knowledge are articulated, as shown in Fig. 2.8. In general, only knowledge that can discuss the performance of a problem is scientific knowledge. Intermediated between philosophical and paradigmatic knowledge, scientific knowledge enters into reality directly. To reach so, there is a need for a range of scientists, including philosophers, scholars, and those who produce statistics that are involved across the hierarchy. The point lines show that philosophical and paradigmatic knowledge cannot directly enter into empirical reality and must pass from the previous stages so as to move to the experimental reality. Figure 2.9 also indicates the role of the mixed methods of qualitative and quantitative methods, which is an indication of the importance of different scientific paradigms in creating scientific knowledge.

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Fig. 2.8 Three ways of creating knowledge (Iman 2008)

Fig. 2.9 Scientific paradigms (Niglas 2001)

Types of Scientific Paradigms The scientific paradigms are classified into the three groups. These enfold positivism, interpretivism, and criticism (Iman 2011). We elaborate the paradigms and the viewpoints of each paradigm to nature.

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Positivism Ontology The ontology of the positivist paradigm implies naive realism (Voros 2008; Ponterotto 2005). Assumed a perceptible fact that is derived from immutable natural laws, the paradigm gives credence to knowledge in the form of time- and- contextfree generalizations and cause and effect laws (Samhalle 2006). Basically, the center of attraction of research is the real state of affairs in the form of reductionism and determinism (Guba and Lincoln 1994). Moreover, human behavior is shaped in the form of a general and pre-written law that is based on causal relations and structural determinism (Iman 2008). In general, positivist paradigm believes in just an empirical observable material reality and takes into account the scientific method merely (Iman 2008). The assumption behind the paradigm is that credible knowledge is formed from atomization, decontextualization, subdividing of objects randomly assigned to observation, and synthesis (Bryant 2011). In the context of environmental knowledge, environmental assessments are conducted in the objective and quantitative manner (Rozema et al. 2012). As Francis Bacon, an English philosopher, states, if “the science” is perceived something for itself, the science will be an empty attempt made by the researchers; rather science must serve the purpose, that is, the goal out of the science itself. Therefore, paying attention to the knowledge of nature is also a means to reach the goal of the outside of nature itself (Hosseini Beheshti 2000). In order to have a positive image of phenomena, positivism makes the assumption of quantification (Bryant 2011), by which the goal of discovering, predicting, and controlling natural phenomena is achievable, and therefore universal laws, that enable people to change nature, are, therefore, developed. Positivism assumes that nature can be understandable, controlled, and dominated by humans by means of science, mathematics, and quantification (Bryant 2011). Bacon believes that science can make use of a technical force over nature; correspondingly, Auguste Comte talks about mastering nature and the ability to change and adapt the nature to fulfill human needs and desires, indicating disrespect and dominating over nature (Aminzadeh 2002; Gaukroger 2004). Therefore, the human thinks of the absolute owner of nature and give allowance for himself to be involved in changing and exploiting nature. Therefore, nature loses its intrinsic value, as would be turned into an inferior object; consequently, environmental crises emerge. As Bryant (2011) declares, it is assumed that ill-considered decisions have the potential to be corrected as the more advancement of knowledge, research, and technologies occur.

Epistemology The epistemology of this paradigm demonstrates objectivism and dualism (Voros 2008), which indicates the independence and separation of the researcher and nature

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(Shahvali et al. 2007). The epistemology enables the researchers to study nature without “influencing” or “being influenced” by the nature (Guba and Lincoln 1994). In positivism, environmental assessments are based on the technical–rational model and value–reality dichotomy; environmental studies are regarded as an applied science (Rozema et al. 2012). According to this view, humans are separate from nature; there is no possibility of combining society with nature (Brown 2011). Modernization is a model of development in which the relation of society with nature is represented by exploitative progress, which is achieved through continuous production, consumption of goods, and material service followed by far-reaching exploitation of resources (Shahvali 2013). Therefore, the separation of man from nature does not necessarily ensure environmental sustainability (Shahvali et al. 2007) because nature is referred to as “other.” In order to be a dominant player on the planet (Taylor 2011; King and McCarthy 2009), the human should torture nature (Heron and Reason 1997). Methodology The methodology of this paradigm is based on experimentation and manipulation (Heron and Reason 1997); questions and research hypotheses are presented in the form of propositions, tested whether be confirmed or not, and threatening conditions that are likely to affect the results of the research are controlled (Guba and Lincoln 1994). This paradigm follows a rational synthesis of data, assuming the conditions for comprehensive information, scientific certainty, and value consent with the contribution of technical and natural stakeholders. In the field of conservation of water resources, it is believed that only scientific research, as water experts and researchers do, gives legal ideas to solve water pollution. Environmental assessments are legalized on the basis of experts’ judgments (Rozema et al. 2012) and public participation, called “local systems of natural resource management,” and standing against these sides is discredited (Brown 2011). Therefore, just scientific methodology is able to correct mistakes, even mistakes made by scientific knowledge itself (Bryant 2011). Therefore, addressing nature through the use of western technologies is taken into account as a scientific insight for solving natural problems. Axiology Positivists eliminate values from science (House and Howe 1999); immaterial objects, such as values, love, and aspirations, cannot be easily and quantitatively tested and not capable of creating legitimate knowledge and legitimizing scientific research (Bryant 2011; Ponterotto 2005). In compliance with the paradigm, myths, religion, and personal experiences are considered to be the main components of common sense, which is non-scholarly knowledge, lacking order, and full of prejudices (Iman 2008). Therefore, to avoid these types of threats, there are some ways to reduce or eliminate the treats. Therefore, research is considered as a one-way mirror, the research should be prevented from being influenced by values and misunderstandings, and assumedly if strategies are followed accurately, repeatable

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findings are true (Guba and Lincoln 1994). In general, the propositional knowledge around the universe is regarded as a goal, which is intrinsically valuable (Voros 2008). The propositional form implies the recognition of conceptual terms that creates the knowledge derived from the description of nature, place, process, and objects (Heron and Reason 1997).

Interpretationism The interpretive paradigm, by accepting the difference in human phenomena with physical chemical-biological phenomena, denies reality outside the mind and gives regard to the outside world as the reflection of mental transitions. Based on the paradigm, research makes an attempt to convey the mental contents of individuals and to discover the meaning of interactions of individuals; generalization of findings is not followed (Iman 2011). Contrary to the positivist paradigm that tests a theory with quantitative data, qualitative data come into use in interpretative style to discover the meaning and abstraction of realities.

Ontology The ontology of the paradigm is relativism (Heron and Reason 1997). The facts are understandable in the form of multiple and unconscious mental constructs that are socially and empirically specific and local (Guba and Lincoln 1994). Meanwhile, the reality is socially constructed and there is no objective reality (Kuusipalo 2008). The social world is not independent of human consciousness, and therefore is not a predetermined reality, but reality is created through meaningful interactive actions of humans (Iman 2008). In the absolute sense, these mental constructions are not more or less correct, but more or less informative and complex, but can be changed (Guba and Lincoln 1994). Contrary to the positivist paradigm that is convinced that human behaviors is formed by a deterministic environment, the interpretive paradigm believes that human awareness, in comparison with other social factors, has a greater impact on social actions; basically, human empowerment plays a specific role in developing creation rather than being influenced by a predetermined context (Iman 2008). According to the ontology of the paradigm, the exploitation and protection of the environment must be seen in terms of values, subjectivities, and perspectives of the various and unintelligible individuals because there are several social realities that are independent of causal relations, and these facts are based on the experiences of individuals and their interactions with nature (Kuusipalo 2008). A scientific hypothesis is based on cognitive virtues and is intertwined with values and facts, and environmental assessments are considered as a civic science (Rozema et al. 2012). In this way, efforts are made to clarify and link the content of science in the context of social issues which involves the intellectual and practical flow of information that is done by scientists, politicians, citizens, and other stakeholders in order to better

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manage the supply, demand, and use of more credible information and knowledge (Weber et al. 2010; Scott 2011). Notwithstanding, knowledge networks are constituted by experts and community citizens, their information commonly used in interactive strategies of pro-environmental movements (Scott and Barnett 2009). Therefore, individuals, groups, and organizations have different attitudes toward nature that are complex in nature, but their knowledge will contribute to resolving environmental crises. To this end, the introduction of human values, environmental ethics, and social justice into the environmentalist paradigm requires a fundamental modification of scientific perspectives and methods (Jager et al. 2009). Epistemology In this paradigm, there is a transactional and subjectivist epistemology (Heron and Reason 1997). It is assumed that the investigator and subject matter are interactively interconnected, so that the findings, when done, are made literally (Guba and Lincoln 1994). In other words, knowledge is the result of interaction between the researcher and the people and object under cognition (Waters 2011). Like the critical paradigm, the distinction between ontology and epistemology does not appear. According to the paradigm, what is needed in scientific knowledge is the perception of the daily lives of people based on common sense knowledge (Iman 2008), which guides social interaction, and this type of knowledge is the basis for the correct understanding of society and social life. Common sense knowledge is a set of facts known by the majority of people in a group that includes the vast majority of human experiences in terms of spatial, physical, social, temporal, and psychological aspects associated with human daily experiences (Dias et al. 2009). Such knowledge is a shared sense and judgment of people (Claire 2003), which is associated with their daily perceptions (Young and Braziel 2006). Methodology The methodology of the paradigm is interpretive and hermeneutical (Heron and Reason 1997). In this paradigm, the reality is a social structure in which meanings are formed by social interaction with nature (Waters 2011). With interpretive and hermeneutical techniques, different constructs are formed and become transparent through dialectic transactions. The ultimate goal is to create a consensus-minded mindset that is more informed and sophisticated than previous mental constructs (Guba and Lincoln 1994). In general, this paradigm is inductive that deals with the symbolic discovery and description of special cases and provides an organized and valid abstraction of the particular cases that the researcher observes (grounded theory), which stem from social life (Iman 2008). Therefore, in the field of conservation of natural resources, the interaction of researchers with local people can be effective in identifying their mentality to resolve environmental crises, and thus, the mentality of different people must be transformed into a conscious and comprehensive mindset. This paradigm emphasizes the

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conservation of natural resources culturally, politically, economically, socially (Jager et al. 2009). Therefore, in environmental assessments, the transactions of social values that are considered in environmental sustainability considerations are based on public participation (Rozema et al. 2012). Axiology In this paradigm, values contribute to the formation of the research results (Waters 2011). The propositional, transformational, and instrumental knowing are tools for social emancipation and are useful and valuable, which is considered as a goal and inherently worthwhile (Voros 2008). Constructivism–interpretivism believes that the values and experience of the researchers cannot be separated from the research process. Researchers acknowledge and describe their values and should not exclude the values. Indeed, the epistemology of the paradigm recognizes the closest and more interpersonal contact with the research participants in order to facilitate their construction and expresses their experience of daily life, a false argument is that one person can exclude whose value orientations in the interactions of the researcherparticipant (Ponterotto 2005).

Critical Theory The critical theory was gradually developed after interpretationism. It was recognized that the mere interaction of individuals with each other is not able to create realities. But the historical contexts of these actions, rules, and laws are contributors to forming the realities. Ontology This paradigm, which stands in the face of positivism (Kuusipalo 2008), is based on historical realism ontology (Heron and Reason 1997; Ponterotto 2005). It is assumed that there is a comprehensible flexible reality, but it is shaped over time by social, political, cultural, economic, moral, and gender values and subsequently in the form of a series of consensual structures are regarded as reality (i.e., natural and unchanging) (Guba and Lincoln 1994). According to this paradigm, the reality is understood only by the people who participate in the research because there are no applicable rules explaining human behaviors (Kuusipalo 2008). According to this paradigm, nature is a function of social circumstances, and since culture and nature and their relationship vary, so the definition of nature changes over time (Shahvali 2013). Hence, nature includes a set of structural and historical insights that change over time. In this paradigm, criticism advocators seek to study and criticize the reality in order to make a fundamental change by resorting to philosophical concepts, such as freedom, truth, equality, and justice, and hence, people are encouraged to interpret the situation and build their better world (Iman 2008). More clearly, critical scholars seek to identify conflicting relationships in

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social reality, which has a type of domination and oppression. Meanwhile, nature involves humans, environments, and actions of both. In general, understanding the reality of nature and environmental crises is possible by identifying the views of the various stakeholders involved in environmental research (Shahvali 2013).

Epistemology The epistemology of this paradigm is based on transactional and subjectivism (Voros 2008). It is assumed that the researcher, in conjunction with his own values, and the subject of the research are in an interactive way that affects the research, and therefore the research findings are a type of value-based mediation (i.e., valuebased). This situation effectively challenges the distinction between ontology and epistemology, which is similar to the positivistic paradigm, and what can be identified is constantly intertwined with the interaction of researcher and the subject under study (Guba and Lincoln 1994). Nevertheless, the relationship between man and nature is interactive, and man is working with the consciousness to change and to escape the obstacles to the sustainability of nature (Shahvali 2013). Therefore, the separation of man and nature in the form that is seen in other paradigms is not regarded in this type of paradigm; humans usually have the opportunity to choose and change their surroundings (i.e., determination) (Kuusipalo 2008). In critical paradigm, some of the questions are established, such as how much cost appear to the nature and the poor as the control and domination of nature takes place (Bryant 2011).

Methodology Dialogue and dialectical represent the methodology of this paradigm (Voros 2008) because the bilateral nature of the researcher and what would be known requires a dialogue between the researcher and the subject under study. To overcome ignorance and misunderstandings (i.e., the adoption of undeniable intermediary historical structures), more informed consciousness is needed (i.e., how structures may be changed and understand that behaviors need to be effective change). Dialogue is in the form of dialectic, called transformative intellectual. The nature of the scenes is called “transformational intellectual.” Clarification and extraction of forms of historical knowledge that are relevant to conflict experiences and contradictions are inevitable (Guba and Lincoln 1994). Therefore, solving conflicts among stakeholders in the exploitation and conservation of natural resources can use this methodology. Transformational scholars show transformational leadership (Abadi 2016). This kind of leadership, in contrast to the transformational leadership, goes beyond the individual’s interests and leads to social exchanges and relationships in which this type of leadership follows a goal, a common approach to the development and promotion of motivation, guidance, and ethical ambition (Simola et al. 2010). Transformational leadership is based on the interaction of the leader with the

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followers and their needs and seeks to promote them to higher moral levels (Anderson and Sun 2017). Generally, this kind of leadership follows the four behavioral patterns of the role of charismatic role modeling, individualized consideration, inspirational motivation, and intellectual stimulation (Ying 2009; Hagiya 2011). Using charismatic role modeling, leadership inspires respect and loyalty and emphasizes the importance of having a sense of group and collective. With inspirational motivation, the leader is able to establish a relationship with the future aspiration and show off to his followers how to achieve prospective goals and make a difference in their beliefs to achieve the goals. Similarly, by intellectual stimulation, these leaders are able to promote their interests in their followers and stimulate them in relation to having a reflection on past issues to create new ways. These leaders use positive emotions to provoke followers. Axiology The propositional and transformational cognition achieve emancipation (Voros 2008). Value certainly affects the process and outcome of the research and, in this sense, one step ahead of the constructivists because the critical paradigm deals with the unequal distribution of power and oppression on controlled groups. A predetermined research objective is empowering participants to change their status quo and get out of dominant oppression, a value for libertarian thinkers (Ponterotto 2005).

Transcendental Paradigm Ontology Nature is defined beyond its objectivity; there is a virtue of spirituality for nature. In order to realize the transcendent place for nature, the tenacity research method is not sufficient; skepticism is accentuated (Abadi 2016). Moreover, according to transcendental wisdom, in addition to experiencing and observing outside, one must rely on the personal and internal experience that this experience should be accessible to all and not be available to researchers. Based on this paradigm, not only empirical observations and rational sciences in the research are the criterion of action but should also be the source of mysticism and ethics as sources of knowledge. In addition, many issues are not objective, and their analysis requires the use of mental components that may be based on individual experience (Shahvali et al. 2007). Epistemology In order to understand the nature, research starts with objectivity and focuses on the interaction of the objectivity and subjectivity, not stopping with the help of just the science. In order to make the reality more transcendent, other sources of knowledge

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are needed, such as sacred sources, traditions, mysticism, and revelation which also resort to the more varied cognitive resources that make the realization of the sustainability of nature more possible (Shahvali 2013). According to the transcendental paradigm, environmental studies are not only possible by describing material realities of the surrounding environment but the realization of a sustainable way of resolving environmental crises based on this paradigm which requires the use of empirical descriptions derived from observation and then their interpretation based on mental evidence, internal explorations. Moreover, the criterion of the effectiveness of scientific interpretations is that these interpretations are based on rational argument and reasoning, the actual revelations, and wisdom derived from divine mercy, while the scientific propositions and human findings are only limited to scientific theories, which is why environmental problems will surely be repeated (Shahvali et al. 2007). Hence, the redefinition of the nature and natural crises becomes important. In this case, the stagnation in environmental studies should be transformed into dynamism, and this dynamic in the maximal epistemology of the interaction of three types of horizontal, vertical, and diagonal dynamics is possible. The purpose of research with horizontal dynamics is to pay attention to environmental crises beyond the present and not only addressing the needs of the present generation and even meeting the expectations of future generations but also committing to the orders of the almighty creator of the world. Similarly, the meaning of vertical dynamics is to consider the issue locally and internationally. Finally, the purpose of the research is diagonal dynamics, attention to environmental crises in terms of appropriate expertise. Achieving this requires the use of human knowledge to integrate with natural and legal teachings that can bring unity of action to resolve environmental crises (Shahvali 2013). Methodology In the methodology of the transcendental paradigm, the environmentalist researcher is encouraged to draw up the future trend of these changes with a wise commentary, by analyzing the logical analysis of past environmental changes so far. In research based on this paradigm, not only attention to the results is important but also the way to achieve them, which is the proper use of reason, mysticism to achieve the results. Holism is also very important in the methodology of this paradigm because if science were to be able to solve environmental crises, science would have achieved this goal (Shahvali et al. 2007). Due to the diversity of the epistemological resources of this paradigm, the methodology requires to consider a range of methods for linking with nature. In this regard, the transcendental paradigm methodology requires that instead of using prescriptive methods in explaining the nature of environmental sustainability, participatory and collaborative methods in which a wide variety of stakeholder groups collaborate to each other to solve environmental problems. The transcendental methodology includes the transition from “peremptory” to “persuasive and consensual,” from “prescriptive to discourse,” from “observer to questioning,” and from

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“computing the results to the analysis and comparison of the results of the research” for the realization of sustainability of environment (Shahvali 2013).

Axiology The ultimate goal of paradigms, such as positivism, is propositional cognition, although interpretive paradigm stays with human flourishing as an ultimate goal, but does not match with the values of the almighty creator of the world. In this case, man’s knowledge should be linked with divine knowledge, which is monotheism. The value framework of divine source of ethics is regarded in this paradigm. Table 2.3 and Fig. 2.10 display a concise description of paradigms and the introduction of new paradigm.

2.4

Integrated Agenda for Management of Water Resources Pollution

In this section, we present an integrated model of hardware and software for managing water pollution. Accordingly, for the integrated agenda to be operational as possible, we will contextualize the newly defined paradigm for pollution management.

2.4.1

Ontology

The transcendental paradigm takes into account the nature of the reality of the pollution of water resources based on objectivism, subjectivism, as well as spirituality. Objective reality plays an important role in the management of water pollution. It is necessary to examine the current situation of the region facing the pollution crisis in order to determine the source of the facts about the causes of contamination and the context in which pollution appear, which give awareness to challenge with the crisis. For this reason, the following questions should be addressed: • Are there infrastructure and pollution control technologies in the area? • Have government agencies done sewage and wastewater treatment projects at local, regional, and national levels? • What areas are under protection? • What is the boundary between the area of sewage removal and intact areas? • What are the weather conditions and hydrological conditions, such as the degree of evaporation? • What is the type of soil composition, permeability, and geomorphological features?

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Table 2.3 Positivism, interpretivism, critical, and transcendental paradigm Critical theory Historical realism shaped by outside forces, material subjectivity

Ontology (i.e., what is the meaning of reality?)

Positivism Objectivist, findings are the truth, realism

Interpretivism Local, relative, co-constructed realities, subjective objectivity, realism

Epistemology (what is the relationship between the knower and what would be known)

The only knowledge is scientific knowledge, which is the truth; the reality is apprehensible

Co-created multiple realities and truths

Findings are based on values, local examples of truth

Methodology

Quantitative-primarily experimental, quasiexperimental

Often qualitative, sometimes quantitative

Usually qualitative, but also quantitative

Axiology

Free of values

Value-laden

Value-laden

Transcendental Nature is defined beyond its objectivity Taking into account spirituality in defining the nature Knowing and changing in nature is not done through universal laws and social norms, but the effect of god should be noticed Axiomatization to skepticism Observation, personal experiences, Islamic ethics Objective-subjective-transcendental resources of knowledge (bible scripts) Horizontal dynamic Vertical dynamic Diagonal dynamic Transdisciplinary A multiplicity of knowledge resources. Instrumental to persuasive, prescription to negotiation, observation to questioning and compute the research to analyze and compare the research results Value-laden, especially spiritual values

Reference: Guba and Lincoln (1994), Abisamra (1998), Ebrahimpour and Najari (2007), and Iman (2008)

Prescription / Command

(i.e., value nature)

Axiology

(i.e., epistemology way)

Methodology:

(i.e., knower and what known relation)

Epistemology

(i.e., reality nature)

Ontology:

Fig. 2.10 Software and hardware components of paradigm for water pollution management

Material-Spiritual Benefit

Negotiation (Dialogue), Consensus, and Persuasive

Materialism (Material benefit)

Insider-Outsider

ObjectivitySubjectivityMeta-physic

Just Outsiders

ObjectivitySubjectivity

Soft-ware water pollution remediation paradigm

Human-objectspiritually value

Result-Process-Based

Skepticism and Mingle(ism)

Simultaneous attention: Objects & Humans

Object-Human-value

Just Result-Based

Tenacity (Superstition) and Dualism

Sole attention: Objects

Soft-ware water pollution remediation paradigm

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• What is the prevailing cultivation in the region, the most commonly used poisons in the region, the traceability of pesticides residues and drainage water, the concentration of heavy metals and pH, toxic compounds, and organic matter? • Are there topographic maps of the region and aerial photography? • What are the consequences of sewage entry into water reservoirs of dams from ecological, sanitary, and social dimensions? • How many villages are covered by sewage management by government and private organizations? • Are water treatment and wastewater treatment regarded in the policies and perspectives of companies and executive agencies? • How are rural households in terms of water per capita? • What is the ratio of water conversion to sewage? • If there are sewage systems in the area, is rainwater mixed with household sewage? • What is the coefficient of exploitation of sewage collection network? Technology transfer and human resource development are closely interrelated (Swanson et al. 1997); as a result, the emphasis should be on the use of pollutionreducing technologies together with human resource empowerment. The use of managerial technologies can also have an impact on the reduction of water resources pollution. For this purpose, the establishment of a sewage treatment system will be beneficial. Nezakati (2015) refers to two examples of sewage systems, (1) centralized and (2) decentralized. Centralized systems are systems that, as a central site, include a central treatment plant, equipment for collecting, disposing, or discharging sewage for the whole of the area. In areas densely populated, these systems are adjacent to agricultural lands. These systems include a sewage collection network, a refinery system, a recovery system, and a sewage disposal system. If the treatment system is equipped with filtration and disinfection, the water can be used to irrigate the farmlands. The decentralized methods of decentralized systems, the health system, are installed and managed at the premises of a home; these systems can also be used to control the sewage of a group of rural houses that are close together. Sewage collection is carried out by several independent networks and transferred to several points outside the village. According to the transcendental paradigm, in addition to experiencing and observing outside, the realities should be extracted from the personal and internal experience that should be accessible to all and not be available to just scholars (Shahvali et al. 2007). In the context of the source of the pollution of water resources, it is necessary to consider the subjective constructs of local people, experts, and elites in the field of water pollution. This is because environmental crises derive from the epistemic crisis (Bryant 2011). For this reason, there are some questions: • What are the diverse viewpoints of rural people about the pollution of water resources and contaminating sources? • What are the diverse views of experts and policymakers about contaminating water sources and contaminating sources?

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Epistemology

A transcendental paradigm takes into account the intermixture of individuals and stakeholders with the realities of water resources pollution. For this reason, all individuals, scholars, elites, and local people are involved in solving the crisis. In addition to commonly used interventions and communications programs (e.g., awareness programs and the use of public participation in rural waste management programs), socio-cultural factors should be considered. The villagers become involved in being aware of the problem. For example, any sewage management program actually needs public participation, so local communities play an important role in the development of the objectives of the sewage management process. In order to institutionalize the pattern of efficient water use and sewage production, the implementation of educational programs is vital. In general, it is noteworthy to address the following questions: • • • • •

What is the relationship between the representatives of rural people and experts? How many experts and elite people visit polluted areas? How much do experts speak about water pollution from dams in news and media? How much do newspapers and local newspapers talk about water pollution? What are the organizations of the people in the region in the area about sewage management and the establishment of local refineries? • What is the level of pressure from people near the dam to set up sewage management plants? • What is the status of participation of villagers in collective activities at the rural level? • What is the status of participation of villagers in collective measures at the regional level? Therefore, multiple and different views of humans must be considered. Therefore, focusing on bilateral learning is important, which indicate how knowledge can be created in special situations (Waters 2011). Because the meaning of humans is formed through their interactions with each other and the environment around them, learning is the result of their interaction with social activities (Kuusipalo 2008). In order to ensure continuous and long-term epistemology in identifying and challenging with the pollution of water resources of dams, it is necessary to monitor the effectiveness of pollution remediation technologies, for example, sewage removal or wastewater treatment systems. Continuous monitoring of the functioning of the equipment should be in the form of regular monitoring programs. To optimize and maintain wastewater management, the use of local people and experts’ skills is necessary (Nezakati 2015). Moreover, regular local meetings with experts are a vital issue.

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Methodology

In order to further develop knowledge about the water pollution crisis, sources of knowledge, such as sacred sources, tradition, mysticism, and revelation, are used, and thus, cognitive resources are far more diverse, making nature more sustainable (Shahvali 2013). Due to the diversity of the epistemological resources of this paradigm, the transcendental paradigm requires a methodology to consider a range of methods for linking with nature. In this regard, the transcendental paradigm methodology requires that instead of using from “peremptory” to “persuasive and consensual,” from “prescriptive to discourse,” from “observer to questioning,” and from “computing the results to the analysis and comparison of the results of the research.”

2.4.4

Axiology

Although positivist researchers precisely prevent values that bias the research results, values are naturally reflected in the choice of the subject under study, for example, the decision of a researcher to study poverty may be due to his sense of social commitment (Ponterotto 2005). In the context of water pollution, propositional and transformational valuation is not sufficient, but spiritual valuation should also have regarded, especially religious values are a part of local people. The value considered can be a criterion for achieving a favorable situation for water resources, which would be in favor of not just human but also natural habitats.

2.5

Conclusion and Remarks

This chapter focuses on the integration of software and hardware technologies to better manage water resources pollution. While reviewing technology theories and providing evidence and statistics on the issue of pollution management at national and global levels, scientific paradigms were introduced as a platform for discussing the causes of water pollution. Since the dominant view in the scientific community of the world is positivism, the defective definition of the environmental reality gives rise to the pollution of water resources, and therefore the inadequate understanding of pollution is created by the respective methodology and value of this type of thinking about the environment because the environmental instability has not yet been resolved. Thus, it is imperative that all scientists, politicians, decision-makers, scholars, and faculty members in the universities, which deal with the study of the environment and environmental management, need to change the way in which these people look

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Table 2.4 Link between integrated agenda and the steps of management model The steps of water pollution management and paradigm components Planning Implementation Evaluation and control

Ontology ✔

Epistemology ✔

Methodology

Axiology

✔ ✔

Reference: Research finding Note: Tick symbol indicates that the component of the paradigm can be used in the respective stage of management

at the nature and water sources; therefore, the change will take place in their intellectual foundations. Therefore, contamination of water resources is an objective reality that cannot be managed solely by technology and scientific methods, especially as these technologies will also create other results that are challenging. Furthermore, the close relationship between scientists and all local communities is a commitment to better management of environmental pollution. Restricting a stratification of the local community will create an incomplete knowledge for pollution management. It is also recommended to use a variety of pollution study methods, such as quantitative and qualitative methods. From the point of view of value, it is also necessary to identify water resources pollution not only for propositional knowledge but also to create a cognition to improve the lives of local and remote people. In general, remediation water pollution technologies should include, for example, the establishment of laboratories, the addition of chemicals to neutralize heavy materials, drinking water treatment plants, and software ones in the form of scientific paradigms (i.e., paradigm shift). At the level of water resources management and contamination of these resources, it is also possible to communicate between the components of the paradigm and stage of management of polluting, which is displayed in Table 2.4.

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Scott D (2011) The role of a spatial civic science in repositioning scientific expert knowledge in society: a case from South Africa. Available at http://www.chance2sustain.eu/fileadmin/ Website/Dokumente/Dokumente/Publications/Chance2Sustain_Opinion_No_2__The_Role_ of_a_Spacial_Civic_Science_in_Repositionning_Scientific_Expert_Knowledge_in_Society. pdf Scott D, Barnett C (2009) Something in the air: civic science and contentious environmental politics in post-apartheid South Africa. Geoforum 40:373–382 Shahvali M (2013) Explanation of transcendental paradigm of agricultural innovation system. In: Proceeding of the second conference on Islamic-Iranian development model. National Library, Tehran, p 101 Shahvali M, Keshvarz M, Sharifzadeh M (2007) Transcendent ethical-philosophical paradigm in environment crises researches. Ethics Sci Technol 2:31–44 Shakeri A, Shakeri R, Mehrabi B (2015) Assessment of As, Cd, Ni and Cr contamination in water, sediments and fish of Shahid Rajaie dam, North Iran. J Environ Stud 41:13–24 Shariat Panahi E, Jabbari M, Fathi A (2008, March 5) Investigation of the environmental impacts of Latian dam basin (with emphasis on the effect of geological and geomorphologic factors on water quality). In: 3th applied geology and environment conference. Azad University, Islamshahr city Simerly RG (1997) Preparing for the 21st century. J Contin High Educ 45:38–51. https://doi.org/10. 1080/07377366.1997.10400328 Simola SK, Barling J, Turner N (2010) Transformational leadership and leader moral orientation: contrasting an ethic of justice and an ethic of care. Leadersh Q 21:179–188 Stoutenborough JW, Vedlitz A (2014) Public attitudes toward water management and drought in the United States. Water Resour Manag 28:697–714 Struger J, Van Stempvoort DR, Brown SJ (2015) Sources of aminomethylphosphonic acid (AMPA) in urban and rural catchments in Ontario, Canada: glyphosate or phosphonates in wastewater? Environ Pollut 204:289–297. https://doi.org/10.1016/j.envpol.2015.03.038 Swanson BE, Bentz RP, Sofranko AJ (1997) Improving agricultural extension. FAO Press, Rome. Available at http://www.fao.org/DOCREP/W5830E/w5830e00.htm#Contents Tabnak (2016) A daily entering of 2,400 m 3 of sewage into Kamal Saleh dam. News code: 334192 Tasnim News Agency (2015) The sewage of the villages upstream of the Ekbatan dam enters the dam. Available at https://tn.ai/898304 Taylor PW (2011) Respect for nature: a theory of environmental ethics. Princeton University Press, Princeton Terzidis K (2004) Algorithmic design: a paradigm shift in architecture? In: 22nd eCAADe conference proceedings, Copenhagen (Denmark) 15–18 September 2004, pp 201–207 Tuijnman A (1989) Paradigms and education professors—comments on scientific foundationalism. Interchange 20:62–67 Vairavamoorthy K, Gorantiwar SD, Pathirana A (2008) Managing urban water supplies in developing countries–climate change and water scarcity scenarios. Phys Chem Earth Parts A/B/C 33:330–339. https://doi.org/10.1016/j.pce.2008.02.008 Verma R, Dwivedi P (2013) Heavy metal water pollution-a case study. RRST 5:98–99 Voros J (2008) Integral futures: an approach to futures inquiry. Futures 40:190–201. https://doi.org/ 10.1016/j.futures.2007.11.010 Waters A (2011) A life in environmental education: an autobiographical case study. Dissertation in philosophy, Deakin University Weber EP, Leschine TM, Brock J (2010) Civic science and salmon recovery planning in Puget Sound. Policy Stud J 38:235–256 Yazdanpanah M, Thompson M, Hayati D, Zamani GH (2013) A new enemy at the gate: tackling Iran’s water super-crisis by way of a transition from government to governance. Prog Dev Stud 13:177–194

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Chapter 3

Anaerobic Biotechnology for the Treatment of Pharmaceutical Compounds and Hospital Wastewaters Ali Khadir

, Afsaneh Mollahosseini, Mehrdad Negarestani, and Ali Mardy

Contents 3.1 3.2 3.3 3.4

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Aerobic and Anaerobic Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Anaerobic Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Common Anaerobic Biotechnologies for Pharmaceuticals Removal . . . . . . . . . . . . . . . . 3.4.1 Anaerobic Digesters (ADs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4.2 Anaerobic Membrane Bioreactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4.3 Up-Flow Anaerobic Sludge Blanket . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4.4 Anaerobic Sequencing Batch Reactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

62 65 66 68 68 69 73 76 78 78

Abstract The frequent occurrence and spread of pharmaceutical compounds in the environment have made their resources a primary public concern. Hospital wastewaters and pharmaceutical production companies release various pseudo-persistent toxic compounds in the environment that may endanger the life of human beings as well as animals, plants, and soil. Since now, various physical, chemical, and biological methods have been employed for their elimination. In most of the

A. Khadir (*) Young Researcher and Elite Club, Yadegar-e-Imam Khomeini (RAH) Shahre Rey Branch, Islamic Azad University, Tehran, Iran e-mail: [email protected] A. Mollahosseini Research Laboratory of Spectroscopy & Micro and Nano Extraction, Department of Chemistry, Department of Chemistry, Iran University of Science and Technology, Tehran, Iran e-mail: [email protected] M. Negarestani Department of Civil and Environmental Engineering, Iran University of Science and Technology, Tehran, Iran A. Mardy Faculty of Civil Engineering, K. N. Toosi University of Technology, Tehran, Iran © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 61 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_3

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current/conventional wastewater treatment plants, biological sections are the heart of the treatment plants that play the primary role in the degradation of pharmaceutical compounds. Biological treatment methods are generally categorized into main groups of aerobic and anaerobic techniques based on the presence or the absence of the oxygen. Taking into account that these wastewaters own high organic content, anaerobic methods have been candidated as sustainable, low cost, low nutrient requirement, and low-area demand and efficient ways for pharmaceutical degradation. Anaerobic digesters, membrane bioreactors, up-flow anaerobic sludge blanket, and anaerobic sequencing batch reactor are the chief techniques. This chapter tries to firstly bring out the removal pathway during anaerobic digestion and then presents the works have been done for pharmaceutical removal. Keyword Anaerobic digestion · Hospital · Pharmaceutical · Membrane, sludge blanket

3.1

Introduction

Rapid population growth, extensive industrialization, and agricultural activities have led to water contamination with various organic/inorganic compounds. Accordingly, access to safe water has become a significant concern in developed, developing, and water-scarce countries (Mirjavadi et al. 2019; Piri et al. 2019; Tshindane et al. 2019). Every year, nearly 2.2 million people die because of exposure to contaminated water (Zhao et al. 2019). No living creatures could survive without water. Under these circumstances, water and wastewater are two chief sources for both clean water supply and prevention of environmental pollution. Among various municipal and industrial wastewaters that currently are treated, pharmaceuticals (or hospital) effluent could pose serious matter. Pharmaceutical wastewater contains multiple organic and inorganic components such as spent solvents, catalysts, additives, and reactants. Table 3.1 shows the characterization of various pharmaceutical wastewaters. These effluents own various fluctuations in physical, chemical, and biological features based on their production source. For instance, herbal pharmaceutical factory (Nandy and Kaul 2001) produces an effluent having chemical oxygen demand of 5000
80000 mg/L; however, of penicillin G pharmaceutical industry (Rodríguez et al. 2005) or bulk drug manufacturing industry (Sreekanth et al. 2009), it is 12500
1070 mg/L and 34400
2000 mg/L, respectively. Apart from high chemical oxygen demand values and their fluctuations throughout a day, the organic matters are considered to be not rapidly degradable, which signify that not every technique could be utilized to treat these wastewaters. The high value of total nitrogen (Shi et al. 2014) is a clear indicator of the fact that N-containing compounds are used as the raw material in pharmaceutical and drug manufacturing companies. Also, total phosphorous concentrations were of

a

–

–

3500

430

–

360

4300–74000

188.3

–

30–120

–

–

5000–80000

18.6

680

6.8

4.2–4.5

– –

8.5

6–7

7.5

–

–

1400

7.5
0.3

–

871
87

38
1.9

605 (N-NH4+) 135–1250

1250
25

12500
1070

6.42

–

–

420

72200

2570

20140

7–8

29450
1209

199
59

7–7.5

7–8

pH 5–5.9

–

251.8

1612
353

16249
714

8500–9000

22168
3757

Total dissolved solids (mg/L) –

2800–3000

388
87

Total suspended solids (mg/L) 100–400

100–120

1763
3.6

Total phosphorous (mg/L) 1–2

6800

120–170

13000–15000

1422
173

15365
1214

Total nitrogen/total Kjeldahl nitrogen

Herbal pharmaceutical wastewater Pharmaceutical wastewater

Fermentation-based pharmaceutical wastewater Pharmaceutical wastewater

Bulk drug pharmaceutical industry Antibiotic production (Singapore) Citric acid wastewater in Anhui province, China Penicillin G pharmaceutical industry Gulab Devi Chest Hospital

Pharmaceutical wastewater Traditional Chinese pharmaceutical wastewater Antibiotics waste

TN/TKNa (mg/L) 60–68

Chemical oxygen demand (mg/L) 45006500

Table 3.1 Characterization of various hospital/pharmaceutical wastewaters

Ng et al. (2014) Sreekanth et al. (2009) Ng et al. (2015) Chen et al. (2014) Rodríguez et al. (2005) Meo et al. (2014) Xing et al. (2014) Buitron et al. (2003) Nandy et al. (1998) (Thakura et al., 2015)

References Chen et al. (2019)

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Fig. 3.1 Main sources and pathways of pharmaceuticals entrance to water media like groundwater in the environment

considerate in some cases (Ng et al. 2014). A proper treatment method must be able to remove total nitrogen and total phosphorous simultaneously. Figure 3.1 shows the main pathways of pharmaceuticals entrance to water media. Among them, the role of treatment plants is of high consideration (Ghenaatgar et al. 2019). Countless studies have shown that pharmaceuticals are entering the wastewater treatment plants and water treatment plants by different ways that hospital and clinical healthcare facilities effluents are the main sources of this issue. In water sources of Brazil, 18 pharmaceuticals were detected in ng/L levels (Reis et al. 2019). Regarding drinking water of Putrajaya (Malaysia), amoxicillin, caffeine, chloramphenicol, ciprofloxacin, dexamethasone, diclofenac, nitrofurazone, sulfamethoxazole, and triclosan were quantified (Praveena et al. 2019). Also, pharmaceuticals were detected in lakes (Nantaba et al. 2020). As a matter of fact, it seems that current wastewater treatment plants are not appropriate for the thorough elimination of these compounds and consequently contribute to the spread of them into the environment (Couto et al. 2019). Biological processes are the most important part of any wastewater treatment plants and it is highly necessary to operate the biological sectors under the optimum conditions to attain maximum removal efficiency. Since pharmaceuticals have a

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significantly high organic content, anaerobic biotechnology is a favorable alternative for their elimination. Recent technological developments demonstrated that the incorporation of an anaerobic process enhances the removal of recalcitrant micropollutants, such as pharmaceuticals (Ghattas et al. 2017). This technique might be introduced as a sustainable source of energy that generates a digestate with the potential to enhance plant growth and soil health (Mortezaei et al. 2018). Methane can be generated as renewable bioenergy under anaerobic conditions. The following chapter tries to summarize the recent works regarding the pharmaceutical wastewater treatment by anaerobic processes.

3.2

Aerobic and Anaerobic Processes

The organic content of any wastewater, including pharmaceuticals, could be reduced by controlling the activity of the organisms in the system. Generally, biological treatment processes might be classified into two main groups: (I) aerobic and (II) anaerobic. In aerobic processes, degradation of organic matters occur in the presence of oxygen. However, in the absence of oxygen, anaerobic degradation pathway is occurred. The aerobic process is a promising alternative for wastewater having low to medium organic content such as municipal wastewater or refinery wastewater; otherwise, once chemical oxygen demand becomes greater than 1000 mg/L, anaerobic process is more preferred. Additionally, less sludge and biomass are remained in anaerobic activities than aerobic. In Table 3.2, major differences of these two processes are listed. Organic matter þ O2 þ aerobes ! carbon dioxide þ water þ NH3 þ new cells Organic matter ! VFAs þ carbon dioxide þ CH4 þ energy þ residuals Table 3.2 Major differences between aerobic and anaerobic treatment processes Parameter Process By-products Applicability Reaction kinetic Sludge yield coefficient (kg VSS/kg COD) Post treatment Foot print Capital cost

Aerobic Treatment In the presence of oxygen Carbon dioxide, water, and excess biomass Low to medium organic content (< 1000 ppm) Decay rate kd ¼ 0.06 day1 0.35–0.45 (relatively high)

Anaerobic Treatment In the absence of oxygen Carbon dioxide, methane, and excess biomass Medium to high organic content (> 1000 ppm) Decay rate kd ¼ 0.03 day  1 0.05–0.15 (relatively low)

Direct discharge, followed by filtration/disinfection 1.0 to 2.4 kgCO2/kg COD removed 12–40 US$/inhab.

Generally done by aerobic methods 0.5 to 1.0 KgCO2/kg COD removed 40–65 US$/inhab.

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Anaerobic Metabolism

Anaerobic processes are of the most favored methods for organic matter degradation, having a much more complicated mechanism than aerobic techniques with more diverse microbial communities. Figure 3.2 shows the simplified anaerobic degradation of organic matters. Temperature is an important parameter in any biological system like anaerobic processes. Generally, the activity of microorganisms in a biological process decreases when the temperature decreases, which leads to a reduction in organic matter degradation. Anaerobic digester is typically run either at mesophilic (30–38  C) or thermophilic (52–55  C) conditions. In terms of pharmaceutical removal, it was reported that under thermophilic digestion conditions, a great number of antibiotic resistance genes were rapidly removed in comparison to mesophilic systems (Burch et al. 2016; Diehl and LaPara 2010). Of course, some believed that there was no difference in antibiotic resistance gene reduction based on the temperature (Zhang et al. 2015). The anaerobic process may be divided into four interdependent biochemical reactions: (1) Hydrolysis Hydrolysis is the initial step of the process that attempts to break down complex polymers (or large organic molecules) such as carbohydrates, proteins and fats into sugars, amino acids, and long-chain fatty acids, respectively, utilizing facultative and/or obligate anaerobic hydrolytic bacteria (Lei et al. 2018). The hydrolysis rate

Fig. 3.2 The main pathway in anaerobic processes for pollutants degradation which contains successive steps and ends in biogas production under accurate operation of the biological system

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depends on the temperature, pH, hydrolyzing mass concentration, and size/type of particulate organic compounds. The following equation shows an example of a hydrolysis reaction: C6 H10 O4 þ 2H2 O ! C6 H12 O6 þ 2H2 (2) Acidogenesis Once the simple molecules are formed as the final product of hydrolysis, the second process is started. In this stage, acidogenic bacteria ferment the simple soluble products (soluble sugars, amino acids, and fatty acids) into short-chain organic acids (formic, acetic, propionic, butyric, and pentanoic), methanol, ethanol, CO2, and H2. This process is called acidogenesis. As a matter of fact, the varieties of final products greatly depend on the operations of the system, especially hydrogen partial pressure. Acetate, carbon dioxide, and hydrogen are suitably formed at low H2 partial pressure; however, by increasing the hydrogen pressure, the conditions favored in terms of higher organic acid. The hydrogen concentration of a digester is an indicator of its health. It must be intended that excessive generations and accumulation of volatile fatty acids in the digester lead to a reduction in pH value, which is not favorable in terms of methanogens activity. Phyla Bacteroidetes, Firmicutes, Chloroflexi, and Proteobacteria are high active bacteria in this step. The following represents three typical acidogenesis reactions: C6 H12 O6 ! 2CH3 CH2 OH þ 2CO2 C6 H12 O6 þ 2H2 ! 2CH3 CH2 COOH þ 2H2 O C6 H12 O6 ! 3CH3 COOH (3) Acetogenesis As its name implies, this step involves the formation of acetate along with H2 and CO2. This process is accomplished by the acidogenesis products such as propionic acid, butyric acid, and alcohols. At low hydrogen pressure, propionate is likely to transform into acetate. Some of the main reactions that occur in the third stage of anaerobic fermentation are shown as: CH3 CH2 COO þ 3H2 O ! CH3 COOH þ HCO 3 þ 3H2 C6 H12 O6 þ 2H2 O ! 2CH3 COOH þ 2CO2 þ 4H2 CH3 CH2 OH þ 2H2 O ! CH3 COOH þ Hþ

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(4) Methanogenesis Methanogenesis is the last and final stage of anaerobic processes that convert acetic acid and hydrogen into methane gas and carbon dioxide. The generation of these gases is the sign of wastewater or sludge stabilization. There are two ways for methane production: (I) 70% of methane is produced from acetate by aceticlastic methanogens such as Methanosarcina and Methanosaeta (Ketheesan and Stuckey 2015), and (II) in the second way, conversion of carbon dioxide and hydrogen leads to methane generation through hydrogenotrophic methanogens (Xie et al. 2014). CO2 þ 4H2 ! CH4 þ 2H2 O 2C2 H5 OH þ CO2 ! CH4 þ 2CH3 COOH CH3 COOH ! CH4 þ CO2

3.4 3.4.1

The Common Anaerobic Biotechnologies for Pharmaceuticals Removal Anaerobic Digesters (ADs)

Anaerobic digesters are mostly an inevitable compartment of municipal wastewater treatment plants and are utilized to destabilize and treat the excessive sludge produced in the system. Since pharmaceutical compounds are detected in the influent of the wastewaters, it is essential to figure out their influence in the reactor. Roxarsone and sulfadiazine are veterinary antimicrobials for growth promotion and disease prevention (Fei et al. 2018). Their effects on biogas production during anaerobic degradation were investigated (Tang et al. 2019). At 120 mg/L of roxarsone, methanogenesis was inhibited by 99.9%. More inhabitation was observed once sulfadiazine and roxarsone were combined with each other. Similarly, in another study, once the concentration of oxytetracycline reached 80 mg/L, biogas generation decreased to 77.79% (Xin et al. 2014). However, at a low concentration of antibiotics (1 mg/L), biogas production was not affected (Feng et al. 2017). Since now, limited studies have been reported about the combined impact of drugs in anaerobic digesters. Yin et al. (2019) intended to figure out the effect of florfenicol, tylosin, and tilmicosin during anaerobic digestion. Based on the obtained results, the removal rates of florfenicol and tylosin under the combined antibiotic condition were no less than those under the individual antibiotic condition. Surprisingly, methane production was enhanced at the combined system. It seems that under these circumstances, the required carbon sources of biogas formation were increased, leading to a higher CH4 accumulation. More importantly, the initial concentration of these compounds must not be underestimated. For instance, the biogas production reduced from 10% to 38% when the tylosin concentration increased from 130 mg/L to 913 mg/L (Mitchell et al. 2013). For florfenicol, the

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biogas production reduced by 75% at the concentration of 210 mg/L (Mitchell et al. 2013). The presence of tylosin and chlortetracycline could adversely affect the digestion of swine manure (Stone et al. 2009). For instance, the generation of carbon dioxide and methane was inhibited by 28.4% and 27.8%, respectively, due to the presence of chlortetracycline. Logically, the type of drug may be a necessary item during experiments. Lins et al. (2015) studied the effects of several antibiotics and 2-bromoethanesulfonate during thermophilic anaerobic digestion. They found that methane production was inhibited by 89% by neomycin at the concentration of 100 μg/mL. Furthermore, kanamycins (100 μg/mL), rifampicin (50 μg/mL), and gentamicin (100 μg/mL) were effective for methane inhibition (27%, 72%, and 81%, respectively). Similar observations were reported by (Amin et al. 2006; Ji et al. 2013). It is fair to suggest that except the initial pharmaceuticals concentration, the type of the drug is of matter.

3.4.2

Anaerobic Membrane Bioreactors

Membrane bioreactor processes are of those treatment methods that integrate membrane with biological wastewater treatment processes. According to the operational situations, membrane bioreactor processes are divided into two groups: (i) aerobic membrane bioreactor and (ii) anaerobic membrane bioreactor. Intending the fact that no oxygen is required during anaerobic digestion, anaerobic membrane bioreactor has a greatly lower-energy input and sludge production in contrast to aerobic membrane bioreactor (Lee et al. 2017; Liao et al. 2006). Table 3.3 numerically Table 3.3 Main differences between AnMBR and AeMBR for wastewater treatment Features Energy consumption (kWh/m3) Biomass concentration (g/L) Organic loading rate (kg COD/ L) Organic removal efficiency (%) Hydraulic retention time (hours) Water flux (L/m2h) Sludge retention time (day)  Operational temperature ( C) Effluent quality Footprint Bioenergy recovery Startup time Temperature sensitivity Alkalinity requirement Nutrient requirement

Anaerobic membrane bioreactor Low (0.03–5.7) 10–40 High (0.17–35.5)

Aerobic membrane bioreactor High (~ 2) 5–20 High to moderate (0.25–0.8)

High (>90) >8 5–12 >100 20–50 High Low Yes < 2 week Low to moderate High to moderate Low

High (>95) 4–8 20–30 5–20 20–30 Excellent Low No < 1 week Low Low High

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Fig. 3.3 Comparison of AnMBR and AeMBR in terms of pharmaceutical compounds (AnMBR anaerobic membrane bioreactor, AeMBR aerobic membrane bioreactor)

compares the main differences between aerobic membrane bioreactor and anaerobic membrane bioreactor (Lin et al. 2013; Song et al. 2018). To elucidate the efficient of the aerobic/anaerobic membrane bioreactor, Fig. 3.3 provides precious results concerning removal efficiencies of the compounds. Generally speaking, anaerobic membrane bioreactor revealed high removal efficiency for most of the compounds; however, for some cases aerobic membrane was more suitable. The combination of anaerobic digestion with membrane filtration is a promising anaerobic technology that initially was utilized for whey processing wastewater treatment by Dorr-Olive in the 1980s in a commercial viewpoint. Since then, this technique has been employed for the treatment of various effluents (low, medium, and high strength wastewater), as shown in Fig. 3.4. The figure indicates that anaerobic membrane bioreactor typically owns high COD removal efficiency. High organic removal, biogas generation, low biosolid production, energy-saving, high permeate quality, and methane conversion rate are the most significant pros associated with anaerobic membrane bioreactor (Maaz et al. 2019). Of course, membrane fouling, cleaning, and controlling are considered as the major issues of the anaerobic membrane bioreactor that may limit the application of anaerobic membrane processes (Robles et al. 2018). Anaerobic membrane bioreactor can be operated under either thermophilic (50–60  C) or mesophilic (30–40  C) conditions (Song et al. 2018). Smith et al. (2015) operated anaerobic membrane bioreactor at psychrophilic temperatures ranging from 15 to 3  C (Smith et al. 2015). In terms of drug removal, various studies have been conducted. Among them, findings of C. Wijekoon et al. (2015) and Monsalvo et al. (2014) are very momentous. The researchers studied the removal efficiency of pharmaceutical compounds

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Fig. 3.4 Removal efficiencies of some compounds by anaerobic membrane bioreactor in two different studies conducted by C. Wijekoon et al. (2015) and Monsalvo et al. (2014)

and personal care products by anaerobic membrane bioreactors. For both studies, it is very evident that removal efficiencies vary from low (less than 10%) to high (nearly 100%) values, which is because of the type and intrinsic features (like biodegradability) of the compound. Virtually, in a similar removal technique, C. Wijekoon (2015) exhibited higher efficiency than Monsalvo et al. (2014). Intending improper temperature and inadequate hydraulic retention time are the chief two main reasons for such lower removal efficiency. Hydraulic retention time is another important factor affecting the performance of any biological systems. It was observed that the chemical oxygen demand removal efficiency was improved from 81.01% to 97.04% once the hydraulic retention time increased from 12 to 18 h (for treating tetrahydrofuran pharmaceutical wastewater) (Hu et al. 2018). Hence, it can be deduced that the in-situ optimization of experiments is an eminent factor affecting the final removal efficiency of anaerobic membrane bioreactor. Do and Stuckey (2019) investigated the removal of ciprofloxacin over a period of 120 days in an anaerobic membrane bioreactor. They found that at a ciprofloxacin concentration of 0.5–1.5 mg/L, the removal efficiency varies between 50% and 76%; however, at higher pollutant concentrations (4.7 mg/L), the efficiency of the process significantly decreased to below 20%. In another study, a pilot-scale anaerobic membrane bioreactor was operated for 435 days for m-Cresol, isopropanol, and N, N-dimethylformamide removal from a pharmaceutical solvent wastewater (Chen et al. 2018). The authors reported a total removal efficiency of >96% for all three pollutants at a psychrophilic condition (15
3  C). In another anaerobic membrane bioreactor, amoxicillin, ceftriaxone, cefoperazone, and ampicillin removal varied from 47.5% to 80.5%, 30.1% to 52.9%, 51.2% to 86.9%, and 21.3% to 42.9%, respectively, and organic loading rate/antibiotic loading rate was intended as the

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highest influential parameters affecting the performance of the bioreactor (Huang et al. 2018). Dutta et al. (2014) used a two-stage anaerobic fluidized membrane bioreactor for municipal wastewater containing pharmaceuticals. It was reported that this system could produce an effluent with near to zero values of biological oxygen demand, total suspended solids, and volatile suspended solids. Among the studied pharmaceuticals, 20 of them were removed in the range of 86–100%. Diclofenac was the only compound that its removal was lower than 80%. Other previous studies have also believed that diclofenac is persistent through anaerobic digestion processes (Carballa et al. 2007). No degradation of diclofenac was observed in the membrane bioreactor system (Kim et al. 2007). Sulfamethoxazole is a commonly consumed antibiotic with the high potential for the spread of the antibiotic resistance genes in the environment, which poses a serious health effect on human health (Beheshti et al. 2019). Biotransformation and removal pathway of sulfamethoxazole were carried out via a lab-scale mesophilic anaerobic membrane bioreactor under a hydraulic retention time of 1 day during a 170-day continuous operation (Wei et al. 2019). During sulfamethoxazole degradation, seven transformation products followed pseudo-first-order reaction kinetics were detected in which initial pollutant concentration was of high concern. Fortunately, the elimination of 97% was recorded for both chemical oxygen demand and sulfamethoxazole at high sulfamethoxazole concentration (10–1000 μg/mL). Normally, high sulfite concentration in the influent and its accumulation in the system could inhibit the activity of anaerobic bacteria, especially in methane production (Lens et al. 1998). To overcome the mentioned issue, Kaya et al. (2017) proposed a hybrid ozonation-anaerobic membrane bioreactor system for a chemical synthesisbased pharmaceutical wastewater treatment with SO24 and SO23 concentration of 4250–5600 mg/L and 3000–5250 mg/L, respectively. The bioreactor was successfully operated up to COD 7500 mg/L with the removal efficiency of 85-90 %. Also, pre-ozonation was effective to obtain high Etodolac removal (up to 99%). Interestingly, some researchers coupled anaerobic membrane bioreactor with adsorbents to improve the removal efficiency of the pharmaceuticals. Xiao et al. (2017) reported that the removal efficiency of carbamazepine, sulfamethoxazole, and diclofenac increased from 0.3
19.0%, 67.8
13.9%, and 15.0
7.2% to 92.4
5.3%, 95.5
4.6%, and 82.6
11.1%, respectively, in an anaerobic membrane bioreactor system after addition of powdered activated carbon (Xiao et al. 2017). It seems that adsorbent could enhance the final efficiency of the process. Besides, the addition of powdered activated carbon could reduce fouling in a membrane bioreactor and consequently improve their lifetime (Torretta et al. 2013). Also, for the removal of other organic micropollutants, anaerobic membrane bioreactor with ultrafiltration/nanofiltration/powdered activated carbon was reported with semi-similar results (Wei et al. 2015). Membrane fouling of anaerobic membrane bioreactor treating pharmaceutical wastewater was studied by (Kaya et al. 2019). Considering the above investigations, it is very much clear that anaerobic membrane bioreactor provides variable removal efficiencies for each pharmaceutical. As a matter of fact, the type of drug is very important for the determination of the final

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efficiency of the process. In addition, it was observed that in situ operational parameters significantly affected the overall process. Anaerobic membrane bioreactor coupling with other treatment methods might be a promising alternative for pharmaceutical compound removal. Apart from the importance of drug removal from pharmaceutical effluents, enough attention must be given to other organic/inorganic compounds/ions such as nitrogen and phosphor (Mohammadi et al. 2019). Sequencing anoxic/anaerobic membrane bioreactor has been proposed for simultaneous removal of N and P from wastewaters. The lab scale of this system was used for pharmaceutical treatment. The results showed the removal efficiency of 98%, 90%, 98.33%, 99.5%, 93%, and 83% for biological oxygen demand, chemical oxygen demand, pathogen, turbidity, nitrogen, and phosphorous, respectively (Al-Hashimia et al. 2013). Internal recycling time between the reactors profoundly affected the efficiency of N and P removal. Other investigations were also carried out by anoxic/anaerobic membrane bioreactor (Ahn et al. 2003; Cho et al. 2005).

3.4.3

Up-Flow Anaerobic Sludge Blanket

Up-flow anaerobic sludge blanket reactors are by far the most robust high-rate anaerobic reactors for domestic and industrial wastewater treatment in which wastewater enters from the bottom of the reactor and flows upward and is treated after being contacted by a suspended sludge blanket (Fig. 3.5). Low cost, flexibility, and high biomass concentration are of the most reasons that engendered the installation of more than 1000 up-flow anaerobic sludge blanket reactors (especially in tropical areas) around the world (Chen et al. 2014; Tiwari et al. 2006). Figure 3.6 shows the main challenges and opportunities regarding this treatment technique. Up-flow anaerobic sludge blanket has been utilized for different wastewaters, and relatively high chemical oxygen demand removal efficiency has been reported in the literature (Fig. 3.7). It is fair to suggest that up-flow anaerobic sludge blanket is a competitive industrial wastewater decontamination method. Hou et al. (2019) used three successive processes of up-flow anaerobic sludge blanket, anoxic/oxic tank, and advanced oxidation technologies for the removal of 18 antibiotics and 10 antibiotic resistance genes from pharmaceutical wastewater over a 6-month operation. The results demonstrated that up-flow anaerobic sludge blanket provided the greatest contribution (85.8
16.1%) for the removal of 18 antibiotics. In the fermentation and chemical synthesis-based pharmaceutical wastewater, 6aminopenicillanic acid and amoxicillin are of the prominent pollutant that Chen et al. (2011) tried to remove them via up-flow anaerobic sludge blanket by varying the chemical oxygen demand loading rate (2.57–21.02 Kg/m3.day) and pH values (5.57–8.26). The operation conditions and wastewater characters are influent chemical oxygen demand (7458 mg/L), hydraulic retention time (23.2 h), influent pH (7.4), influent amoxicillin (92 mg/L), and influent 6-aminopenicillanic (192 mg/L).

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Fig. 3.5 The schematic of an up-flow anaerobic sludge blanket reactor in which the influent enters from the bottom of the reactor then goes up

The results demonstrated that the up-flow anaerobic sludge blanket average percentage reductions for COD, 6-aminopenicillanic acid, and amoxicillin were 52.2%, 26.3%, and 21.6%, respectively. N,N-dimethylformamide is a toxic, carcinogenic, thermally stable, and weak degradable solvent that has been extensively used in industrial activities, such as the pharmaceutical industry (Nisha et al. 2015; Vidhya and Thatheyus 2013). Accordingly, its removal was studied by up-flow anaerobic sludge blanket and anaerobic membrane bioreactor (Li et al. 2019). At a low organic low rate, both up-flow anaerobic sludge blanket and anaerobic membrane bioreactor showed high CH4 production and pollutant removal; however, at the extremely high organic rate (greater than 6), the performance of both reactor decreased sharply. It was observed that anaerobic membrane bioreactor had less tolerant than up-flow anaerobic sludge blanket since its membrane was damaged by excessive sludge production (Li et al. 2019).

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Fig. 3.6 The most important uniqueness and challenges of up-flow anaerobic sludge blanket treatment method

Aromatic sulfonates include a broad group of chemicals applied in the dye and pharmaceutical industries that are water-soluble and could generate high strength wastewater (Tan et al. 2005). Treatment of p-acetamidobenzene sulfonyl chloride via up-flow anaerobic sludge blanket was investigated by Li et al. (2015). This study demonstrated that thorough degradation of p-acetamidobenzene sulfonyl chloride (50 mg/L) was feasible by bacterial group phyla Proteobacteria at organic load rate

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Fig. 3.7 The removal efficiency of different wastewaters by up-flow anaerobic sludge blanket reactors

and HRT of 7 g COD/L/day and 0.5 h, respectively. Also, phylum Clostridia was responsible for desulfonation. In other experiments regarding dilute pharmaceutical wastewater, Firmicutes, Bacteroidetes, Thermoplasmata, and Methanobacteria were dominant microbial community groups at high organic loading rates (Chen et al. 2014). Anaerobic packed bed reactors were first proposed as a treatment process by Young and McCarty and are similar to a trickling filter, having advantages like a reasonable reactor volume, low hydraulic residence times, acceptable growth of anaerobic microorganisms, and biofilm production (Bakhshi et al. 2011). In terms of pharmaceutical removal by anaerobic packed bed reactor, Comoglu et al. (2016) conducted an investigation and reported that under hydraulic retention time and organic loading rate of 2.5–4 day and 0.6–2.2 g chemical oxygen demand/d, the chemical oxygen demand removal efficiency was 93–97%.

3.4.4

Anaerobic Sequencing Batch Reactors

Anaerobic sequencing batch reactors, firstly developed and patented at Iowa State University, include five successive cycles (fill, react, settle, decant, idle) that the biochemical reactions occur in a single reactor under the absence of light and oxygen. Reactor geometrics, feeding strategy, the substrate to biomass concentration, mixing, temperature, and hydraulic retention time are the most important

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Fig. 3.8 The removal efficiency of different wastewaters by anaerobic sequencing batch reactors

parameters influencing the final efficiency for any pollutant degradation (Kannan and Singaram 2012). Recently, the effluents of palm oil mill (Maaroff et al. 2019), cassava processing (Mari et al. 2019), flushed dairy manure (Zeb et al. 2019), biodiesel production (Pereira et al. 2019), and acid mine drainage (Giordani et al. 2019) have been studied by anaerobic sequencing batch reactors with the aim of wastewater treatment or bioenergy generation (Fig. 3.8). The results in the figure demonstrate that anaerobic sequencing batch reactors could significantly reduce the COD level of the industrial wastewater. Landfill leachate, for instance, the removal efficiency of COD was greater than 80%. The same results also was reported for Hence, it is considered that even for pharmaceutical wastewater it would be a promising alternative. In real environment media, pharmaceuticals are not lonely existed, and in fact they are combined with other pollutants or other groups of drugs. The effects of mixtures will be different from the single compounds, which can be antagonistic or synergistic, and often higher combination effect is observed in the literature (Cleuvers 2004). Based on such behavior, Aydin et al. (2014) tried to figure out the simultaneous effect of erythromycin and sulfamethoxazole in anaerobic sequencing batch reactors. Firstly, biogas/methane generation and substrate/chemical oxygen demand consumption decreased by the enhancement in pollutant concentration. Furthermore, the inhibitory effects of the pharmaceutical compounds affect Gramnegative bacteria much more than Gram-positive bacteria. The combined effect of erythromycin, tetracycline, and sulfamethoxazole was also studied (Aydin et al. 2015). In another investigation regarding the saline pharmaceutical wastewater treatment, Methanobacteria, Methanomicrobia, and Thermoplasmata were introduced as the predominant archaeal groups involved in the process (Shi et al. 2015).

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Like any other biological system, the initial concentration of pharmaceuticals is greatly important. In the case of anaerobic sequencing batch reactors, it was seen that the sulfamethoxazole wastewater concentration threshold was 40 mg/L, indicating that up to this level, the anaerobic sequencing batch reactors were a suitable method (Cetecioglu et al. 2015). In other words, at higher 40 mg/L, the microbial activity was impaired, and so biogas/methane ratio was inhibited. In terms of sequencing reactors, a pilot reactor anaerobic/aerobic sequencing batch biofilter packed with a porous volcanic stone was used to treat pharmaceutical hospital (Buitron et al. 2003). They found that up to the organic load of 6 kg chemical oxygen demand/m3.d the removal efficiency was satisfactory and super to 6 the quality of the treated wastewater started to decline. This finding is in good agreement with (Seif et al. 1992). From the total organic matter removal, 78.5% and 20.3% were removed in the anaerobic stage and aerobic stage.

3.5

Conclusion

This chapter reviews the recent works conducted on hospital and pharmaceutical wastewater treatment employing anaerobic systems. The release and danger of pharmaceuticals have forced researchers to implement more efficient methods to mitigate their effects on the environment. Briefly, it was observed that high organic/ inorganic contents of these wastewaters, along with their variations in the influent, had made them a hard and strength wastewater in terms of treatment. Anaerobic biotechnology was seen as a promising and appropriate alternative. Anaerobic digesters proved that the type of pharmaceuticals greatly influenced the dominant microbial community in the system. High treatment efficiency was observed by anaerobic membrane bioreactors under the conditions that the optimum operational values are accurately considered in the reactor. Similar results were observed in up-flow anaerobic sludge blanket and anaerobic sequencing batch reactors. As a matter of fact, to reach high removal efficiency in any anaerobic techniques, the types of drugs and in-situ operational conditions are of great matter.

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Chapter 4

Bacterial Metabolites for Removal of Toxic Dyes and Heavy Metals Sriparna Datta, Dipanjan Sengupta, and Ishika Saha

Contents 4.1 4.2 4.3 4.4 4.5

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Involvement of Bacteria in Removal of Toxic Dyes and Heavy Metals . . . . . . . . . . . . . . . Biosorption: A Practical Measure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Involvement of Bacterial Metabolites in Biosorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Factors Affecting Biosorption of Dyes by EPS or Bioflocculants . . . . . . . . . . . . . . . . . . . . . . 4.5.1 Effect of Carbon Source and Initial Dye Concentration . . . . . . . . . . . . . . . . . . . . . . . . 4.5.2 Effect of pH, Temperature, and Contact Time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5.3 Effect of Bioflocculant or Extracellular Polymeric Substance Concentration . . 4.5.4 Effect of Metal Stress . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5.5 Effect of Salts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.6 Regeneration of Bacterial Cells: Desorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7 Mechanisms and Factors Involved in Bacterial Binding of Heavy Metals: At a Glance 4.8 Involvement of Bacterial Genetic Factors and Genetic Manipulation for Enhancement of Biosorption of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.9 Live Bacterial Cell, Dead Biomass, and Bacterial Metabolite in Bioremediation: A Comparison . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.10 Future Scope of Bacterial Metabolite in Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.11 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Clean, consumable water and usable soil are basic necessities for sustenance of mankind. But unfortunately owing to unprecedented urbanization and industrialization, global ecosystem and its resources are under constant threat and are gradually losing sustainability. Natural water bodies and soil are progressively becoming unusable due to relentless emission of potentially toxic substances. Out of several factors augmenting such crisis, toxic dyes and heavy metals, drained as by-products of textile, tannery, chemical, petroleum, and other related industries, play alarming roles. So far several technologies have been implemented for

S. Datta (*) · D. Sengupta · I. Saha Department of Chemical Technology, Rajabazar Science College, University of Calcutta, Kolkata, India © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 85 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_4

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mitigation involving physical or chemical means of detoxification. However, due to the high cost-efficiency ratio, feasibility becomes a practical question. In such circumstances, various biological techniques involving nonpathogenic bacteria and their metabolites are gaining importance in the domain of environmental research. This chapter highlights the details of some potential bacterial strains which are involved in such remediation process. The mechanisms involved in production of bacterial metabolites for biosorption have been focused upon. Moreover, the physicochemical and genetic factors involved in the production of such potent bacterial metabolites have been studied as well. The importance of genetic manipulation in augmenting the metabolite production has also been highlighted. The contrasting efficacy of viable bacterial cells, bacterial metabolites, and dead biomass in biosorption process of heavy metals and dyes has been explored. The reusability of such biosorbents has also been considered which makes such bioremediation measures quite promising in futuristic applications. Keywords Toxic dyes · Heavy metals · Biosorption · Bioremediation · Bacterial metabolites · Environmental · Biodegradable · Exopolysaccharide · Genetic factors · Desorption

Abbreviations ABC CBA CDF EPS FSSAI LPS OPX PAH PCP PCR PEI PhoK PhoN

4.1

ATP Binding Cassette Chemiosmotic antiporter Cation diffusion facilitators exopolysaccharides Food Safety and Standards Authority of India Lipopolysaccharide Outer membrane polysaccharide export protein Polycyclic aromatic hydrocarbon Polysaccharide co-polymerase Polymerase chain reaction Polyethylene amine Periplasmic acid phosphatase Extracellular alkaline phosphatase

Introduction

Through years global population and rapid urbanization have relentlessly increased the rate of pollution. The effects of pollution are detrimental to both human health and ecology. One of the main pollutants are the heavy metals and dyes which are used considerably in industries like chemical, petroleum and oil, textile, wool, and leather. The interaction of microorganisms and/or their metabolites with potentially

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toxic compounds is indeed an interesting as well as essential arena of study. Considering the current scenario of the ecosystem and the ecological balance, we are facing a serious threat against a sustainable environment as far as the future generations are concerned. Such ecological imbalance is often being expressed by the accumulation of toxic substances in the environment that demands a fair disposal or degradation. So far we have adapted several physical as well as chemical means that proved to be partially effective for such purposes. However, the cost that is incurred in the process is sometimes not permissible. In such a scenario, the application of microorganisms or their metabolites for proper treatment of heavy metals and other toxic substances like dyes or non-biodegradable plastics is constantly gaining attention. According to the Central Water Commission, 42 rivers in India have already surpassed the permissible toxicity limit of at least 2 potentially toxic heavy metals (Pandey et al. 2018). This is quite alarming as majority of the population use water directly from these rivers for household purposes. Furthermore, according to Consumer Reports, presences of heavy metals like arsenic and lead have been detected in 45 packaged fruit juices (CBS News 2019). Recently, this situation has extended so much so that even baby foods are not devoid of toxic metals (Welch 2019). Similar is the picture with toxic dyes that are often being used as a coloring agent in edible products to enhance the aesthetic appeal. Until the mid-nineteenth century, natural dyes from plants, animals, and minerals were used to color food. However, things have changed after the discovery of artificial dyes, and the reckless use of color to attract customers has not bypassed food industries as well. Such dyes have severe toxic effects. From an FSSAI (Food Safety and Standards Authority of India) report, it is quite alarming to note that even the Indian sweets are not devoid of the banned toxic dyes like rhodamine B, orange II, metanil yellow, malachite green, and auramine (Sachan 2013). Even toxic dyes used in textile and other industries often release heavy metals into the water bodies and contaminate the soils and crops. For instance, in Sanganer, Rajasthan, a significantly high level of chloride and copper between 7 and 53 mg/g in vegetables were detected which were much higher than the limits set by Bureau of Indian Standards, New Delhi (Down to Earth 2015). Dyes are mostly soluble organic substances which can impart color to textiles, paper, leather, and other substances such that they are not readily altered away by washing, heat, or light. Pigments on the other hand are inorganic and finely ground powder in nature. Unlike pigments, dyes can chemically bond to the applied material. The basis of the color of dyes is mainly due to the following: They absorb light in the visible region (400–700 nm) and possess at least one chromophore or color-bearing group, a conjugated system (bearing alternate double and single bonds), and structure bearing resonating electrons which stabilize organic compounds. Some dyes function together with auxochromes (color helpers) examples of which are carboxylic acid, sulfonic acid, amino and hydroxyl groups along with chromophores. The auxochromes are not directly responsible for imparting color but are known to shift the wavelength of the chromophore group and also enhances dye solubility (IARC Monographs on the Evaluation of Carcinogenic Risks to Humans).

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The chromophores in anionic and non-ionic dyes mostly consist of azo groups or anthraquinone types (Moody and Needles 2004; Srinivasan and Viraraghavan 2010). Metal complex dyes are found among reactive, direct, and acidic dyes of anionic class depending upon mode of applications. These dyes form a mono azo structure which is formed by complexes of selective metals (nickel, chromium, and copper) with dyes possessing additional groups like hydroxyl, carboxyl, and amino groups. Generally, the dyes are stable organic compounds and thus can persist in the environment for years. For example, anthraquinone-based dyes are more resistant to degradation due to their fused aromatic structures (Srinivasan and Viraraghavan 2010). On top of that, an increase in urbanization has amplified the human exposure to such pollutants to a great extent. The current physicochemical strategies for removal of heavy metal and dyes from the industrial effluents are mainly done by numerous conventional methods like chemical precipitation, reverse osmosis, ultrafiltration, ion exchange, electrowinning, and carbon adsorption. But these processes are too expensive and energy requiring and do not guarantee the complete removal of the pollutants (Pereira and Alves 2012; Aksu and Balibek 2010). Researchers are seeking suitable alternative cost-effective approaches to combat the harmful effects of these recalcitrant pollutants in a sustainable manner. The use of living organisms and their metabolites for degradation and subsequent removal of these pollutants from the ecosystem is popularly termed as bioremediation. “Bioremediation” is the subject of immense interest by the researchers as it is effective, cheap, and environmentally friendly. In recent years, many studies have been reported on bioremediation by various genera of microorganisms like bacteria, fungi, and algae (Ghosh et al. 2015). On the other hand, bacteria-mediated remediation of polluted lands and water is by far a cheap, efficient, and environment friendly method. There are a number of strategies by which a bacterium can detoxify its surroundings for its survival: (1) biosorption and bioaccumulation, (2) bacterial secondary metabolite production, and (3) biodegradation by enzymes or bioreduction (Bhatia et al. 2017). Among these three bacteria-mediated processes, the chapter mainly focuses on the ability of a bacterium to remediate (i) directly depending upon the nature of its cell surface and (ii) via production of secondary metabolites.

4.2

Involvement of Bacteria in Removal of Toxic Dyes and Heavy Metals

Among the diverse bacterial consortia, many display beneficial properties like biodegradation and/or bioremediation which are of immense importance in the environmental facets. In such context, we cannot help but mention cyanobacteria, one of the first life forms of the Earth (circa 4 billion years ago) and its enormous role in the sustenance and restoration of an ecosystem. It has been postulated that it was largely due to this life form which produced oxygen as a by-product that our planet evolved into an oxidative atmosphere. Furthermore cyanobacteria help improve the

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soil health by its potency to fix atmospheric nitrogen (Johnson 2017). However, the uncontrolled and limitless usage of heavy metals which are often naturally occurring has aggravated the damage to such an extent that it is gradually turning into a Frankenstein against our own ecosystem. This often results in bioaccumulation in the food chain especially due to the non-biodegradability of such heavy metals (Vhahangwele and Khathutshelo 2018). Scientists are constantly in search of microorganisms and other life forms that can aid in restoration of the ecosystem. Such context bacterial species like Bacillus subtilis, B. cereus, Lactobacillus bulgaricus, Micrococcus luteus, Pseudomonas putida, and P. aeruginosa have shown appreciable biosorption of metals like lead, copper, and cadmium (Hossain and Anantharaman 2006; Murthy et al. 2012; Sedighi et al. 2012; Puyen et al. 2012; Pan et al. 2007; Uslu and Tanyol 2006: Pardo et al. 2003). The zinc-resistant bacterium Brevibacterium sp. has shown appreciable biosorption of zinc (Taniguchi et al. 2000). P. aeruginosa and Bacillus cereus also display the biosorption of zinc (Joo et al. 2010). Even relatively less popular bacterial species like Achromobacter sp. isolated from oil refineries have been reported to show zinc biosorption (Subudhi et al. 2014). It can be well speculated that microorganisms especially bacteria that thrive well in the industrial vicinity often possess such remarkable propensity to utilize the industrial wastes and in the process also degrade the toxic heavy metals. Such potential strains require further investigation and should be studied in detail. Halophilic bacteria often act as potent tool in the biosorption of heavy metals. For instance, Halomonas sp. impressively caters to the biosorption of lead and cadmium (Amoozegar et al. 2011). Sometimes it is also observed that bacterial species like Brevibacterium sp. have the potency to adsorb several heavy metals (Igiri et al. 2018). However the affinity of the bacterial cell to the type of metal ion varies accordingly. For instance, Brevibacterium sp. can bind lead, copper, and cadmium, but the order of affinity is Cu > Pb > Cd. Often it is also observed that the presence of a particular metal either accentuates or hinders the binding of another metal. In case of Brevibacterium sp., the presence of lead inhibited the removal of cadmium and copper (Vecchio et al. 1998). Basically the polarizable functional groups (like phosphate, amino, carboxyl, and hydroxyl groups) present in the multilayered bacterial cellular envelope (comprising capsule, cell wall, membrane) creates a field affinity for the cations to bind. On the other hand, the activity of bacterial oxidoreductases and other enzymes, the accumulation of metabolite, metal binding proteins, and other related factors often play as manipulative factors in the bioavailability of metal ions (Vecchio et al. 1998). From FTIR studies of Alcaligenes sp. biomass in biosorption of lead, it has been observed that the hydroxyl, carbonyl, and phosphate groups on bacterial cell surface are involved in the process (Jin et al. 2017). However there is a subtle difference between the “bioaccumulation” and “biosorption” in accordance to bacterial cell– metal ion interaction. Bioaccumulation, unlike biosorption, is a metabolism dependent mechanism and involves several enzymes and related proteins. So biosorption often involves non-viable cells unlike that in case of bioaccumulation.

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Often it is found that dead bacterial cells cater a better potency of biosorption of heavy metals in comparison to live cells. It was found that dead cells mass of Bacillus cereus proved to be a better biosorbent of cadmium ions than that of live cells (Huang et al. 2013). This might be due to the fact that large populations of the bacterial strains have a natural tendency to allow only a permissible limit of metal ions into the intracellular matrix for cellular metabolism. Often, this cellular mechanism is an energy-dependent process. Thus, amounts of metal ions exceeding a certain limit are often rejected by bacterial cells. Moreover individual bacterial strains have a certain toxicity limit to metal ions. On one hand, a specific bacterial species has an affinity for a specific metal ion, while on the other certain metal ions are readily rejected. Thus the selection of viable bacterial strain as biosorbent of a particular metal is an essential prerequisite. Other important factors that often play modulatory role in the biosorption of heavy metals are pH, temperature, salinity, media shaker speed and oxygen transfer, and various others. Therefore, optimization of both physical and chemical factors involved in maximization of the biosorption process by a bacterial species is a necessary prerequisite. The pH of medium for bacterial biosorption often affects the functional groups and the metal binding sites on the bacterial cell surface. Furthermore it alters the metal ion solubility in water. An important factor determining the biosorption activity of heavy metal by bacteria is pH. A particular bacterial strain often exhibits an effective biosorption of a particular metal ions at a definite pH, while the biosorption of some other metal by the same strain might happen at a different pH altogether. This can be well understood by the metal biosorbing tendency of Bacillus thuringiensis. Nickel and chromium biosorption by the bacterial strain takes place at an optimum pH of 7, while that in case of cadmium, copper and lead is 6 (Oves et al. 2013). Moreover, the cell wall architecture of the bacteria also plays the pivotal role in adsorption of dyes. The cell wall is located outside the cell membrane and it provides the necessary mechanical strength and rigidity to the cell. Generally, two classes of bacteria exist depending upon their cell surface morphology: Gram positive and Gram negative bacteria. The cell wall of a Gram positive bacterium mostly comprises (90%) of peptidoglycan and teichoic acids. The teichoic acids can also be covalently linked to the peptidoglycan via a lipid moiety. In case of Gram negative bacteria, there is presence of an additional outer layer or otherwise termed as outer membrane outside of the cell wall. It mainly comprises of lipopolysaccharide (LPS) and phospholipids (Bruslind 2017). It was further proved that the teichoic acid and peptidoglycan in Gram positive bacteria and the lipopolysaccharide and phospholipid in Gram negative bacteria play vital role in imparting a charged surface for binding of the dyes (Vijayaraghavan and Yun 2009). Charged functional groups like carboxylic, amino, hydroxyl, and phosphates present in the cell wall are mainly responsible for its binding capacity. The chemical structure, type, and number of chemical substituent of the dyes also influence the adsorption process (Wang et al. 2007; Zabochnicka-Świątek 2014). A large number of reports suggest that dyes can be utilized by the bacterial species as sole carbon sources; still an optimum dye concentration is necessary for its survival (Sarnaik and Kanekar 1999). Azo dyes in particular have shown maximum

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resistance to degradation by microorganisms (Sudha et al. 2014; Michaels and Lewis 1985). For example, Basic Violet 3 and Basic Green 4 are used as inhibitory agents in microbiological media. Still some bacterial species have been reported that utilize these dyes as sole carbon and nitrogen sources because of the enzymes are present intracellularly (Michaels and Lewis 1985). Triphenylmethane dyes contain methyl groups in their structure and hence can be readily be metabolized by the bacteria (Michaels and Lewis 1985). Another impressive bacterial trait is to co-metabolize in their natural environment where these co-substrates like glucose, yeast extract, and enzymes act as an electron donor. They induce the biodegradation process and reduce the overall decolorization time (Jadhav et al. 2008; Hu 1998; Banat et al. 1996). The inhibitory effect of dyes on live cells was further proved in case of Bacillus subtilis (IFO 3022) isolated from activated sludge. Triphenylmethane dyes have shown to inhibit bacterial growth in very low concentrations. The dyes methyl violet, acridine orange, and astrazon red with increasing concentrations have shown to decrease the RNA/DNA concentrations in the Bacillus sp. in the logarithmic phase. The mean growth rate of the cells and cell concentration decreased with increasing concentration of methyl violet in the logarithmic and stationary phase, respectively. Thus it can be said that these dyes more actively inhibit protein synthesis than cell division. Contrarily, biosorption normally involves the utilization of dead pre-treated bacterial biomass, and the whole process is dependent upon the surface of the cell wall; hence it is metabolism independent (Ogawa et al. 1988). A study on Pseudomonas luteola grown in presence of few azo dyes reported that cellular viability is necessary for the decolorization process (for activation of azodoreductases). But if the culture conditions and concentration of dyes cumulatively or alone fails to restore cellular viability, then biosorption becomes more prominent than decolorization due to lack of metabolic activity (Chen 2002). So it can be said that higher dye concentrations in the medium often induce biosorption process and later reach saturation at higher equilibrium condition. This mostly happens due to presence of lesser binding sites available on the cell surface for the dye molecules to get adsorbed when the dye concentration is increased (Vijayaraghavan and Yun 2007; Arunarani et al. 2012). Many studies have claimed that pre-treating the cells by chemical agents, oven heating, or autoclaving, genetic modification, etc. can enhance the biosorption capability of the cells. In many cases, it has been observed that dead bacterial cells have proved to be better adsorbents than the live cells. Heat-treated Aeromonas sp. and Pseudomonas sp. strain DY1 exhibited greater biosorption ability of the dyes than the live cells. It was suggested that the heat treatment increases the surface area of the cells and thus improves adsorption (Hu 1992). This is mainly due to the fact that permeability of the cell wall increases with increasing the temperature. Elucidation of various characterization studies speculated that after the heat treatment, the cell wall becomes thinner suggesting the enhancement of permeability. The possible mode of action could be penetration of dye macromolecules through the permeable cell wall structure and binding of the dyes at specialized functional groups located inside the bacterial cell wall. The intracellular proteins play the key role in adsorption of the dye. The various characterization studies before and after the incubation suggested

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that the amines mainly act as the functional group in binding of dye (Du et al. 2012). Powdered biomass of Corynebacterium glutamicum cross-linked with polyethylene amine (PEI) shows 2.8 times enhanced biosorption of Reactive Red dye via electrostatic interaction of the amine groups present in the cell wall. The PEI increases the presence of primary and secondary amine groups in the cell wall. Chemically treating the cells with acid, alkali, salts, ethanol, and acetone enhances sorption capacities (Mao et al. 2008). Citric acid-treated Corynebacterium glutamicum shows excellent sorption capacity of basic dyes. On heating citric acid dehydrates to form a reactive anhydride which when present in reaction mixture combines with hydroxyl groups of the cell wall to form citrate-biomass adduct. The extra binding sites provided by the modified biomass act as suitable binding sites for the Basic Blue 3 dye (Mao et al. 2009). Quite similar to this, another study revealed that Corynebacterium glutamicum biomass shows maximum uptake of Reactive Black 5 dye (195 mg/g) when treated with HNO3 with the initial dye concentration 500 mg/l. As mentioned before, the amine functional groups played a vital role in binding with the dyes (Vijayaraghavan and Yun 2007). The pH of the surrounding medium exerts direct influence on the electrochemistry of the dyes, the dissociation of the poly-electrolytes, and hence their structural confirmation in the solution (Somasundaran and Runkana 2005). At high pH, the positive charge at the dye solution interface decreases and the biomass adsorbent acquires negative charge. Hence, the biosorption of cationic or basic dyes increases when pH increases. In contrast, at low pH conditions, the positive charge at the dye solution increases, and the biomass adsorbent appears positively charged. Thus, the biosorption of anionic dyes increases at low pH condition (Salleh et al. 2011). Bacterial strains Aeromonas sp., Pseudomonas luteola, and E. coli adsorbed negatively charged reactive dyes. Decarboxylated Corynebacterium glutamicum adsorb Reactive Black and Planococcus sp. adsorbs Black TSB when the pH was maintained at acidic range. The carboxyl and amine groups attain a positive charge which facilitates the interaction between the positively charged bacterial cell and negatively charged dye anions thus enhancing biosorption process (Hu 1996; Vijayaraghavan et al. 2007; Choudhary et al. 2015). On the other hand, Bacillus catenulatus JB-022 shows maximum biosorption of cationic basic dye, Basic Blue 3 (BB3) at alkaline pH. The carboxyl groups present in the cell wall become negatively charged at pH higher than 5 thus enhancing electrostatic interaction between the dye and the biosorbent (Kim et al. 2015). Hence it can be said that along with the charged surface of the bacterial strain, the adsorption mechanism is also dependent on the nature of the dyes (anionic or cationic). In case of live bacterial cells, it is often seen that the removal of dyes is dependent upon the optimum growth temperature of the bacterial strain (35–45
C). Generally, the removal of color increases with increasing the temperature. However, after a certain temperature is reached, the removal activity decreases. This can be attributed to the loss of cell viability or heat denaturation of azo reductase enzyme (Pearce et al. 2003; Chang et al. 2001). The biosorption of Reactive Black 5 by pre-treated Corynebacterium glutamicum was observed at temperature range (25–40
C). High temperature facilitates enhanced biosorption, the optimum being at 35
C.

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This indicates the endothermic nature of the biosorption process (Vijayaraghavan and Yun 2007). In another investigation, the optimum temperature favoring the biosorption process occurs at low temperatures. Streptomyces fradiae shows maximum adsorption at 20
C thus inferring the nature of the process as exothermic (Velkova et al. 2017). Contact time is one of the important parameters in the biosorption process. The first 40 min showed rapid uptake of the dyes by the strain Streptomyces fradiae followed by the secondary slow phase until equilibrium was reached at 80 min and 70 min for methylene blue and Congo red, respectively. The presence of adequate binding site suggests rapid uptake of dyes initially. Gradually as the binding sites become saturated, the bioadsorbent losses its effectiveness (Velkova et al. 2017). In another instance, contact time and temperature cumulatively favor the biosorption process. Temperature range between 20 and 25
C is found to be optimum for maximum uptake of dyes by the bacterium Paenibacillus macerans suggesting the phenomenon to be characteristically similar to physical adsorption and hence is exothermic (Çolak et al. 2009). It is important to maintain the stability of the bioadsorbents in the whole process. Free cells are not resilient enough to undergo industrial operations mainly due to poor mechanical strength and rigidity. Immobilization of cells provides a proper frame work, mechanical strength, porosity, and rigidity making the process ideal for continuous mode of operation (packed and fluidized bed reactors) in industries. Some of the techniques for immobilization include adsorption on inert supports, entrapment in polymer matrix, and cross-linking of the cell to other compounds (Veglio and Beolchin 1997). Decarboxylated Corynebacterium glutamicum biomass immobilized on polysulfone matrix showed efficient removal of Reactive Black 5 dye (180.7 mg/gm dry solid beads) when the initial dye concentration in the solution is 500 mg/L. Column experiments established the reusability of the bio sorbent with over 74 mg/gm of dry bead even after progressive cycles. Although the usage of immobilized bio adsorbent comes with its benefits, there are some practical problems likely to be encountered. The biomass will be retained inside the polymeric matrix, and thus the mass transfer resistance will come in the forefront. This mass transfer resistance slows down the attainment of equilibrium. Even though the process of equilibrium slows down, a successful immobilization will usually have the functional group sites of the adsorbent activated for the uptake of dyes (Vijayaraghavan et al. 2007). In a similar study, immobilization of cells has resulted in better biosorption of reactive dye Procion blue 2G, with maximum dye uptake of 1.648 mg g1 in case of P. aeruginosa and 1.242 mg g1 for P. chrysosporium when the initial dye concentration was 100 mg L1 (Saravanan et al. 2013). Immobilization of cells enhances decolorization of dyes by imparting high stability to the enzymes present in the whole cells. Further, it protects the cell from the recalcitrant compounds that can be toxic to the living cells (Khan and Banerjee 2010). Sodium alginate immobilized bacterial strains P. putida and B. licheniformis maximally decolorized reactive dyes in comparison to poly acrylamide immobilized cells in both static and shaken conditions (Suganya and Revathi 2016).

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At high concentrations of biomass, removal of dyes from solution is influenced. However, much higher concentrations can lead to cellular aggregates that can disturb the whole process. Streptomyces fradiae biomass showed enhanced biosorption with the optimum being 2 gm dm3 resulting in removal of 91.20% and 73.10% in case of methylene blue and Congo red respectively (Velkova et al. 2017). It was concluded that the greater surface area and presence of adequate binding sites is the reason behind the phenomenon. Similar results were obtained with Paenibacillus macerans and Pseudomonas putida (Arunarani et al. 2012: Çolak et al. 2009).

4.3

Biosorption: A Practical Measure

The method of adsorption has been found to be most effective for removal of dyes from polluted environments. Activated carbon is widely used as an adsorbent in multiple arenas, but it is too expensive. Thus the exploration of low cost adsorbents is gaining momentum in research of which non biological sources include peat, bentonite, steel-plant slag, fly ash, china clay, and silica (Crini et al. 2019). Leaf mold, rice husk, groundnut husk, coconut husk and palm pressed fibers, coconut shells, coconut jute, coconut tree sawdust, cactus, olive stone cake, wool, and pine needles are among the biological sources that can be used as adsorbents. Other than these materials, in recent times suitable biological biomass (bacteria, fungi, and algae) has been reported as a potential substitute for effective removal of heavy metal ions and dyes (Joshi 2017). The term “biosorption” can be defined as the ability of the biological material to uptake the heavy metals and dyes from polluted environments through metabolically mediated or physicochemical pathways. Biosorption is advantageous compared to other conventional techniques because of the following: • • • • • •

Low cost High efficiency No use of chemical reagents No additional nutrient requirement Regeneration of the biomass Possible recovery of heavy metal and dyes (Kratchovil and Volesky 1998; Veligo et al. 1997; Gadd and White 1993)

Biosorption is passive in nature and does not essentially require living cells. Various studies have been conducted in both batch and continuous mode with the biomasses that can either be living, dead, pre-treated dead, or immobilized cells for the elimination of the metals and dyes (Joshi 2017). The implementation of technologies involving the use of microorganisms especially bacteria and/or their secondary metabolites is comparably cost-effective yet efficient in the alleviation of the toxic levels of heavy metals and dyes. The mechanisms involving such technologies are diverse and also differ from one species to another. Furthermore the structural and function attributes of the metabolites

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produced by the microorganism in removal of toxic metal ions from a contaminated source is also diverse. In comparison to common adsorbents like ion exchange resins and activated carbon which are high in cost and also lower in efficiency, biomassbased adsorbent or biosorbents are most attractive alternative due to their lower cost of production and high efficiency. Looking back in 1902 with the uptake of copper by the fungal spore Tilletia tritici and Ustilago crameri, the process of biosorption was first demonstrated. Subsequently in 1922, the uptake of silver, copper, selenium, and mercury was reported by utilizing corn smut fungi. In 1949, Ruchloft reported that activated sludge was capable of removing radioactive metals like Polonium-239 from wastewater. The first patent of biosorption for wastewater management was awarded to Ames Crosta Mills & Company in 1973. In 1982, the first patent was awarded for removal of uranium and thorium ion from water (Karcher et al. 2001). From such remarkable achievements in the field of biosorption, we can definitely apprehend the importance of microorganisms and their metabolites in such processes. In 1978 Nakajuma reported broadly the order of metal uptake by different microorganisms as follows: actinomycetes>bacteria>yeast>fungi (Manu and Chaudhari 2003). However our focus of concern here will mainly be based on the diverse array of bacteria, their metabolites, the mechanism involved, and several other factors in context to biosorption of toxic dyes and heavy metals.

4.4

Involvement of Bacterial Metabolites in Biosorption

Bacterial metabolites can be categorized as primary and secondary metabolites. Secondary metabolites are independent of cellular growth and development. Bacterial secondary metabolites as already discussed above are involved in certain bacterial functionality that are not directly involved in the bacterial growth and development processes. Instead, they find their applicability in terms of the bacterial defense mechanism and are significantly functional during detrimental conditions of the bacterial vicinity. They are generally produced during the late exponential phase or during the stationary phase of bacterial growth. In context to biosorption of heavy metals, secondary metabolites like bacterial exopolysaccharides (EPS) prove to be an effective tool. There are other secondary metabolites involved in the process of bacterial metal biosorption like siderophores which are compounds secreted by the bacteria that exhibit a high iron binding affinity and are effective in chelating iron from the environment. However, by far, bacterial EPS have been most efficient as a secondary metabolite in adsorbing diverse array of toxic heavy metals and dyes due to its non-specificity in binding. Among the metabolites produced, exopolysaccharides happen to be a domain of current research in bioremediation of heavy metals and dyes. The term exopolysaccharide was coined by Sutherland (1972) to describe high molecular weight carbohydrate polymers produced by marine bacteria. They are mainly metabolites loosely attached to the cells or often secreted into the external environment. When provided with favorable conditions like medium composition (carbon and nitrogen), pH, and temperature, the cells

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produce these extracellular metabolites during late log (exponential) or at the onset of stationary phase (Sengupta et al. 2018). Sometimes, EPS are also termed as extracellular polymeric substances due to their chemical constitution (Nichols et al. 2005). In the environment a single bacterial cell cannot solely survive alone. Often these cells exist in aggregates and enclose themselves in a polymer matrix while adhering to a surface. These cellular aggregates with the polymer matrix are termed as “biofilm.” EPS build up the primary structural framework of the biofilm (Smirnova et al. 2010). They are popularly termed as “adhesive polymers” because of their maintenance of adhesion in between the cells and also to the solid surface (Czaczyk and Myszka 2007). Apart from the functional attributes like cell-to-cell adhesion, inhibition of cellular desiccation, and protection against toxic extracellular substances and antibiotics, bacterial secondary metabolites especially the extracellular polymeric substances, exopolysaccharides, and biosurfactants play effective role in the biosorption of heavy metals. In addition, EPS also prevents desiccation, provides resistance to antibiotics, and protects the cells against harsh conditions of pH, salinity, radiation, and also sequester nutrient in extreme starvation conditions. These polymers are predominantly made up of carbohydrates and other biomolecules like proteins, lipids, phospholipids, nucleic acids, etc. (Pal and Paul 2008). Due to the presence of these biomolecules, few charged species like succinate, phosphate, and acetate play an important role in biosorption process. These charged/ionizable groups impart ligand like properties to these polymers which enable them to bind efficiently with the charged groups of the dyes (Bhaskar and Bhosle 2006; Moppert et al. 2009). Generally, EPS are high molecular weight bacterial polymers which make them the ideal sorbents due to some of the attributes like possessing stronger van der Waals forces, more functional groups compared to that of low molecular weight sorbents. The more the linearity in the structure, the more ionizable/functional groups are present in EPS (Zhang et al. 2008). The intrinsic structural moiety of bacterial EPS depends largely on the genetic constitution of the producer bacteria as well as several physicochemical parameters (Sengupta et al. 2018). Several bacterial signaling cascades are involved in the synthesis and translocation of different types of bacterial EPS (Schmid et al. 2015). The Wzx/Wzy proteindependent pathway is involved in the synthesis of highly diverse types of bacterial exopolysaccharides (especially the heteropolymeric EPS). In this pathway several glycosyltransferases and especially the Wzx and Wzy protein come into play. The Wzx is involved in the translocation of the repeating units of EPS to the cytoplasmic membrane, whereas Wzy protein is involved in the polymerization followed by the transport of EPS polymers to the periplasmic space. The final transport to the extracellular medium is mediated by PCP (polysaccharide co-polymerase) and the OPX protein (outer membrane polysaccharide export protein). On the contrary, the homopolymeric EPS involves ABC transporter (ATP-binding cassette)-dependent pathway for synthesis and translocation. ABC transporter is involved in translocation of capsular polysaccharides. Also, the bacterial protein synthase is dependent in production of homopolymeric extracellular bacterial polysaccharides.

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The time involved in the biosorption of heavy metals varies from one bacterial strain to another. Certain bacterial strains like Methylobacterium organophilum exhibited remarkable adsorption of copper and lead ions non-specifically within 30 min at neutral pH (Gupta and Diwan 2016). Another interesting finding is that in certain cases the toxic heavy metals induce the formation of biofilm as well. The biofilm production by Herminiimonas arsenicoxydans, a Gram negative bacterium, depends on the presence of arsenic (Muller et al. 2007). Also the bacterial biofilm serves as effective biosorbent of arsenic. Similar is the role of biofilm produced by Thiomonas sp. in bioremediation of arsenic. The presence of the ionizable functional groups like carboxyl, acetyl, and phosphate as the heteropolymeric structural moieties of diverse types of bacterial EPS often imparts an overall polarized group for the cationic metal ions to bind. This imparts an active site in the EPS to bind metal ions. On the contrary homopolymeric EPS are polyanionic due to association of same repeating units forming the structural backbone of the EPS (Gupta and Diwan 2016). Thus heteropolymeric EPS are more efficient in biosorption of heavy metals in contrast to homopolymers. As already mentioned homopolysaccharides being less pronounced as a potential biosorbent often require the attachment of a matrix or a derivative for the desired purpose. Considering the first discovered bacterial homopolysaccharide dextran in such a context, we often find that dextran does not happen to be an appreciable biosorbent unless a matrix or a derivative is bound to it. For example, diethylaminoethyl dextran is a potential biosorbent for adsorption of heavy metal ions (Demirbilek and Dinç 2012) in an order of affinity Zn2+ > Mn2+ > Pb2+ > Cd2+. More and more bacterial strains are being discovered from the diverse plethora of natural resource that cater to efficient remediation of potentially toxic heavy metals. For instance, psychrotrophic arctic glacier soil borne Pseudomonas sp. exhibited a metal biosorption potency up to 90% in the order Fe2+ > Cu2+ > Mg2+ > Zn2+ > Mn2+ > Ca2+ (Sathiyanarayanan et al. 2016). The functional groups present in a bacterial exopolysaccharide predominantly involved in the process of biosorption of heavy metal and dyes are an important domain of study. Often techniques like Fourier transform-infrared spectroscopy or nuclear magnetic resonance of the exopolysaccharide might help in understanding and comparing the constitutional moiety involved as the biosorbent. This is especially important for relatively uncommon bacterial metabolite biosorbents. In case of exopolysaccharide of Bacillus thuringiensis, the FTIR studies revealed that amino, carboxyl, hydroxyl, and carbonyl groups may be predominantly involved in the biosorption of heavy metals. The biosorption was to a remarkable extent – nickel (94%), copper (91.8%), and cadmium (87%) (Oves et al. 2013). Similarly, in case of Alcaligenes sp., the FTIR spectra of the bacterial metabolite revealed the functional groups involved in biosorption (Jin et al. 2017). Similar is the case of fungal metabolites as well. For instance, in case of Verticillium insectorum, the FTIR results indicated that –NH, –CH, –OH, C¼O, and C–O groups of the exopolysaccharide participate in the biosorption of lead and zinc (Feng et al. 2018).

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EPS can sometimes be termed as extracellular polymeric substances and it’s wellknown for its remarkable characteristics thus serving as an excellent adsorbent when it comes to removal of dye molecules from a heterogeneous aqueous solution. Proteus mirabilis TJ-1 was inoculated into a production medium to stimulate EPS production. The dried EPS has shown excellent results in binding up of Basic Blue 54. The kinetic data so obtained deduces the initial sorption of the dye into the cell wall followed by internal assimilation hence suggesting that initial sorption is necessary for the intake of the dyes into the cell (Zhang et al. 2006, 2007, 2008; Jang et al. 2001). Furthermore, the presence of heavy metals in the bacterial niche either plays an augmentative or inhibitory effect on the production of bacterial secondary metabolite. The secondary metabolites are basically a class of compounds that are generally polymeric in nature and are involved in a wide spectrum of bacterial activity like conglomeration or adhesion of bacterial cells to one another promoting an architectural framework of biofilm. Often, the effectiveness of bacterial secondary metabolites in effective biosorption supersedes the potency of that of the dead bacterial biomass. Thus bacterial metabolites are an effective tool in bioremediation by lowering the toxicity of heavy metals from the environment. Another manifestation of heavy metal toxicity to bacterial cell is the eventual production of metabolites and polymeric substances to the extracellular medium acting as a defense mechanism. The extracellular polymeric substances that often bind and adsorb to the metal ions even before reaching the peptidoglycan layer render a change in the metal ion concentration that eventually enters the intracellular matrix. Thus the bacterial extracellular polymers behave as a blanket surrounding and protecting the cell from the toxic metals. Flocculants are used widely for their effectiveness is removal of organic pollutants in water bodies. They can be classified mainly into three subdivisions: (a) inorganic flocculants like aluminum sulfate and poly-aluminum chloride; (b) organic flocculants like polyacrylamide derivatives and polyethylene imine; and (c) naturally occurring flocculants like alginate, chitosan, and microbial sources (Salehizadeh and Shojaosadati 2001). However, their prolonged use results in huge sludge formation that can result in waste disposal problem. Additionally, alum poses harmful health effects and the polymeric residues formed are carcinogenic in nature. But on the other hand, the bacterial flocculants can mitigate environmental pollution without simultaneously affecting the environment. Because of their several disadvantages of both the inorganic and organic flocculants, bioflocculants can be a useful substitute. The phenomenon “bioflocculation” results from the synthesis of extracellular polymeric substances (EPS) or extracellular biopolymeric flocculants (EBPFs) which are generally produced by microorganisms present in activated sludge and soil. These microbial -flocculants are cheap, can be abundantly produced in large scale, are biodegradable in nature, and can be easily recovered from the fermentation broth. Also, the side products formed are not fatal to the environment making them a popular alternative (Salehizadeh and Shojaosadati 2001). In recent studies, marine bacteria Aerococcus urinaeequi HZ strain produces an EPS (EPS-A) in its fermentation broth which has shown excellent wastewater flocculation activity

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(70.90%) (Wang et al. 2018). The exact mechanism of biological flocculation has not been thoroughly investigated. However, a preliminary idea was reported that relies mainly on two ways: bridging and charge neutralization. Bridging occurs when the biopolymeric flocculent extends from the particle’s surface into the solution for a distance greater than the distance over which the interparticle acts. As a result of this, biopolymers adsorb to each other and form flocs. Charge neutralization occurs when there are differences between the polymeric surface and dye molecules (Wang et al. 2018; Levy et al. 1992; Hantula and Bamford 1991).

4.5 4.5.1

Factors Affecting Biosorption of Dyes by EPS or Bioflocculants Effect of Carbon Source and Initial Dye Concentration

The initial dye concentration plays a pivotal role in production of exopolysaccharide (EPS) and also its sorption activity. The bacterial strains Pseudomonas sp. and Ochrobactrum sp. showed maximum EPS production when media enriched with 100 mg l1 Remazol Blue was utilized by them. However, with higher concentration of the dye, the production of EPS by both the bacterial strains significantly decreased (Koçberber and Dönmez 2012). This effect might be due to the toxicity of the dyes on the overall bacterial metabolism. It has been reported earlier that the increased concentration of some dyes inhibits protein synthesis and cell division (Ogawa et al. 1988). This puts emphasis on the fact that when maximum dye concentration is present in the media, genes involved in EPS synthesis are aborted thus lowering the yield of EPS production. On the contrary, in a study by Zhang et al. it has been reported that the sorption activity of extracellular polymeric substance (EPS) produced by the strain Proteus mirabilis TJ-1 increases with increase dye concentration (Zhang et al. 2008). Stress might upregulate some of the genes involved in production of EPS which in turn aid in sorption of the dyes. Along with the concentration of dyes, the carbon and nitrogen sources of the culture medium also favor the overall bacterial growth and the production of EPS (Salehizadeh and Yan 2014; Li et al. 2009). In various studies, different carbon sources have been utilized for the production of bioflocculant or EPS (Zhang et al. 2008; Gong et al. 2008). The nature of the carbon sources utilized is mainly dependent upon the bacterial strain involved. In case of Alteromonas sp., the bioflocculant produced exhibits maximum flocculating activity (1515 U/ml) when glucose (30 g/l) was present in the fermentation medium. The bioflocculant so obtained was able to remove Congo red, Direct Black, and methylene blue at efficiencies of 98.5%, 97.9%, and 72.3%, respectively (Chen et al. 2017). In few investigations however, it has been reported that some unconventional carbon sources can also promote bacterial growth and subsequent production of EPS which can have excellent flocculating activity. Acinetobacter baumannii

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YNWH 226 was inoculated in a culture medium with Congo red as the sole carbon source. It was observed that the strain has a potential of producing EPS with a high flocculating activity and dewaterability on textile dyeing sludge (Li et al. 2015).

4.5.2

Effect of pH, Temperature, and Contact Time

pH of the solution directly effects the structural confirmation and charge of the dyes along with the polymer surface charge. As already mentioned before, the acidic and alkaline pH along with the dye’s nature controls the adsorption process (Salleh et al. 2011). Indigenous bioflocculant-producing bacteria isolated from activated sludge show maximum removal of the medi-blue and whale dyes at neutral pH conditions. However, the same bioflocculant shows maximum removal of mixed and fawn dyes at alkaline range (pH 10). It was suggested that the hydroxide group of the flocculant adsorbs the dye molecules in this state (Buthelezi et al. 2012). On the other hand, the bioflocculant produced by Alteromonas sp. shows optimal removal of the anionic dyes Congo red and Direct Black between pH range 3 and 11. Positively charge amide groups impart the cationic nature of the bioflocculant thus adsorbing the anionic dye molecules (Chen et al. 2017). Temperature is also an important parameter to consider and the range between 35 and 40
C was found to be optimum for the removal of textile dyes from wastewater effluents by the bioflocculant-producing bacterial isolates. When the temperature was increased further, the dye decolorization rate decreases. The reason for this phenomenon might be due to the loss of surface activity of the flocculant or loss of viability in case of living cells (Buthelezi et al. 2012). In most bacterial cells, the dye removal occurs rapidly in first few minutes and then equilibrium is attained and the adsorption rate decreases. This might happen due to saturation of the binding sites on the biomass surface (Zhang et al. 2008).

4.5.3

Effect of Bioflocculant or Extracellular Polymeric Substance Concentration

The dye removal efficiency increases with increasing the concentration of adsorbate. The capability of the bioflocculant produced by Alteromonas sp. was measured at different concentrations (20–200 mg/l) against the dyes Congo red, methylene blue, and Direct Black. The optimum bioflocculant dosages vary with the respective dyes. However, after attaining a saturation point, no further addition of the adsorbate decreases the dye removal efficiency. Although with high concentrations of the flocculant, the adsorption surface area increases, but a decrease in the activity after reaching a maximum value can be attributed to the destabilization of the flocculant structure due to repulsive forces between the small flocs (Buthelezi et al. 2012).

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Effect of Metal Stress

Out of all the heavy metals, the usage of heavy metals like cadmium, chromium, and copper is most common in textile and tanning industries. The heavy metals are mainly used during the dyeing, processing, and production of textile dyes (Deepali and Gangwar 2010; Halimoon and Yin 2010; Batool et al. 2014). Often the dyes are mixed with heavy metals to enhance color fastness and stability (Maddhinni et al. 2006). Both the heavy metals and dyes are recalcitrant compounds and highly toxic in nature. Some microorganisms show their resistance to the presence of heavy metals by employing several strategic mechanisms. The production of an extracellular polymeric substance (EPS) around them helps in protecting the cells (Anderson and O’Toole 2008; LeChevallier et al. 1988). Several reports have claimed the adsorption potential of the bacterial EPS in removal of heavy metals (Morillo et al. 2008; Nogueira et al. 2005; Zhou et al. 2010). Rhodotorula mucilaginosa UANL001 L when exposed to varied concentrations of transition and post-transition metals like cadmium, lead, zinc, nickel, copper, and chromium showed enhanced EPS production. The presence of zinc stimulated maximum EPS production by the strain. Moreover, the EPS produced was able to adsorb methylene blue twofold greater than the existing biosorbents reported in the literature (Gonzalez et al. 2016). In another study, EPS produced by Bacillus subtilis MB378 showed highest dye binding efficiency when malachite green dye was present in combination with chromium (0.66 mg/g) followed by cadmium (0.64 mg/gm) (Tahir and Yasmin 2019).

4.5.5

Effect of Salts

It has been reported in several studies that the bacteria are evolutionarily able to protect themselves from osmotic stressful conditions. The production of EPS is one of the survival mechanisms which act as a necessary protective barrier (Sengupta et al. 2018; Sandhya and Ali 2015; Wood 2015). The production of bioflocculants is also significantly influenced in the presence of cations (Salehizadeh and Yan 2014). Dyes are applied on fabrics in presence of salts to decrease dye solubility; hence the presence of salts in industrial effluents is quite ordinary. Hence, it is an absolute necessary to know whether the presence of salts interfere with the sorption process. It has been reported that the presence of salts MnCl2, CaCl2, MgSO4, and CTAB enhanced the dye removal capacities of bioflocculants by the indigenous bacterial isolates. The cations present play a vital role in neutralizing and stabilizing residual negative charges of functional groups and thereby forming bridges between particles. The trivalent or divalent cations increase the initial adsorption on suspended particles by decreasing negative charges on the surfaces of both the flocculant and the dye particle (Buthelezi et al. 2012; Levy et al. 1992). In a similar study, Alteromonas sp. shows highest flocculating activity 30–40 g/l of sea salt and exhibited removal of Direct Black, Congo red, and methylene blue dyes. However,

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too much salinity can reduce the flocculating activity which may be due to the fact that bacteria losses its activity after saturation (Chen et al. 2017).

4.6

Regeneration of Bacterial Cells: Desorption

The biomass or metabolite employed in the process of the biosorption should be regenerated for further reusability. Incineration and land filling of the spent biomass only amplifies the amount of accumulated waste. The regeneration of biomass is possible if the dye is desorbed from it. Desorption is the phenomenon of releasing an adsorbed material from the surface of the adsorbent. The process enables regeneration of any costly material thus reducing process cost and also dependency on the continuous supply of biosorbent. The process of desorption depends upon the type of the elutant, the nature of the biomass, and biosorption mechanism. A few parameters on choosing the suitable elutant must be checked like as follows: (i) should not interfere with the biomass, (ii) cost-effective, (iii) eco-friendly, and (iv) efficient (Vijayaraghavan and Yun 2009). Mao J. et al. performed several experiments with the same bacterial biomass Corynebacterium glutamicum but different natures of dyes. It was reported that, in case of biosorption of basic blue dye, solution was shifted to an acidic range (pH 3) where adsorption is lowest. This process enables complete regeneration which was efficient for up to four sorption/desorption cycles (Mao et al. 2009). Similar experiment was performed by utilizing same biomass but with an anionic dye Reactive Red 4. Complete regeneration of biomass was also obtained here by shifting the solution to an alkaline range and it was feasible for reusability for up to four cycles. Although shifting the pH by addition of acids or alkali ensures reusability of the adsorbent, some problems regarding deterioration of the biomass were observed (Vijayaraghavan and Yun 2007).

4.7

Mechanisms and Factors Involved in Bacterial Binding of Heavy Metals: At a Glance

The mechanisms employed by the bacterial cells for the metal affinity and binding are diverse and quite interesting to study. Broadly the metal binding activity by a viable bacterial cell can be categorized as direct and indirect mechanism. Direct mechanism involves the binding of the cationic metal ions to the bacterial peptidoglycan cell wall mostly by means of passive transport, while the indirect mechanism involves the activity of bacterial enzymes in binding of metals ions. The latter is governed by regulatory genes and transcription factors at large. Metal ions are often required by the bacterial cell as trace elements for several metabolic and biochemical reactions. However, when such ionic concentration surpasses a certain threshold level, the manifestation of their toxicity comes into play. The manifestation can

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either result in the inhibition of bacterial growth or overexpression of certain genes resulting in elevated production of enzymes involved in imparting resistance to such toxicity (Ferris et al. 1987). In certain cases of uncontrolled toxicity of heavy metals, the bacterial cell can accumulate and localize the metal ions in specific zones in cytoplasmic matrix which are referred as nucleation site that often results in eventual cell death (Ferris 2000). This suicidal mechanism of the cell imparted by certain bacterial species can often be regarded as a tool in the remediation of toxic heavy metals from the environment. Passive and active (metabolically driven) transport of heavy metal show contrasting differences in the context that metals like gold, iron, and others often exhibit passive microbe–metal interaction, while on the other the reduction of radioactive metals like uranium often involves the precipitation and degradation by gradual metabolic pathways (Kulkarni et al. 2016). The pattern of uranium biosorption is quite interesting to study. On comparing the adsorption by Bacillus pantothenticus, B. megaterium, Pseudomonas putida, and P. chlororaphis with both live and dead cells, the viable cells of B. megaterium and P. chlororaphis showed highest adsorption (20 μg/ml and 30 μg/ml, respectively). Both the strains exhibited uranium adsorption at relatively acidic pH of 6 and 4, respectively. This is indicative of the fact that although the strains show adequate growth at neutral pH, the adsorption of radioactive metals like uranium occurs at an acidic pH. This might be due to the fact that on thriving in acidic environment, the bacterial cell might render a cell surface polarity thereby expressing metabolically active proteins on the bacterial cell surface thus incorporating the passage of uranium in an energy dependent manner. Often uranium is retained in the membrane bound protein. This mechanism is well expressed in Deinococcus radiodurans and even in E. coli. Both the bacterial cells express two types of proteins in orchestration – PhoN (a periplasmic acid phosphatase) and PhoK (an extracellular alkaline phosphatase). Thus recombinant bacterial cells overexpressing phosphatases have been prospective tool in adsorption of a radioactive metal like uranium. It has been observed that both the strains showed six to tenfold more biosorption under carbonate-deficient condition under neutral pH in comparison to carbonateabundant, alkaline conditions (Suzuki and Banfield 1999). However, in contrast to bioaccumulation, the biosorption efficiency of uranium by bacterial dead cells happens to be much better at an extent of 615 mg per gram of dry cell mass (Strandberg et al. 1981). A comparison of the uranium uptake potency by Pseudomonas aeruginosa and Saccharomyces cerevisiae revealed that cell-bound uranium reached a concentration of 10–15% of the dry weight of cell, but only 32% of S. cerevisiae cells and 44% of P. aeruginosa cells possessed uranium deposits (Hu and Zhao 2007).

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Involvement of Bacterial Genetic Factors and Genetic Manipulation for Enhancement of Biosorption of Heavy Metals

In almost all prokaryotic cells, there is a definitive efflux system for maintaining a controlled inflow of metal ions and also for the prevention of excess uptake. This regulatory system is principally governed by a cascade of several protein families. The over- or under-expression of such regulatory proteins are in turn controlled by the transcription of the regulatory genes involved. Cadmium which is a relatively toxic heavy metal even at a very low concentration is regulated by the energydependent cadmium efflux mechanisms which involve a family of three major bacterial proteins – CBA-type chemiosmotic antiporter, P-type ATPases, and cation diffusion facilitators (CDF) (Silver 1996). Genes involved in the translation of such proteins are generally identified by insertion mutagenesis studies followed by inverse PCR (polymerase chain reaction). Since bacterial resistance to metal ions is a function that is not involved directly with growth or development of the bacterial cell, the loci of such genes are generally found in the extra-chromosomal DNA, most predominantly in the bacterial plasmid (Jain and Bhatt 2014). For instance, the czc gene located in the ~5 Kb plasmid of Pseudomonas sp. is involved in bacterial metal efflux mechanism conferring resistance to cobalt, zinc, and cadmium. Such significant genes often require amplification by molecular techniques like PCR and vector mediated gene cloning by the advent of recombinant DNA technology (Zhang et al. 2015). Apart from its prospect in remediation of oil spills, genetically engineered bacterial strains of Pseudomonas putida pose to be important tool in bioremediation of cadmium-contaminated soil (Diep et al. 2018). But prior to genetic engineering of potentially important bacteria, the necessary prerequisite is to acquire a comprehensive knowledge regarding the heavy metal ion import system of the viable cells. Largely bacterial transport system can be categorized into basic three types: channels (which are basically α-helical proteins), secondary carriers (which are single component protein – either uniporters, symporters, or antiporters in nature), and primary active transporters (these are multicomponent protein transporters). Generally, E. coli, Pseudomonas sp., and Streptomyces sp. follow the first mode of transport, while Helicobacter pylori and Staphylococcus aureus have secondary carriers for transport, and finally bacteria like Lactobacillus follow primary active transport (Péant et al. 2005). It can be stated broadly that bioaccumulation research began circa 1990 and is an energy-dependent process unlike biosorption. Furthermore, bioaccumulation is a slow process and also an irreversible one (Becker et al. 2002). Thus often the activity of bacterial metabolite and/or dead bacterial biomass comes into play in terms of remediation of heavy metals. As already mentioned the yield in bacterial exopolysaccharide (EPS) depends both on the physicochemical parameters of culture media and the genetic constitution of the producer organism (Barnett and Long 2017). If we study the chromosomal DNA of Lactobacillus rhamnosus, we find six potential genes that have been

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identified in the production of heptasaccharide repeat units with high proportion of rhamnose (Morris and González 2009). Similarly, in Sinorhizobium meliloti two types of EPS (succinoglycan and galactoglucan) have been identified that have potential in protection from abiotic stress and biofilm formation (Xia et al. 2018). However, the production of the EPS-1, one of the two acidic EPS, is dependent on protein coding gene emmD that codes for a part of EmmABC three-component regulatory circuit. Furthermore, mutations of such regulatory genes in the organism form mucoid colonies that show overproduction of EPS (Upadhyay et al. 2017; Jittawuttipoka et al. 2013). Even the genetic constitution of less familiar bacterial strains like Mucilaginibacter pedocola in conferring the bacterial resistance to and the adsorption of zinc and cadmium is being studied. The putative heavy metalresistant gene associated with the EPS production involved in metal adsorption has been located in the 3132 protein coding genes of the 7,035,113 bp bacterial genome. Similar is the case of the Pseudaminobacter manganicus which has been reported to exhibit multiple heavy metal resistance. The strain significantly remediates manganese and cadmium ions not only by itself but also by producing exopolysaccharide as a potent metal biosorbent. The 4,842,937 bp genome comprises 4504 protein coding genes and 71 genes in the bacterial RNA among which a large number of genes are involved in exopolysaccharide production and heavy metal resistance (Moorthy et al. 2015). Often mutational studies involving the deletion of potential genes help identify the genes involved in the synthesis of protein involved in regulation of bacterial EPS production. Here we can state the example of the fluorescent Pseudomonas strain Psd which in addition to being a plant growth-promoting strain, also express high zinc-adsorbing potency. The gene alg8 which codes for a subunit of Alginate polymerase interestingly showed significant reduction in the EPS production by the organism. Such findings were also supported by the fact that exposure to zinc lead to augmented alginate production thereby accentuating the formation of biofilm (Hu et al. 2018). Cyanobacterial species are also being widely studied for the detection of potential genes involved in the expression of proteins involved in the regulation of EPS that cater to protection against various toxic metal ions. This can be well estimated by the EPS-depleted mutants that have deleted sll1581 and slr1875 genes. Such mutants have exhibited impaired metal biosorption especially that of cadmium and iron (Polanska et al. 2014). Thus the EPS-depleted mutants can be exploited as important tool in the investigation of cell-to-cell aggregation leading to the formation of biofilm matrix which in turn promotes tolerance against environmental stress.

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Live Bacterial Cell, Dead Biomass, and Bacterial Metabolite in Bioremediation: A Comparison

Bioremediation can be referred to as the purposeful introduction of microorganisms for the utilization and breakdown of environmental pollutants. Environmental pollutants can be industrial waste like petroleum or crude oil wastes, chemical wastes, and wastes from textile tannery or related industries and finally range up to household wastes as well. In such a context, the basic constituents of waste happen to be a determining factor for the selection of the type of tool for bioremediation. For instance, too much exposure to polycyclic aromatic hydrocarbons (PAH) can lead to cancer, cardiovascular diseases, and poor fetal development (Chauhan et al. 2008; Xu et al. 2018; Dussud et al. 2018). The cleanups of such toxic substances by invasive technologies often prove to be expensive and even damaging to the natural resources. So implementations of bioremediation techniques mainly by viable bacterial strains have come out with appreciable results (Srinath et al. 2002). However, for this a thorough knowledge of the bacterial metabolic pathways and its relation with the utilization of PAH is necessary. Similar is the case of the bioremediation of petroleum hydrocarbons and aromatic hydrocarbons like benzene, toluene, ethylbenzene, or xylene by viable strains of Pseudomonas (Selatnia et al. 2004). Often bacteria utilize such complex hydrocarbons in order to breakdown into utilizable forms and in the process make the non-biodegradable wastes into a biodegradable counterpart. For instance, non-biodegradable and biodegradable plastics are being utilized by the marine bacteria (Karimpour et al. 2018). Thus the most important aspect pertaining to the bioremediation of wastes is the understanding of the nature of the waste and subjecting it to utilization by the related biomaterial. Toxic dyes, which are often a by-product of the textile or tannery industry when treated with oil-utilizing microorganisms, might not cater to sufficient degradation. Thus a suitable form of biosorbent is necessary. Toxic dyes from tanneries or textile industries often deposit substantial amounts of chromium that might get dissolved in the water bodies rendering them hazardous for consumption. On comparing the potential of living and dead cells of Bacillus coagulans in the remediation of chromium – Cr (IV), 23.8 and 39.9 mg Cr/gm dry weight was biosorbed, while on the other hand in case of living and dead cells of Bacillus megaterium, it was 15.7 and 30.7 mg Cr/gm dry weight (Srinath et al. 2002). Similarly, dead bacterial mass of Streptomyces rimosus happens to be effective biosorbent of lead in aqueous solutions (Selatnia et al. 2004). Remarkable efficacies of 87% and 98.5% biosorption of cadmium and lead, respectively, by dead mass of Pseudomonas aeruginosa were observed (Karimpour et al. 2018). Thus broadly it can be stated that in contrast to viable cells, often dead cells happen to be better biosorbents of heavy metals and dyes. Moreover, the remediation of metals by viable cells often requires a specific culture condition and certain physicochemical factors pertaining to the culture media. Such monitoring of the culture conditions is also a matter of concern and often incur certain costs. Furthermore, a specific bacterial strain might be specific in bioremediation of a specific type

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of heavy metal. So a sound knowledge regarding the bacteria–metal binding specificity is a prerequisite. On the contrary bacterial dead mass happen to be better alternative for biosorption of majority of heavy metals and dyes. It requires a comparatively lesser time for metal binding and remediation. Also the amount of metal adsorbed by bacterial cell mass is often higher than that by viable cells. However, the pretreatment of the bacteria like desiccation or heat treatment for biosorption by dead cell mass is an important criterion. Unlike viable cells, dead biomass often does not require specificity in the process of biosorption. Thus often bacterial cell mass can remediate multiple metals from a cocktail of potentially toxic wastes. The implementation of bacterial metabolites is by far the most scientific biotreatment of metal contaminants. This is due to the fact that certain bacterial exopolysaccharides are often having minimalistic chances of pathogenicity in contrast to viable cell-mediated treatment. Furthermore, after extraction of a bacterial metabolite like exopolysaccharide, the judicious purification is a necessary step. The purification often involves repeated treatment by ethanol and/or acetone followed by centrifugation and other techniques. The metabolites that act as building blocks of bacterial biofilms can be considered as a filter binding the metals thereby might be a potential alternative in purification processes. A comparative study on biosorption of the anionic dye Procion Red MX 5B by all these forms of adsorbent was conducted to provide some insight into the process. It was reported that the dried EPS and the free cells showed better sorption efficiency than the other two, i.e., immobilized cell or immobilized EPS. The immobilization done with the aid of calcium-alginate beads increases the mass transfer resistance between the adsorbent and adsorbate thus lowering its performance (Binupriya et al. 2010). The use of bacterial metabolite is an upcoming tool that is low in cost and also these can be produced by bacterial utilization of low-cost, household, or even waste substrates like bagasse, rice husk, starch, molasses, jaggery extract, fruit pulp, and others. Thus the cost incurred in the production of metabolites like exopolysaccharides is quite low and economic in comparison to other filters and purifiers involving techniques.

4.10

Future Scope of Bacterial Metabolite in Bioremediation

Biosorption offers a new avenue in tackling the problems involving the treatment of industrial effluents, polluted lands, and municipal wastewaters where the conventional technologies have proven to be quite expensive, time consuming, and inefficient. Much focus has been given to the bacterial enzymes which aid in degradation of the pollutants. However, the potential of only the bacterial cells and their metabolites like EPS in energy-independent passive binding of the dyes was largely unexplored. But, several recent reports have finally revealed the efficiency of the bacterial cells and their metabolites in sorption of several dyes and heavy metals

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used in industries. In spite of their efficiencies, the bacterial biosorption processes and their background require further detailed information. The exact amount of bacterial biomass required for a particular dye or metal removal needs to be quantified before performing the experiment. Also, the optimum physicochemical parameters for better removal of dyes and metals need to be observed. Further studies are required regarding genetic manipulation of the bacteria ensuring higher biosorption capacity. Often the real conditions differ significantly from the laboratory conditions. So, industrial effluents can be taken into consideration when performing the experiments. In that way, the bacterial cells and its metabolites can be commercialized for the bioremediation of the dyes and other pollutants in the future.

4.11

Conclusion

Bioremediation is an eco-friendly, cost-effective approach in mitigation of toxic substances like heavy metals and dyes that are not readily degradable and often requires invasive measures that are not cost worthy or practical. Thus biomaterials especially whole bacterial cell or dead bacterial cell mass or even bacterial metabolites stand out to be an appreciable alternative to such measures. Often we find biosorption to be a better alternative than bioaccumulation in the process of bioremediation of metals and dyes as the former is usually energy-independent and requires less pronounced physicochemical conditions. However a thorough study regarding bacterial genetic constitution for production of bacterial metabolite as a potential biosorbent is necessary. Acknowledgment The authors are grateful to the Department of Chemical Technology, University of Calcutta, for contemplating the study.

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Chapter 5

Bacterial Biofilms in Bioremediation of Metal-Contaminated Aquatic Environments Rafig Gurbanov

and Feride Severcan

Contents 5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2 The Collective Behavior of Microbial Cells: An Overview of Bacterial Biofilms . . . . . . 5.3 The Communication Between the Bacterial Cells: The Common Concepts of Quorum Sensing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4 The General Aspects of Bacterial Biofilms in Environmental Cleanup . . . . . . . . . . . . . . . . . . 5.5 Bacterial Biofilm-Assisted Strategies for the Decontamination of Metal-Polluted Aquatic Environments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.6 Conclusion and Future Prospects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract The main focus of this chapter is to review the current information on the application of bacterial biofilm-based approaches in the remediation of metalpolluted aquatic environments. In a present day, the metal pollution due to anthropogenic and mostly industrial activities possesses growing health threats to all aquatic and terrestrial life including humans. From the environmental health perspective, the main strategy is to reduce the availability, toxicity and spread of metal contaminants from the natural habitats. In view of this, the microbial bioremediation approach that is the utilization of wild or recombinant microorganisms to degrade or

R. Gurbanov (*) Department of Molecular Biology and Genetics, Bilecik Şeyh Edebali University, Bilecik, Turkey Biotechnology Application and Research Center, Bilecik Şeyh Edebali University, Bilecik, Turkey e-mail: rafi[email protected] F. Severcan Department of Biological Sciences, Middle East Technical University, Ankara, Turkey Department of Biophysics, Faculty of Medicine, Altinbaş University, Istanbul, Turkey Biomedical Sciences Graduate Programme, Altinbaş University, Istanbul, Turkey e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 117 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_5

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sequester the environmentally dangerous chemicals presents promising scientific and thus technological achievements. The bacterial cells are commonly specified as planktonic, whereas in natural habitats, bacteria manage to exist as communities by settling on the surfaces inside the biofilms. Here, we summarize the common mechanisms standing behind the bacterial biofilms and bacterial regulatory circuit so-called quorum-sensing system and precautions to consider for the actuation of this unique collective behavior in the bioremediation of metals from the aqueous ecosystems. In this chapter, first, knowledge on the collective behavior of bacteria is given, and then, the characteristic mechanisms of crosstalk between bacteria are explained. In the following section, the general aspects of the utilization of bacterial biofilms in bioremediation are mentioned. In the final section, the current applications and future potential of biofilm-based strategies for the remediation of metal contaminated aquatic environments are deliberated. Keywords Aquatic environments · Bacteria · Bacterial swarm · Biofilm · Biosurfactant · Bioremediation · Exopolysaccharide · Metals · Pollution · Quorum sensing

5.1

Introduction

Bacterial cells in their natural environment tend to cooperate, and by remodeling their cumulative behavior, these communities form a slimy initiation known as a biofilm. In turn, the biofilm formation is strictly controlled through the molecular mechanisms resulting in the unique communication pattern between the cells. This intercellular signaling pathway is named as quorum sensing and materialized through the small chemicals sensing the required cell density. After sensing this quorum, the genetic traits upregulating the communal behavior under a particular environment are expressed. This kind of bacterial socialization is crucial for the creation, composition, and coordinative behavior of biofilm-forming microbial units (Shrout and Nerenberg 2012). Accordingly, knowledge on the collective behavior of bacteria is presented in the scope of Sect. 5.2, whereas the characteristic mechanisms of crosstalk between bacteria so-called quorum sensing are explained in Sect. 5.3. Composed of water, exopolysaccharides, and microbial cells, the biofilm-based bioremediation technologies are exciting research areas in environmental science. In this regard, the biofilm exopolysaccharides with anionic character should be emphasized considering their strong complexation abilities with cations. Exopolysaccharides also play a determining role in the attachment of the cell to solid surfaces which are the very first stage of this genesis. Indeed, both of these functions are influenced by the quorum-sensing circuit (Bellich et al. 2018; Berne

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et al. 2018). The general aspects of the utilization of bacterial biofilms in bioremediation are considered in Sect. 5.4. Occupying a very large proportion of our planet, the aquatic ecosystem is a crucial environment for all the living kingdom. As an inevitable end of the developing civilization, the anthropogenic and industrial toxicants largely pollute our aquatic ecosystem. The first-level environmental threat for water resources is chiefly heavy metals and radionuclides. Scientists are in search of economical, fast-acting, and sustainable remediation of metals from polluted environments. In this regard, the cleanup technologies coupled with the bacterial biofilms are promising (Dar and Bhat 2020). Therefore, in the final Sect. 5.5, the current applications and future potential of biofilm-based strategies for the remediation of metal contaminated aquatic environments are deliberated.

5.2

The Collective Behavior of Microbial Cells: An Overview of Bacterial Biofilms

Long since, the bacteria are believed to be unicellular organisms, and the collective behavior of the entire population considered as the summation of all the effects of these cells. Over the last decade, this microbiological paradigmatic opinion has been abandoned, and today the scientists have evidential clues on how a variety of small molecules pave the web of communication between the bacterial cells (Parsek and Greenberg 2005; Jeckel et al. 2019). Bacterial swarms and biofilms are interesting phenotypes which are considered as totally different phenomena since the regulatory genes and happening cellular events are unique processes. In addition, the swarms include extremely mobile bacteria moving cooperatively to explore the environment for food, while the biofilms are consisting of motionless cells kept together on the support of extracellular matrix (Zhang et al. 2010; Flemming et al. 2016; Jeckel et al. 2019). The biofilms are ordinary life form of bacteria which are extracellularly and autogenously generated structural and functional microbial communities with particular integrities formed and attached or floated in a wide range of solid and aqueous environments, biological tissues or fabricated surfaces (Vasudevan 2014; Berne et al. 2018). The live microorganisms inside the biofilm organization refine their duties to work in synergy for the modulation of growth activity in support of the collective behavior. The behavior inside these communities is strictly regulated through the genetic mechanisms generating a systematically well-organized biofilm bundle rather than the dispersed organization (Flemming and Wingender 2010; Vasudevan 2014). Although biofilms are defined as extracellular organizations sometimes are considered as internal microbial constituents because of their roles in protection, nutrient supply, the regulation of cellular metabolism, and cell-to-cell interaction (Vasudevan 2014).

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In the development of biofilms on solid surfaces, the bacterium itself has first to gain very close contact with a solid surface, and only then the attachment process starts. In the case of aquatic biofilms, a slightly different but unique strategy is employed. In aquatic ecosystems, the diverse range of bacterial metabolites and environmental minerals and organic substances physically precipitates on the surface to create so-called conditioning films. The content of conditioning film is very diverse since any compound present in the aquatic environment can be precipitated and participated in the film development. These compounds alter both the physical potential and chemical charge of surface. By this way, convenient attachment and binding sites for bacteria are generated through the formation of strong interactions between the bacterium and chemical groups like phosphates of the conditioning film (Dunne 2002; Geng and Henry 2011; Lorite et al. 2011; Friedlander et al. 2013; Berne et al. 2018). The typical biofilm is composed of 97% water, 2–5% microbial cells, 1–2% polymeric compounds such as polysaccharides, 1–2% proteins and enzymes, 1–2% DNA and RNA, and ions in the free or bound state (Dar and Bhat 2020). Biofilms are estimated to be up to 90% of the total carbon content of the matrix (Vasudevan 2014). Primarily composed of polysaccharides, the polymeric compounds are responsible for matrix integrity. In addition to the roles in horizontal gene transfer and cell-to-cell communication, polymeric compounds also furnish the cells with essential nutrients and protect them from environmental threats (Oliveira and Cunha 2008). The diverse range of polysaccharides takes part in all the attachment, formation, and maturation steps in biofilm architecture (Limoli et al. 2015). Because of a complex structure, polysaccharides often considered as “dark matter of biofilm” (Shukla et al. 2014). Of course, under different environmental conditions, the polysaccharide biochemistry is also modulated to reshape the biofilm performance. Accordingly, the biofilms are acknowledged to be “an active tool in the hands of bacteria” rather than the lifeless matrix housing bacterial colonies (Bellich et al. 2018). Structurally and functionally these major biofilm elements are divided into aggregative, protective, and architectural polysaccharides. Generally speaking, aggregative polysaccharides serve as adhesive substances to build a cohesive scaffold with the surfaces in which the bacteria are trapped. As a result, the settlement and expansion of bacterial colonies are enhanced, and physical pressure generated by motion dynamics of water is endured. Protective polysaccharides serve to confront a variety of environmental stresses like drought and immune cells, whereas architectural ones laminate the colonies to give a stratum or thickness and create a food and by-product densities (Limoli et al. 2015). The charge characteristics of biofilm polysaccharides are also variable depending on the prevalence of Gram-negative or Gram-positive bacterial species. In the acquisition of net charge, the other bound molecules like uronic and teichoic acids and calcium and magnesium ions play an important role (Donlan and Costerton 2002). In addition, biomolecules such as proteins and nucleic acids are also included in the biofilm content. It is relevant to note that the biofilm scaffold is abundant with water channels that form hydrogen bridges with multilayers of bacterial

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microcolonies and provide food flow, oxygen, and microbial motion throughout the intact biofilm (Tolker-Nielsen and Molin 2000). Thanks to the moisture environment granted by water channels, the risk of evaporation or drought is also curtailed presenting a growth-promoting condition for microbial life (Vasudevan 2014).

5.3

The Communication Between the Bacterial Cells: The Common Concepts of Quorum Sensing

Communication is vital to the progress of coexistence, and in bacteria, this communication is known as quorum sensing. Being a fundamental and common feature of bacterial life, it includes the exchange of small chemicals between bacteria (Reuter et al. 2016). To define, it is an intercellular communication process based on the generation, secretion, and detection of autoinducer signals to regulate gene expression in response to changes in population density (Hawver et al. 2016). As the bacterial population density increases, the autoinducers accumulate in the surrounding environment (Papenfort and Bassler 2016). When a certain cell population density, that is, autoinducer concentration, is reached, the gene expression program of the bacterial cells is altered, and transcription of certain genes is switched on or off. Thus, bacteria can regulate genes that are advantageous for survival, while adapting their behavior to a variety of environments (Reuter et al. 2016). For example, quorum-sensing-regulated processes such as bioluminescence, secretion of virulence factors, sporulation, and biofilm formation are ineffective when performed by a single bacterial cell but operative when performed by a group (Bassler and Losick 2006). Quorum sensing is utilized by a variety of Gram-negative and Gram-positive bacteria. Regardless of their genus differences, all bacterial quorum-sensing systems fulfill the following basic principles (Vadakkan et al. 2018). In the cell densitydependent response to autoinducers, first autoinducers are secreted extracellularly, and their quantity continues to increase as bacterial populations develop (Kaplan and Greenberg 1985; Hawver et al. 2016). In the cell wall of bacteria, the specialized receptors are recognizing and responding to the autoinducer concentrations. Subsequently, the interaction of these receptors with the autoinducers initiates the mechanisms of quorum-sensing circuit (Novick et al. 1995; Seed et al. 1995). Despite this similarity, the autoinducer molecules differ between Gram-negative and Grampositive bacterial species (Reuter et al. 2016). The key elements of the quorumsensing system, autoinducers, are examined in three different classes according to their structure and particular functions. These classes are acyl-homoserine lactones, autoinducing peptides, and autoinducer-2 (Hense and Schuster 2015). Acylhomoserine lactones are small diffusion molecules with core-negative lactone rings and acyl side chains, responsible for facilitating signaling in Gram-negative bacteria (Manefield and Turner 2002).

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In Gram-positive bacteria, quorum sensing was found to be mediated by autoinducing peptides, which are short peptide chains produced intracellularly. Autoinducing peptide requires special membrane transport proteins since it is not transported freely across the cell membrane (Kolibachuk and Greenberg 1993; Sturme et al. 2002). On the other hand, autoinducer-2 is a furanone-derived signaling molecule found in both Gram-negative and Gram-positive bacteria (Coulthurst et al. 2004), demonstrating both acyl-homoserine lactone and autoinducing peptide properties (Xavier and Bassler 2003). In spite of their functional and structural differences, autoinducers have some common properties, such as a high receptor specificity and transport that may be active or passive through the cell membrane (Vadakkan et al. 2018). Quorum-sensing systems exhibit several mutual characteristics in all Gramnegative bacteria (Ng and Bassler 2009). First, autoinducers in such systems are acyl-homoserine lactones or else compounds derived from S-adenosylmethionine which are easily diffusible from the membrane. Subsequently, autoinducers interact with specific receptors, either found in the interior part of the membrane or the cytosol. Next, quorum sensing modifies many genes that underlie diverse cellular events. Finally, in a mechanism named as auto-reduction, the auto-stimulated activation of quorum sensing increases the synthesis of the auto-stimulant, which creates a proposed feed-forward cycle to stimulate synchronized gene expression in the community. Gram-negative bacterial acyl-homoserine lactones are the greatest group of autoinducers. Numerous Gram-negative Proteobacteria species possess luminescence R-type transcription factors interacting with acyl-homoserine lactone signals (Manefield and Turner 2002). Since there is excessive specificity between these proteins and cognitive acyl-homoserine lactone signals, these systems are predominantly used in cell-to-cell communication (Waters and Bassler 2005). The luminescence systems have a core N-acylated homoserine lactone ring and an acyl chain of 4–18 carbons which may include modifications. The length of the acyl chain may affect the signal specificity (Von Bodman et al. 2008). The luminescence I-type signal synthase enzymes are present in many bacteria yielding acyl-homoserine lactones (Case et al. 2008). The synthases yield acyl-homoserine lactone by getting the lactone moiety from S-adenosylmethionine, and often, the specific acyl chain is acquired from the metabolites of fatty acid synthesis pathway (Vadakkan et al. 2018). Structural studies of these enzymes demonstrated that each has an acyl binding pocket that fits precisely to a particular side chain fragment (Watson et al. 2002; Gould et al. 2004). This structural property apparently provides specificity in signal generation. Therefore, each enzyme produces a high quality, accurate signaling molecule (Waters and Bassler 2005). The luminescence R-type proteins also have acyl binding pockets that allow each transcription factor to bind only acyl side chains and therefore only be activated by own simultaneous signal (Vannini et al. 2002; Zhang et al. 2002). Therefore, in mixed species environments where multiple acyl-homoserine lactone signals are present, each species can only distinguish, measure, and respond to its own signal. When a certain threshold level attained, the acyl-homoserine lactone diffuses from the cell. At high concentrations, it is returned to the cell by interacting with luminescence R-type proteins. Non-acyl

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homoserine lactone-bound, free transcription factor decomposes in the bacterial cell (Zhu and Winans 2001). The transcription factor – acyl homoserine lactone complex – binds to the luminescence promoter region to initiate quorum-sensingregulated functions (Stevens et al. 1994; Devine et al. 2006). Gram-positive bacteria commonly employ oligopeptides and two-component signal transductions composed of membrane sensory kinase receptors and cytosolic transcription factors regulating the changes in gene expression (Papenfort and Bassler 2016). Autoinducing peptides are produced as pro-autoinducing peptides in the bacterial cell and are processed and replaced within or outside the cell membrane of the organism (Okada et al. 2005). Since the autoinducing peptides are impermeable across the cell membrane, their transport is generally mediated by membrane sensory kinases (Lyon et al. 2002; Bouillaut et al. 2008). The cell-to-cell signaling system, found uniquely in Gram-positive bacteria, is based on the prototypic accessory gene regulator system, first described in Staphylococcus aureus. This unique Gram-positive system uses a polypeptide signal rather than a small molecule. These polypeptides play a dual role. Polypeptide signal acts as an autoinducer for the organism that produces it while exhibits an inhibitory effect against other organisms (Parker and Sperandio 2009).

5.4

The General Aspects of Bacterial Biofilms in Environmental Cleanup

Nowadays environmental pollution has become the number one problem in the earth, due to the increasing urbanization and the advancements in industry and economy (Das 2014). Along with the presence of increasing amounts of toxic contaminants, there is a growing consciousness to develop efficient technologies to cleanse natural ecosystems (Megharaj et al. 2011). The application of microbial communities for biological degradation of dangerous pollutants to a lesser toxic or harmless substance is called bioremediation (Das 2014). Despite our theoretical knowledge and field applications, the technology is still in its infancy. As known, the essential properties of biofilm are controlled through quorum-sensing mechanisms. These mechanisms coordinate the different stages of bacterial biofilm development such as horizontal gene transfer, catabolic repression, the production of exopolysaccharides and cell surface bioactive substances, mobility, and chemotaxis which are crucial target events in collective behavior to explore for the advancement of environmental cleanup technologies (Huang et al. 2013; Mangwani et al. 2017). The bacteria with the ability of biofilm production enhance also the bioremediation success, on behalf of the synthesis of biofilm components like exopolysaccharides and biosurfactants (Dash et al. 2014). Therefore, the potentials of bacterial biofilms are enormous as it can be coupled with the cleanup technologies for bioremediation purposes. Several studies have already shown this potential in which the industrial phenolic pollutants were utilized as carbon sources by biofilms of Pseudomonas

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fluorescence and, as a result, reduced to less toxic forms (Sgountzos et al. 2006; Vasudevan 2014). It is worth mentioning that wild-type bacteria-originated biofilms possess a high chance of success compared to the predefined laboratory strains due to their better adherence capacities to wide range surfaces (Costerton et al. 1995). Therefore, the extent of molecular changes occurring in aquatic isolates of Bacillus and Pseudomonas species and laboratory strain of Escherichia coli ATCC 8739 acclimated for growth in inhibitory concentrations of cobalt was compared in a previous study by our group (Kardas et al. 2014). The contrasting changes in molecular parameters between the environmental isolates and laboratory strain shed light on how distinct genera of bacteria develop adaptations for survival in heavy metal-polluted aquatic environments (Kardas et al. 2014). Recently, we have shown the different adaptation mechanisms between the Gram-negative, namely, Escherichia coli ATCC 8739, and Gram-positive, namely, Staphylococcus aureus ATCC 6538 strains in response to silver exposure (Gurbanov et al. 2018). To note, we have uncovered the opposite epigenetic modulations between these bacteria that are 34% hypomethylation in Escherichia coli and 40% hypermethylation Staphylococcus aureus, in response to silver exposure (Gurbanov et al. 2018). Most recently, we have also deliberated the different methylation profiles and conformational forms in DNA of freshwater isolate Gordonia species acclimated to live and grow at high concentrations of heavy metals like cadmium, lead, and silver (Gurbanov et al. 2019). Biofilms are verified for wastewater refinement systems in which the pollutants are distilled from the polluted water stream by biofilm layers located in the special filters (Vasudevan 2014). The biofilms are also contributing to the maintenance of harmony in aquatic environments as biofilms precipitate the metals from the water except for the oxygen and degrade the plant and fish remains (Hunter 2008). Recently, the decomposition of polycyclic aromatic compounds, such as phenanthrene and pyrene, was achieved via the utilization of marine isolate of Pseudomonas aeruginosa N6P6. It was found that the upregulation of acyl-homoserine lactone synthase genes intensifies the production of exopolysaccharides and eventually biofilm formation (Mangwani et al. 2015). The biofilm of another marine isolate Stenotrophomonas acidaminiphila NCW-702 enhanced the decomposition of phenanthrene and pyrene in the laboratory environment (Mangwani et al. 2014). Naphthalene decomposition was achieved by the biofilms of Pseudomonas stutzeri T102 in a soil microcosm (Shimada et al. 2012). Several studies used different types of biofilm bioreactors for the degradation of persistent industrial chemicals like chlorinated hydrocarbons, which are the most common pollutant of soil and groundwater (Mangwani et al. 2017). The nondegradable plastics such as polyethylene are another dangerous contaminant that is extensively used for almost all domestic and industrial purposes. The microorganisms or their enzymes are able to decompose the plastics, but the process takes place quite slowly. In addition, polyethylene is first needed to be oxidized with the help of physical and chemical processes (Chatterjee et al. 2010). On the contrary, the biofilm of Pseudomonas species AKS2 accomplished the degradation process in only 45 days without any preprocessing step (Tribedi et al. 2015). Other persistent pollutants are nitro-aromatic substances,

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which are extensively used in the industries producing textile, agrochemicals, and pharmaceuticals. The biofilm bioreactors were found to be 90% efficient for the instant removal of dinitrotoluenes (Lendenmann et al. 1998). More comprehensive information on the biofilm-associated quorum-sensing mechanisms and their implications in bioremediation of environmental contaminants was recently presented by Mangwani and colleagues (Mangwani et al. 2017). The production of surface bioactive compounds or biosurfactants is one of the strategies of microorganisms that is important for access to surrounding hydrophobic chemicals (Marchant and Banat 2012). Biosurfactants consist of a range of substances such as glycolipids, lipoproteins, polymers, neutral, and ionic lipids. Glycolipids and polymeric ones succeeded in the decontamination of organic contaminants, while the lipoproteins and ionic lipids generally suited for pollutants (Pacwa-Płociniczak et al. 2011). The lipopolysaccharides of Gram-negative bacteria are important for attachment to surfaces. The acyl-homoserine lactone synthase gene regulates the rhamnolipid biosynthesis in Pseudomonas aeruginosa which is confirmed to be useful in bioremediation as rhamnolipids increase the availability of hydrophobic oil pollutants (Wang et al. 2007). Sphingomonas species exhibited 99% decomposition of polycyclic aromatic compounds in the presence of biosurfactant. The decomposition of diesel from soil was found as 77% efficient with the help of the cocktail of different bacterial species augmented with rhamnolipid (Owsianiak et al. 2009; Mangwani et al. 2017).

5.5

Bacterial Biofilm-Assisted Strategies for the Decontamination of Metal-Polluted Aquatic Environments

Microorganisms have an ability to reduce the metal ions by the oxidation–reduction reactions, for their energy metabolism, thus eliminating their toxicity. During the anaerobic respiration of bacteria, oxidized environmental metals act as electron acceptors and reduced to the less toxic form, as can be exemplified for mercury and chromium reduction (Igiri et al. 2018). Although numerous microbial species have been utilized for the purpose of heavy metal removal from water as well as soil, their efficiency is still limited because of the exceeding amounts of metal concentrations in an environment which is beyond the tolerance levels of a particular microorganism. Even genetically engineered and metal-resistant microbial species isolated from extreme habitats have their tolerance limits to withstand high concentrations of these contaminants. The practical outcome is inefficiency in bioremediation because of the eventual microbial death. Biofilm mode of growth plays an outstanding role in the remediation of metal contaminated aqueous ecosystems since the threshold levels of this consortium are much greater than freely suspended cells that are the planktonic mode of growth.

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Apart from being pollution-sensitive biological indicators, the bacterial biofilms can be implicated to remediate the heavy metal as well as radioactive metalcontaminated environments, in which understanding the reciprocal relationship between the bacteria and particular metal is important. Recent and past studies were targeting this cooperation extensively (Singh et al. 2006). The complexation of simultaneously enhanced exopolysaccharides with metals is beneficial to entrap the metals from the environment. It was shown that uranium salts can be precipitated with their interactions with the lipopolysaccharides of the bacterial cell wall (Macaskie et al. 2000). Besides, sphalerite deposits were developed in a native biofilm of sulfate-reducing bacteria (Labrenz et al. 2000). The biofilms are used as a component of urban wastewater purification systems for decades. In addition to wastewater pools and wetlands, the widely involved unit is biofilm-assisted bioreactors. The different types of biofilm bioreactors exist such as trickling and denitrification filters, submerged fixed bed biofilm reactor, aerobic moving bed biofilm reactor, denitrification fluidized bed biofilm reactor, membrane biofilm reactor, and biofilm-based microbial fuel cell (Dar and Bhat 2020). The absorption capacity of a reusable biofilm-assisted rotating biological contactor was demonstrated in the decontamination of synthetic wastewater containing cadmium, copper, and zinc (Costley and Wallis 2001; Singh et al. 2006). Previously, the microbial fuel cell was also shown to be successful in the reduction of chromium from the electroplated wastewater containing hexavalent chromium and in the production of electricity (Li et al. 2008). The recent laboratory-scale study demonstrated the 95% effectiveness of microbial biofilm in the removal of mercury, copper, and lead compared to corresponding planktonic culture (Grujić et al. 2017). Biofilms were also proved to be more tolerant against copper, lead, and zinc ions with respect to a planktonic cell in a bioreactor system (Teitzel and Parsek 2003). In another study, the effectiveness of biofilm-assisted bioreactor was compared with the common membrane bioreactors in the purification of wastewater. The biofilm-assisted one demonstrated a 6% increase in the removal of total nitrogen (Subtil et al. 2014). Furthermore, the continuously stirred bioreactors based on the biofilms of metalabsorbing bacteria were successfully employed to purify the heavy metal from the wastewaters (Acheampong et al. 2010). In this context, the process so-called the metal removal by sand filter inoculation was designed in which the biofilms of metalprecipitating bacteria were inoculated on a filter support composed of sand. During the contact with industrial rinsing water containing high concentrations of zinc ions, the biofilm absorbs and precipitates the metal ions. As water purifies in this wise, the metal-loaded top layer of biofilm is discarded from the sand matrix, and the left bottom layer can be grown again for the next loop of events (Pümpel et al. 2001). The continuous system composed of extractive membrane bioreactor coupled with sulfate-reducing bacteria was prospered in the remediation of zinc-containing wastewaters with at least 90% efficiency (Chuichulcherm et al. 2001). Previously, copper was also concentrated as a copper sulfide using the biofilms of sulfate-reducing bacteria in a continuous system (White and Gadd 2000). Although the biofilmassisted bioreactors hold several restraints such as high cost, the need for regular surveillance and frequent care, and fouling of pipes, these conditions can be resolved

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by periodic cleaning, rubbing, and other applications. The biofilm-assisted bioreactors are the best available technologies for water purification systems because of their ecological safety, authenticity, and performance (Dar and Bhat 2020). Numerous studies have demonstrated the potential of biofilm grown on a variety of surfaces and biofilm exopolysaccharides in the bioremediation of heavy metals (Pal and Paul 2008). Biofilm tolerance against heavy metals can be attributed to increased bacterial exopolysaccharide production under metal stress. In our previous study, we have shown the enhanced production of polysaccharides in the lead and cadmium-resistant Gram-negative and Gram-positive bacteria (Gurbanov et al. 2015). Another most recent study conducted by our group revealed an elevated production of exopolysaccharides in heavy metal-acclimated Brevundimonas, Gordonia, and Microbacterium species isolated from the aquatic environment (Kepenek et al. 2019). The exopolysaccharide production was found higher under acute cadmium exposure, in contrast with a gradual one. In the case of lead exposure, a different picture was observed, in which the gradual exposures were found more effective than those of acute ones (Kepenek et al. 2019). Our data are valuable for the development of efficient metal removal strategy by triggering the production of polysaccharides and accordingly biofilm (Gurbanov et al. 2015; Kepenek et al. 2019). Different marine strains of Pseudomonas aeruginosa indigenously expressing cobalt, zinc, and cadmium efflux system genes are showing high resistance and efflux features against the cadmium by forming heavy biofilms (Zeng et al. 2012; Chien et al. 2013; Chakraborty and Das 2014). It was proven that the cadmium tolerance takes over these efflux pumps and subsequently is removed out by binding to biofilm-related exopolysaccharides (Chakraborty and Das 2014; Chellaiah 2018). The exopolysaccharides can act as a biofilm safeguard to reduce metal-driven oxidative stress whether by binding and neutralizing them or diminishing their dispersion inward the matrix, which can be accounted for the aforementioned high metal resistance properties of biofilms (Pal and Paul 2008). The factors affecting the interaction of exopolysaccharides with metals are enclosing pH, the amount of metal, organic substances, and cell density. The carbon to phosphorus ratio is also crucial to measure the biofilm growth for the reduction of heavy metals from the wastewater. When exopolysaccharide interacts with metal ions like copper, lead, and nickel, the carbon to phosphorus ratio decreases (Jang et al. 2001). A bacterium producing enhanced biofilm, Herminiimonas arsenicoxydans, isolated from the contaminated aquatic environments was reported to diminish high-level arsenic contamination by sequestering arsenic ions with the help of their exopolysaccharides (Weeger et al. 1999; Marchal et al. 2010). Enhanced generation of surface biofilms and arsenic ion accumulation was reported in Thiomonas species CB2, exposed to arsenic (Marchal et al. 2011). Herein, the exopolysaccharide of marine bacteria is interesting as these exopolysaccharides possess high uronic acid, fortifying the anionic character of the whole biofilm for complementation with metal cations (Gupta and Diwan 2017). High capacity metal-chelating exopolysaccharides were also extensively characterized in Cyanobacteria species. As an example, the highest affinity for manganese ions was shown in exopolysaccharides of Anabaena spiroides (Freire-Nordi et al. 2005).

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The immobilization of bacterial cells to different surfaces has been confirmed to enhance the exopolysaccharide synthesis by a number of investigations except for increases in cell density (Vandevivere and Kirchman 1993). Previously, the utilization of calcium alginate beads to trap the Chryseomonas luteola cells and exopolysaccharides triggered the highest partition of cadmium, cobalt, and copper, as well as nickel ions from aqueous solutions (Ozdemir et al. 2005a, b). The structure of bacterial exopolysaccharides can be modified to amplify their binding functions. This can be accomplished by the chemically integrating phosphoryl, acetyl, sulphonyl, methyl, and carboxymethyl functional groups (Gupta and Diwan 2017). Although these modified polymers applied rarely in metal removal studies, phosphorylated bacterial cellulose from Acetobacter species was previously shown to augment the precipitation of different metals, compared to the innate one (Oshima et al. 2008). The significant concentrations of metals can be precipitated from the polluted water by the biofilms. Biofilms can precipitate fivefold higher cadmium from the stream water with respect to stagnant water. Interestingly, the photosynthesis ability of massive biofilms was not affected even from the highest cadmium concentrations, whereas the metals were found to induce toxic effects on photosynthesis in slender biofilms (Hill et al. 2000). Recently, the aqueous solutions of cadmium, lead, copper, and zinc were precipitated through biosorption using biofilm-forming bacteria isolated from mining sites in the Philippines. According to the investigators, these biofilm-producing isolates hold the potential to be further prospered for the economical bioremediation of actual metal contaminated wastewaters (Villegas et al. 2018). In another study, the 11 biofilm-forming bacterial isolates were isolated from the metal-contaminated wastewaters of different industrial facilities in Bangladesh. Increased production of proteinaceous curli fimbriae and main exopolysaccharide component, cellulose, and accordingly enhanced production of biomass biofilm were quantified in all these isolates. Additional application of metal stress at certain concentrations further resulted in noticeable biofilm production (Mosharaf et al. 2018). Various mercury-resistant bacteria were isolated from different polluted ecosystems (Poulain et al. 2007; Mirzaei et al. 2008; Deng and Wang 2012). Recently, the isolation of biofilm-producing marine bacterium Bacillus thuringiensis carrying plasmid for mercury resistance operon was reported. Since the isolate was shown to reduce over 90% of the residual mercury under different pH, salinity, and temperature, it promises the great usage in the decontamination of mercurycontaminated aqueous environments (Dash et al. 2014). The bacterial surface substances have also been implicated in the precipitation of cadmium, lead, and zinc from the environment (Maier and Soberón-Chávez 2000; Mulligan 2005). It was also reported that a class of such biosurfactant, rhamnolipids, make powerful complexes with assorted metals (Ochoa-Loza et al. 2001; Neilson et al. 2003). A recent study demonstrated a firm attachment of these Pseudomonas aeruginosa-derived “green materials” to a variety of environmentally hazardous metals including uranium, lead, copper, cadmium, aluminum, and others. Moreover, the bounding affinities of rhamnolipids to heavy metals were much higher than the common soil and water cations, indicating their potential for the bioremediation of

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heavy metal contaminants (Hogan et al. 2017). This stable interaction can be associated with the activation of quorum-sensing mechanisms since the cadmium exposure was found to be positively correlated with rhamnolipid synthesis with the induction of acyl-homoserine lactone synthase gene in Pseudomonas aeruginosa (Neilson et al. 2010; Chellaiah 2018).

5.6

Conclusion and Future Prospects

Being the frontier study area of microbiology, the potential of bacterial biofilms in the remediation of metal contaminated aqueous environments is promising and necessitates the conduction of comprehensive work especially focusing the molecular mechanisms that are a complex network of regulatory pathways and crosstalk standing behind this phenomenon. Speed rising in omics technologies will eventually lead to the discovery of novel and unknown metal-resistant genes, their transcripts, functional products, as well as intact organisms. With the help of recombinant DNA technologies, these items can be tailored to create a live system targeting the decomposition and sequestration of environmentally hazardous metals. The utilization of biofilms produced from the recombinant or indigenous metaltolerant and resistant bacterial communities will open new horizons of limitless possibilities in the design of smart bioremediation practices with low cost and rapid effect. Indeed, the biochemical engineering of biomolecules constituting the biofilm will also be acknowledged to bring the new perspectives in the decontamination strategies of metal-polluted areas. Consequently, the technologies developed at a laboratory or microcosm scale can be transformed into the actual in-field decontamination strategies.

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Chapter 6

Laccase-Mediated Bioremediation of Dye-Based Hazardous Pollutants Muhammad Bilal

, Syed Salman Ashraf

, and Hafiz M. N. Iqbal

Contents 6.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2 Laccases and Their Physicochemical Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3 Soluble/Purified Laccases for Dyes Degradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4 Immobilized Laccases and Biocatalytic Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5 Carrier-Supported Immobilized Laccases for Dye Remediation . . . . . . . . . . . . . . . . . . . . . . . . . 6.6 Carrier-Free Immobilized Laccases for Dye Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.7 Dye Degradation Mechanism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.8 Conclusions and Prospects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Increasing urbanization and uncontrolled development of textile amenities together with the dearth of efficient removal and/or treatment strategies led to discharge wastewater effluents loaded with various toxic, recalcitrant, and oncogenic dyes. Mostly, such hazardous pollutants are being excluded into main water streams either with or without partial and/or incomplete treatments. Owing to the technoeconomic restrictions, several of in practice chemical/physical routes are either inefficient or deterring. Enzyme-assisted bioremediation is an easy, eco-sustainable, and widely pursued approach to biodegrade and biotransform a range of organic contaminants and recalcitrant xenobiotic compounds. Laccases have demonstrated great bioremediation potential owing to their excellent catalytic efficiency and high

M. Bilal (*) School of Life Science and Food Engineering, Huaiyin Institute of Technology, Huaian, China S. S. Ashraf Department of Chemistry, College of Arts and Sciences, Khalifa University, Abu Dhabi, United Arab Emirates e-mail: [email protected] H. M. N. Iqbal (*) Tecnologico de Monterrey, School of Engineering and Sciences, Campus Monterrey, Monterrey, Mexico e-mail: hafi[email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 137 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_6

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chemo-, regio-, and enantio-selectivity under slight environmental settings. Enzyme immobilization increases the half-life, catalytic efficiency, and thermal steadiness of the enzymes. Several techniques extending from simple reversible adsorption and ionic bonding to the irreversible covalent attachment have been employed to immobilize laccase onto various support materials. These engineering strategies create immobilized biocatalyst of variable catalytic stability and reusability attributes by manipulating the surface microenvironment of the carrier support. In this chapter, an effort has been made to address laccase immobilization (carrier-supported as well as carrier-free immobilization) as a futuristic eco-friendlier and practical choice for the abetment of dye-based effluent wastewater by appraising its physicochemical properties and catalytic performance. The concluding remarks, current challenges, and future scenarios for the remediation of dyes are also discussed. Keywords Green biotechnology · Industrial waste · Laccase · Immobilization · Carrier-free immobilization · Immobilized enzymes · Enzymatic remediation · Textile wastewater · Dye degradation

Abbreviations ABTS CLEAs COD GLU GO HBT LMS MB MG MGO RBBR WRF

6.1

2,20 -Azino-bis(3-ethylbenzothiazoline-6-sulphonic acid) Cross-linked enzyme aggregates Chemical oxygen demand Glutaraldehyde Graphene oxide 1-Hydroxy benzotriazole Laccase-mediator system Methylene blue Malachite green Magnetic graphene oxide Remazol Brilliant Blue R White-rot fungi

Introduction

Sustainable water quality management has recently become a subject of immediate research due to the exponential increase in water contamination problems. As part of these attempts, the abatement and elimination of toxic environmental chemicals, such as dye pollutants, have been acknowledged as prime tasks by the scientific community, environmental engineers, and worldwide agencies. Figure 6.1 represents various sources of environmental pollutants and their adverse effects that can be remediated by the use of native or insolubilized laccase (Bilal et al. 2019a). Consequently, extensive progress has undeniably been made to expound the adverse

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Fig. 6.1 Various sources of environmentally related contaminants that can be tackled with free and/or immobilized laccase. General distribution and adverse effects are given based on contaminants. (Reprinted from Bilal et al. (2019a) with permission from Elsevier. Copyright (2019) Elsevier Ltd.)

consequences of dye pollutants on the ecosystem and health of humans (Vikrant et al. 2018). Dye pollution appears as one of the all-time high threat because of widespread consumption as colorants in multiple industries like paper, tannery, textile, food, cosmetics, and pharmaceutics. During the dyeing operations, approximately, 15% of dyes remain unfixed and are lost in the wastewater effluents (da Silva Leite et al. 2016). The wastewater harboring such type of unfixed dye molecules substantially raises the ecological pollution creating serious health menaces when expelled in the ecosystem (Mishra and Maiti 2019). These dyes along with intermediate degradation products are documented to be highly poisonous with carcinogenic and mutagenic properties. Due to high coloration and chemical oxygen demand, dyes can reduce water transparency and oxygen solubility resulting in potential inequity in the development of aquatic biota (Bilal et al. 2018a, 2019b). Attempts have been made to develop innovative and promising technologies for dyes removal to overcome this issue. Various chemical, physical, and advanced oxidation approaches have been used to remediate recalcitrant dyestuffs from industrial wastewater. Nevertheless, the degradation of dyes using these methods can generate highly toxic and detrimental aromatic amines. In this regard, bio-based methods have gained researchers interest owing to their greater dye degradation efficacy, eco-sustainability, and cost-effectiveness. The utilization of whole-cell bacteria, algae, and fungi have been examined, but the production of secondary sludge, longer evolution cycle, and susceptibility to shock loads hamper the applicability of system (Zamora et al. 2003). The abovementioned inadequacies can be addressed by applying dye-degrading enzymes that are demonstrated to yield encouraging results in a shorter time without any ecological impact. Biocatalysts or enzymes produced by various microbial strains can transform complex toxic

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Fig. 6.2 A schematic pathway of enzyme screening to immobilization with novel features to scale up to an industrial level. (Reprinted from Bilal et al. (2018a) with permission from Elsevier. Copyright (2018) Elsevier B.V.)

chemical compounds into simpler harmless species in a broader pH and temperature ranges and organic solvents. Novel features of enzymes from the industrial implementation viewpoint are shown in Fig. 6.2 (Bilal et al. 2018a). In recent years, enzymes have gained a great deal of research attention in various fields of green chemistry including biofuel production, biosensors development, and decolorization of textile-related dyes (Fig. 6.3) (Bilal et al. 2017; Asgher et al. 2014, 2016). A range of enzyme biocatalysts, i.e., peroxidases, mono-/dioxygenases, and laccases from various microorganisms and plants, have shown potential for biodegradation and biotransformation of diverse organic contaminants including textile dyes. Among these enzymes, the use of laccases has recently drawn enormous attention as an attractive choice for the effective decolorization of waste effluents by enzymatic oxidation of the dye. Their high cost and low stability hamper the application of enzymes in industrial bioprocesses. However, the biosynthesis cost for microbial biocatalysts can be diminished in contemporary genetic engineering approaches, whereas the use of carrier-insolubilized enzymes can improve their catalytic efficacy and lifespan (Bilal et al. 2017). Recent literature reports have demonstrated greater dye removal efficiencies by the immobilized enzymes than that of their free counterparts. Therefore, the immobilized form of enzymes can serve as superior environmental candidates for bioremediation of a range of toxic dyes. Immobilized enzymes display some desirable features such as convenient handling, facile separation, greater catalytic activity, stability and repeatability, specificity, selectivity, and improved tolerance to inhibitions (Zdarta et al. 2018; Bilal et al. 2019a, b). This study explicitly investigates the capability of the free and insolubilized laccases to remove a variety of recalcitrant and toxic dyes from

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Fig. 6.3 Application of laccase in various fields of biotechnology

contaminated wastewater. The report would be beneficial to the investigators engaged in the arena of bioremediation of wastewater containing various dyes.

6.2

Laccases and Their Physicochemical Characteristics

Laccases (EC.1.10.3.2) are multi-copper oxidases that are extensively disseminated in eukaryotes and prokaryotes. Nevertheless, laccases from microbial origins, in particular, from white-rot fungi, have attracted increasing attention owing to great oxidation ability to multiple compounds and a wider spectrum of substrate specificity (Gasser et al. 2014). Based on copper centers, these enzymes are categorized into groups including type 1 (blue), type 2 (normal), and type 3 or coupled binuclear. Type 1 and type 2 exhibits one Cu atom each, whereas type 3 possesses two Cu atoms (Ouzounis and Sander 1991). The biocatalytic efficacy of laccases relied on the oxidation-reduction potential of the type 1 copper ion, where oxidation of substrate occurs. The higher redox

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potential of the microbial laccases than plant-based enzymes indicates their higher catalytic efficiency and activity relative to plant-sourced enzymes. On the other hand, type 2 and type 3 form a tri-nuclear cluster (T2/T3) for molecular oxygen reduction to water by transferring electrons from T1 to the tri-nuclear site (Falade et al. 2018). Usually, the biocatalytic reaction of laccases implicates four substrate oxidations together with oxygen reduction to two molecules of water. Consumption of atmospheric oxygen as a source of electron acceptor is beneficial in laccase-mediated catalytic reaction than the use of hydrogen peroxide by peroxidases. Nonetheless, these enzymes necessitate the use of redox mediators, i.e., acetosyringone, 2,20 -azino-bis(3-ethylbenzothiazoline-6-sulphonic acid), 1-hydroxy benzotriazole, and vanillin for abatement of numerous recalcitrant and of non-phenolic environmental contaminants. In spite of this shortcoming, the use of laccases has gained noteworthy interest due to their tremendous biotechnological applications. Remarkably, there has been a paradigm revolution from the laccases application in conventional lignin depolymerization to remove emergent contaminants and organic micropollutants (Falade et al. 2018).

6.3

Soluble/Purified Laccases for Dyes Degradation

Laccases are lignin-degrading enzymes with a pronounced capacity to decolorize and remove environmental pollutants in wastewater effluent. In addition to lignin modification, they have revealed promising potentials for the biotransformation and elimination of numerous kinds of recalcitrant aromatic compounds with toxic properties such as alkylphenols, polychlorinated biphenyls, polycyclic aromatic hydrocarbons, and pesticides. The catalytic efficiency and activity of laccases have recently been well-inspected for the elimination of dye pollutants in wastewater. Table 6.1 represents a summary of recent studies for various dyes removal by the free form of laccases. A novel laccase isoform LacA purified from the culture extract of WRF Cerrena unicolor BBP6 showed excellent decolorization abilities for various recalcitrant dyes. In the presence of 1-hydroxy benzotriazole, the newly isolated LacA caused 96.6%, 98.9%, 97.1%, and 84.6% decolorization of azure blue, methylene blue, safranin, and simulated textile effluent, respectively. It also presented a considerable potential in denim bleaching, when employed in combination with manganese peroxidase, Mn2+, and HBT (Zhang et al. 2018). Navada and coworkers (2018), for the first time, isolated laccase enzyme Phomopsis sp. and employed for biotransformation and detoxification of a highly recalcitrant dye Remazol Brilliant Blue R. The partially purified enzyme transformed more than 85% of dye molecules into harmless species in 5 min without the use of any mediators. It retained catalytic efficiency in a broad pH and temperature range of 4–8 and 25–45  C, respectively. Spectral analysis, FT-IR and LC-MS data revealed the dye degradation and proposed mechanism by laccase following oxidation, hydroxylation, deamination, and ring cleavage (Fig. 6.4) (Navada et al. 2018). After enzymatic treatment, the chemical oxygen demand of the tested dye solution

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Table 6.1 Decolorization studies of different kinds of dyes by free laccases from various microbial sources Decolorization (%) 75.01%

Source of enzyme Peroneutypa scoparia

Name of dye Acid Red 97

P. scoparia

Acid Brown 282

69.15%

6h

P. scoparia

Acid Black 210

65.12%

6h

P. scoparia

Acid Green 16

66.17%

6h

P. scoparia

Acid Yellow 42

62.73%

6h

P. scoparia

Acid Blue 7

69.17%

6h

Myrothecium roridum

Amaranth

Bacillus pumilus

Evans blue

80.35%

24 h

B. pumilus

Reactive Brilliant Orange K-7R Evans Blue and Reactive Black Remazol Brilliant Blue R Malachite green

71.44%

24 h

64.72%

24 h

B. pumilus Phomopsis sp. Pleurotus pulmonarius BPSM10 P. pulmonarius BPSM10

90%

85%

Time 6h

24 h

5 min

68.61%

1h

Naphthol blue black

55.26%

12 h

P. pulmonarius BPSM10

Orange G

58.58%

36 h

P. pulmonarius BPSM10

Victoria Blue B

16.03%

24 h

P. pulmonarius BPSM10

Phenol red

22.22%

36 h

P. pulmonarius BPSM10

Congo red

61.14%

24 h

P. pulmonarius BPSM10

33.12%

36 h

Cerrena unicolor

Coomassie Brilliant Blue Bromothymol blue

84.9%

12 h

C. unicolor

Evans blue

71.8%

12 h

C. unicolor

Methyl orange

68.4%

24 h

C. unicolor

Malachite green

86.2%

24 h

References Pandi et al. (2019) Pandi et al. (2019) Pandi et al. (2019) Pandi et al. (2019) Pandi et al. (2019) Pandi et al. (2019) Jasińska et al. (2019) Xia et al. (2019) Xia et al. (2019) Xia et al. (2019) Navada et al. (2018) Leo et al. (2018) Leo et al. (2018) Leo et al. (2018) Leo et al. (2018) Leo et al. (2018) Leo et al. (2018) Leo et al. (2018) Wang et al. (2017) Wang et al. (2017) Wang et al. (2017) Wang et al. (2017) (continued)

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Table 6.1 (continued) Decolorization (%) 75.4%

Source of enzyme Leucoagaricus naucinus LAC-04 L. naucinus LAC-04

Name of dye Bromothymol blue

Time 10 h

Eriochrome Black T

64.7%

10 h

L. naucinus LAC-04

Evans blue

53.2%

10 h

L. naucinus LAC-04

38.3

10 h

L. naucinus LAC-04

Remazol Brilliant Blue R Fuchsin Basic

37.6%

10 h

Trametes sp. LAC-01

Bromothymol blue

65.4%

24 h

Trametes sp. LAC-01

Malachite green

75.8%

24 h

Pichia pastoris

Malachite green

94.1%

20 h

P. pastoris

Bromophenol blue

96.2%

20 h

P. pastoris

Methyl orange

80.2%

20 h

P. pastoris

Crystal violet

62.0%

20 h

References Ning et al. (2016) Ning et al. (2016) Ning et al. (2016) Ning et al. (2016) Ning et al. (2016) Ling et al. (2015) Ling et al. (2015) Zhuo et al. (2015) Ling et al. (2015) Ling et al. (2015) Ling et al. (2015)

P. scoparia, Peroneutypa scoparia; B. pumilus, Bacillus pumilus; P. pulmonarius BPSM10, Pleurotus pulmonarius BPSM10; C. unicolor, Cerrena unicolor; L. naucinus LAC-04, Leucoagaricus naucinus LAC-04; P. pastoris, Pichia pastoris

was significantly diminished (60%), and the resulting transformed dye products were found to be harmless on the growth of plants and microorganisms. Of most recent, a green and sustainable approach was adopted for the removal of eight different leather dyes by the laccase enzyme extracted from Peroneutypa scoparia. Maximum color removal was recorded in Acid Red 97 (75%), followed by Acid Black 210 (69%), Acid Violet 54 (69%), Acid Green 16 (66%), Acid Brown 282 (65%), and Acid Yellow 42 (62%). However, Acid Blue 193 and Acid Blue 7 were found to be resistant with a lower removal percentage of 44 and 59%. LC/MS and 1H NMR analyses proposed the oxidative degradation of a phenolic group of AR 97 dye to naphthalene 1,2-dione and 3-(2-hydroxy-1-naphthylazo) benzenesulfonic acid by the action of laccase (Fig. 6.5) (Pandi et al. 2019). Trovaslet and colleagues (2007) achieved 52 and 80% decolorization of Acid Yellow 36 and Acid Blue 62, respectively, by incubating the enzyme with dyes for 24 h. Up to 90% of amaranth dye was decolorized by laccase from a non-ligninolytic fungus Myrothecium roridum after 24-h redox mediator-assisted incubation. It also entirely degraded the simulated effluent composed of many dyes, reducing agents, detergents, and metal ions (Jasińska et al. 2019). An engineered laccase rLAC-EN3-1 obtained from P. pastoris presented a promising ability for biodegradation of malachite green.

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Fig. 6.4 Proposed hypothesis for degradation of Remazol Brilliant Blue R dye by a laccase enzyme electron beam irradiated (0.2 kGy) endophytic fungus Phomopsis sp. (Reprinted from Navada et al. (2018) with permission from Elsevier. Copyright (2018) Elsevier Ltd.)

Moreover, the enzyme-treated dye solution showed considerably lower toxicity compared to the original dye sample (Zhuo et al. 2015)

6.4

Immobilized Laccases and Biocatalytic Properties

Enzyme immobilization is an attachment of native or soluble form of the enzyme onto a support material to hold in suitable reactor geometry for improved catalytic efficiency, stability, and economical recycling of enzymes even under unfavorable reaction conditions (Bilal et al. 2018a, b). Several immobilization procedures such as adsorption on the glass, alginate matrix or microspheres, encapsulation, entrapment, covalent coupling, and cross-linking by the use of bi-functional reagents have been

146 Fig. 6.5 Degradation pathway of Acid Red 97 by laccase from Peroneutypa scoparia. (Reprinted from Pandi et al. (2019) with permission from Elsevier. Copyright (2018) Elsevier Ltd.)

M. Bilal et al.

OH -O S 3 N

OH N

N Acid Red 97

N

SO3Cu (II)

-e , H +

O

Cu (I)

-O S 3 N

O N

N

N

SO3Cu (II)

-e , H +

O

Cu (I)

-O S 3 N

O N

N

N

SO3– +OH (H2O)

O H O

-O S 3

N

O N

N

N

SO3-

O H O

-O S 3 N=NH +

O

3-(2-hydroxy-1-naphthylazo) benzenesulfonoic acid Naphthalene 1,2-dione

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147

Fig. 6.6 A schematic illustration of cross-linked enzyme aggregates and combined cross-linked enzyme aggregates development in the presence of cross-linker and Fe2O3 particles, respectively. (Reprinted from Bilal et al. (2018b) with permission from Elsevier. Copyright (2017) Elsevier B.V.)

developed and reported to attach a broad variety of industrially relevant enzymes. Notably, the carrier supports selected for enzyme attachment should exhibit greater surface area, less expensive, and circumventing diffusional limitations of substrate and product for enzyme catalytic reactions. Immobilized laccases have presented their initial activities for a longer time than their pristine counterparts and could easily separate and recycle for several successive cycles (Bilal et al. 2017). Laccases in immobilized form may exhibit economic advantages to degrade and detoxify numerous dye pollutants and xenobiotic compounds because of their catalytic steadiness and repeatability. They also demonstrate improved activity in broadspectrum pH and temperature conditions. The thermal stability of intracellular enzymes that are not operative in a cell-free system can also be increased by immobilization technology. Figure 6.6 illustrates a diagrammatic representation of cross-linked enzyme aggregates and combined cross-linked enzyme aggregates (Bilal et al. 2018b).

6.5

Carrier-Supported Immobilized Laccases for Dye Remediation

Intensely polluted wastewater can markedly reduce the catalytic performance, selectivity, and stability of the soluble enzymes. Therefore, enzymes in immobilized states are well suited and gained popularity to circumvent these drawbacks based on their improved stability and durable shelf life features. In recent years, a wide variety of supporting matrices in the form of beads, gels, or membranes have been pursued for enzymes immobilization to achieve desirable

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attributes such as rapid enzyme separation, enhanced stability, durability, and resistance to extreme pH, temperatures, and elevated substrate concentrations. Decolorization studies of different kinds of dyes by the immobilized laccase enzymes are summarized in Table 6.2. Trichoderma harzianum sourced an extracellular laccase was entrapped in copper alginate, calcium alginate, sol-gel matrix, and calcium alginate-chitosan microspheres. Optimally entrapped enzyme in sol-gel showed maximum activity in wider pH range from 4.0 to 7.0 and found to be thermo-tolerant (50–70  C) compared to its soluble counterpart. Immobilized laccase was used for biodegradation of structurally diverse synthetic dyes, namely, Congo red, methylene blue, and malachite green (200 mg/L). Using HBT as a mediator, it caused 60%, 90%, and 100% decolorization of Congo red, methylene blue, and malachite green after an incubation time of 20, 18, and 18 h, respectively. In addition, the phytotoxicity test revealed the nontoxicity of the by-products of degraded dyes against Phaseolus mungo with regard to native dye samples (Bagewadi et al. 2017). Very recently, Kashefi et al. (2019a) integrated unique characteristics graphene oxide nanosheets with superparamagnetic properties of the CuFe2O4 nanoparticles to fabricate MGO. The resultant magnetic graphene oxide was functionalized and cross-linked with 3-aminopropyl trimethoxy silane and glutaraldehyde, respectively, and employed for covalent attachment of laccase. The immobilized nanobiocatalytic system led to the 95.33% color removal of Direct Red 23 dye under optimal conditions of dye concentration (19.60 mg/L), biocatalyst amount (290.23 mg/L), and reaction pH (4.23). Particularly, the magnetic features of nanobiocatalytic support allowed its rapid and facile separation from the complex reaction media. In contrast to native enzyme, the newly synthesized immobilization support ensures enhanced laccase availability and its repeated uses. Similarly, laccases loaded on poly(hydroxyethyl methacrylate-co-vinylene carbonate) microbeads results in 100% elimination of Cibacron Blue 3GA dye within 120 min. After 1 h of incubation with insolubilized laccase preparation, dye and its by-products completely vanished from the reaction mixture. The dye degradation rate was higher in the presence of a mediator than those without the addition of a mediator compound. The toxicity profile examined using Daphnia magna and a green micro-algal (Chlorella vulgaris) growth inhibition confirmed the nontoxic nature of samples after treatment with the enzyme. However, initial intermediates appeared to be highly toxic than the dye products formed later (Bayramoglu et al. 2019). Carbaryl pesticide and MB dye were completely removed from the aqueous solution by the laccases conjugated on acrylate-based microbeads using acetosyringone as a mediator. The immobilized enzyme lost only 8.0 and 2.0% of its original activity for removal of MB dye and carbaryl pesticide, respectively, when operated continuously in a fluidized bed reactor for 24 h (Fig. 6.7) (Bayramoglu and Arica 2019). Malachite green dye-based effluent accompanied by cadmium was efficiently decolorized (over 98%) by Trametes versicolor laccase physically adsorbed onto kaolinite using 3, 5-dimethoxy-4-hydroxybenzaldehyde. It preserved higher than 50% of its original catalytic performance and 80% dye color removal efficiency after five consecutive batch cycles (Wen et al. 2019).

Brevibacterium halotolerans N11 Aspergillus oryzae

Trametes versicolor

Magnetic graphene oxide

Genetically modified Aspergillus Trametes versicolor

Ca-alginate, agarose-agar, agar-agar, alginategelatin mixed gel and polyacrylamide gel Layered double hydroxide/alginate biohybrid beads

Graphene oxide nanosheets separation membrane Graphene oxide nanosheets separation membrane

Cellular loofa sponge carrier

Hybrid inorganic Nanoparticles/polymer-biomacromolecules vesicles Acrylate-based microbeads

Trametes hirsute

Trametes versicolor

Graphene oxide nanosheets

Genetically modified Aspergillus Genetically modified Aspergillus –

Graphene oxide nanosheets

Enzyme support Kaolinite

Source of enzyme Trametes versicolor

Cross-linking

Entrapment

Adsorption

Covalent attachment Covalent attachment Adsorption

Covalent attachment

Type of immobilization Physical adsorption Covalent attachment Covalent attachment Encapsulation

Malachite green

Reactive Brilliant Blue KNR Reactive Brilliant Blue K-GR Congo red

Remazol Brilliant Blue R Direct Red 23

Methylene blue

Congo red

Acid Blue 92

Name of dye Malachite green Direct Red 23

90%

85%

More than 40%

More than 40%

17.76 mg treated dye/g of loofa 95.33%

100%

99.5%

48.7%

88.7%

Degradation (%) 98%

Table 6.2 Decolorization studies of different kinds of dyes by immobilized laccases from various microbial sources

1.5 h

24 h

2h

2h

–

5h

2h

3h

1h

1h

Time duration 5h

(continued)

Reda et al. (2018) Huang et al. (2018)

Bayramoglu and Arica (2019 Mohammed et al. (2018) Kashefi et al. (2019b) Xu et al. (2018) Xu et al. (2018)

References Wen et al. (2019) Kashefi et al. (2019b) Kashefi et al. (2019b) Wu et al. (2019)

6 Laccase-Mediated Bioremediation of Dye-Based Hazardous Pollutants 149

Source of enzyme Trichoderma harzianum strain HZN10 Trichoderma harzianum strain HZN10 Trichoderma harzianum strain HZN10

Table 6.2 (continued) Type of immobilization Entrapment

Entrapment

Entrapment

Enzyme support Sol-gel matrix

Sol-gel matrix

Sol-gel matrix

Congo red

Methylene blue

Name of dye Malachite green

60%

90%

Degradation (%) 100%

20 h

18 h

Time duration 16 h

Bagewadi et al. (2017)

Bagewadi et al. (2017)

References Bagewadi et al. (2017)

150 M. Bilal et al.

Fig. 6.7 Schematic representation of bioremediation of pollutants in the reactor system. (Reprinted from Bayramoglu and Arica (2019) with permission from Taylor & Francis. Copyright (2019) Taylor & Francis Group, LLC)

6 Laccase-Mediated Bioremediation of Dye-Based Hazardous Pollutants 151

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Catalytic treatment of dye wastewater by gold nanoparticles or laccaseincorporated hybrid vesicles biocatalysts has received exceptional research consideration. As a part of this effort, Wu et al. (2019) designed a stabilized building block with inorganic gold nanoparticles/nanoconjugates and characterized by various techniques. These assembled hybrid giant vesicles served as a novel support for laccase encapsulation and employed to degrade CR dye. Remarkably, the nanobiocatalytic system presented high catalytic decolorization performance achieving over 98% of color removal coupled with tremendous reusability. A possible dye degradation mechanism by laccase ⊂ AuNPs@vesicle catalyst was proposed and shown in Fig. 6.8. CR dye molecules enter into the cavities of the hybrid giant vesicles and form the electrondeficient reaction center by the catalytic action of laccase in these vesicles. This carbocation ion is extremely reactive intermediate resulting in asymmetric fragmentation of azobenzene bonds to form diazine and p-dihydroxy biphenyl intermediate. Through the subsequent oxidative steps, Congo red is completely broken down into nontoxic hydroxyl and hydroperoxyl metabolites. Though the individual immobilization of the laccase or mediator has been inspected. Nevertheless, only scarce reports have explored the co-immobilization of laccase and redox mediator, and investigations reporting the reusing options of a laccase-mediator system are also still lacking. Some researchers attempted to immobilize redox mediators with an objective recover and reuse these for repeated dye decolorization cycles. For example, Huang et al. (2018) develop a novel laccasemediator biocatalytic system by encapsulating both laccase and ABTS on layered double hydroxide/alginate biohybrid beads (Fig. 6.9) and evaluated its potential for decolorizing malachite green dye. Results revealed that the free and immobilized laccase-mediator systems achieved 90% and 92% decolorization yields of malachite green within 1.5 and 2 h, respectively. MALDI-TOF-MS analyzed and identified the intermediate metabolites generated during the dye degradation process and resulting degradation products speculated an N-demethylation-dependent oxidative degradation of malachite green dye. The N-demethylation results in sequential intermediates, including desmethyl, didesmethyl, and tridesmethyl MG, which reduced the characteristic absorption peaks. Subsequently, these intermediates were further fragmented into smaller molecules with one benzene ring (Fig. 6.10) (Huang et al. 2018). In addition, the malachite green dye solution was completely detoxified following treatment with the insolubilized enzyme. It retained 79% of an original decolorization activity after eight successive cycles for dye degradation and found to be completely stable after storing for 10 days. Mendoza et al. (2011) corroborated that 2,2,6,6-tetramethylpiperidine 1-oxyl mediator combined with a poly(ethylene glycol) molecule could be recycled for laccase-mediated degradation of azo dye. Similarly, ABTS-modified silica nanoparticles have also been successfully employed for dye decolorization by laccase (Liu et al. 2015). Moreover, ABTS coupled with MOF MIL-100(Fe) was found to be a proficient redox mediator for laccase-assisted degradation and detoxification purposes (Liu et al. 2017).

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Fig. 6.8 The self-assembly process of laccase ⊂ AuNPs@vesicles and proposed catalytic mechanism for biodegradation of Congo red by laccase ⊂ AuNPs@vesicles. Reprinted from Wu et al. (2019) with permission from Elsevier. Copyright (2018) Elsevier B.V

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Absorbance (A.U)

1.5 Blodegration

1.0 MG

0.5 H3C

CH3

CH3

N

N

CH3

0.0 200

300

400 500 600 Wave length (nm)

0 min 60 min 80 min

700

800 329.20179 m/z

100 min 120 min H3C

(a)

CH3

CH3

N

N

315.18578 m/z

10 329.20204

Intens´106

8 6 4 2

125 150 175 200 225 250 275 M/Z (b)

300 325 350

Intens´106

2.0

1.0 0.5

CH3

N

N

N

H

301.17001 m/z

H

H

N

N

H H

O

158.98768 315.18596

274.27420 301.17033 212.03162 245.12850 204.06520

0.0 225 250 M/Z (c)

N

301.17001 m/z

H

125 150 175 200

H

CH3 H3C

H or

0

1.5

CH3 H

172.03890 190.04970

CH3

275 300 325 350

N

274.27449 m/z

CH3 H

212.03168 H m/z Further degradation

(d)

Fig. 6.9 (a) UV-Vis spectra of the malachite green dye during the enzymatic degradation process at different time intervals. Positive-ion matrix-assisted laser desorption/ionization mass spectrum (b) before enzymatic degradation. (c) Two hours after enzymatic degradation. (d) Proposed pathway of malachite green degradation after treatment with the immobilized enzyme. (Reprinted from Huang et al. (2018), an open-access article distributed under the Creative Commons Attribution License)

6.6

Carrier-Free Immobilized Laccases for Dye Remediation

In recent years, carrier-free immobilization technologies are gaining an accelerating interest in green chemistry because of their low costs, practical simplicity, and eco-friendliness. Laccases CLEAs from Agaricus bisporus, C. polyzona, S. putrefaciens, T. villosa, and T. versicolor for dye-containing wastewater treatment have been described in the literature (Table 6.3). All these developed CLEAs-laccases have revealed improved thermal, storage, and functional stabilities

6 Laccase-Mediated Bioremediation of Dye-Based Hazardous Pollutants

+ ABTS

2+

3+

Lac/ZnCr-ABTSLDH

ZnCr-ABTSLDH

Zn , Cr

Encapsulation

Adsorption

Co-precipitation

155

Alginate beads (Im-LMS)

(a) O2

Lac (Red)

H2O

Lac (Ox)

+

ABTS

Dye (Red)

Recyle ABTS

Dye (Ox)

Use Alginate Im-LMS

Laccase

ABTS Dye (Red) Dye (Ox)

Biodegradation (b)

Fig. 6.10 Schematic illustration of the strategy for preparation of Lac/ZnCr-ABTS layered double hydroxides/alginate beads and malachite green biodegradation. (Reprinted from Huang et al. (2018), an open-access article distributed under the Creative Commons Attribution License)

against deactivation by heat, organic solvents, or autoproteolysis. Additionally, they possess higher catalytic efficiencies and are readily recoverable and recyclable. For instance, laccase produced by Cerrena sp. strain HYB07 was immobilized by forming CLEAs. A high activity recovery (68.1%) was achieved under optimal ammonium sulfate (precipitant) and glutaraldehyde (cross-linker) concentrations for 3 h at room temperature. Insolubilized laccase showed enhanced resistance to metal ions, NaCl, and organic solvents. It also demonstrated high catalytic performance for decolorizing RBBR than that of the soluble enzyme. The resulting CLEAs-laccase recorded over 90% decolorization in 2 h using 80 mM NaCl. Increasing discrepancies in the half time of RBBR between free and CLEAs-laccase by increasing concentrations of NaCl, highlighting the elevated NaCl tolerance by immobilized laccase derivative (Yang et al. 2016). Sinirlioglu and coworkers (2013) identified a novel laccase from S. putrefaciens, expressed in E. coli, and synthesized its CLEAs for decolorization of malachite green dye. Saturated ammonium sulfate and GLU solution served as the precipitant and cross-linker, respectively. The as-prepared laccase CLEAs demonstrated superior catalytic performance and elevated thermal properties under harsh environments of thermal and chemical agents than with free laccase. In addition to excellent repeatability, they also presented about 90% decolorization of tested MG dye after 24 h. Amino-functionalized magnetic nanoparticles coupled to CLEAs-laccase rapidly decolorized 61–96% of MG, RBBR, and RB5 at their initial concentrations of 50 mg/L at pH 7.0 and 20  C. Further, more than 90% decolorization of RBBR using a continuous lab-scale perfusion basket bioreactor for 10 h accompanied by a negligible decrease in CLEAs activity indicates its suitability for wastewater treatment.

Propanol Propanol Ammonium sulfate t-butanol t-butanol t-butanol t-butanol

Trametes versicolor Trametes versicolor Shewanella putrefaciens

Pleurotus ostreatus Pleurotus ostreatus Pleurotus ostreatus Pleurotus ostreatus

Trametes versicolor

Precipitant Acetonitrile Ammonium sulfate Propanol

Laccase source Trametes versicolor Cerrena sp. strain HYB07

Glutaraldehyde Glutaraldehyde Glutaraldehyde Glutaraldehyde

Glutaraldehyde Glutaraldehyde Glutaraldehyde

Glutaraldehyde

Cross-linker Glutaraldehyde Glutaraldehyde

Acid violet Basic red Reactive violet Reactive orange

Name of dye Trypan blue Remazol Brilliant Blue Reactive Remazol Brilliant Blue Reactive Malachite green Reactive black 5 Malachite green 65% 70% 60% 61%

96% 61%h 90%

96%

Degradation (%) 100% >90%

2h 2h 6h 6h

4h 4h 24 h

2h

Time 270 min 40 min

Kumar et al. (2012) Kumar et al. (2012) Kumar et al. (2012) Kumar et al. (2012)

Kumar et al. (2014) Kumar et al. (2014) Sinirlioglu et al. (2013)

Kumar et al. (2014)

References Seow and Yang (2017) Yang et al. (2016)

Table 6.3 Decolorization studies of different kinds of dyes by cross-linked enzyme aggregates of laccases from various microbial sources

156 M. Bilal et al.

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6.7

157

Dye Degradation Mechanism

Multiple studies have been conducted to elucidate the dye degradation mechanism by enzymes. For this, researchers utilized advanced analytical techniques including HPLC, GC-MS, FTIR, and NMR to investigate and characterize the degradation pathway intermediate metabolites from enzyme extracts. Due to structural diversity and recalcitrance, enzymatic degradation of dye pollutants is very challenging. Two simple mechanisms, i.e., adsorption and enzymatic conversion, are predominantly involved for dye elimination from the aqueous media in case of insolubilized enzymes. The anticipated mechanism for oxidative destruction of azo dyes by an enzyme follows the biotransformation of aromatic amines into non-innocuous nitrogen and depolluted compounds. The oxidation capacity for azo dyes could be expanded by incorporating redox mediators, i.e., 1-HBT and ABTS (Torres et al. 2003). Shabbir and colleagues (2017) reported that the azo dye degradation by periphyton shows the cleavage of the azo linkage and conversion of sulfonic compounds into simple aliphatic hydrocarbons.

6.8

Conclusions and Prospects

Until now, a list of various physicochemical methods has been recognized for the abatement of dyes in the aqueous milieus. Nevertheless, the use of all these techniques is hindered due to elevated operational costs, the creation of toxic by-products, and sludge, and necessities of huge amounts of eco-unfriendly chemicals and energy consequences. The direct application of biocatalysts (e.g., laccase) is a versatile biotechnological tool for the effective exclusion of dyes from water that can dramatically reduce the time associated with the production steps of microbes. Laccase-mediated bioremediation approach offers several benefits such as safe, economical, and eco-friendly operation without generating any sludge. Moreover, carrier-supported or carrier-free immobilized biocatalysts are industrially lucrative options, since they are highly stable and can be reused in multiple dye removal cycles to reduce operating costs significantly. Further technological advancements in this arena require active collaborations among enzyme biotechnologists, biochemical engineers, material researchers, and genetic engineers to meet the technological inadequacies and to apprehending dye degradation bioprocesses in terms of stability, kinetics, mechanisms, and functioning abilities. These concerted efforts might lead to developing an ideal bioremediation approach to remove and control dye-based pollution. Acknowledgments Authors are grateful to their representative universities/institutes for providing literature facilities.

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Conflict of Interests Authors declare no conflict of interest in any capacity including financial and competing.

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Chapter 7

Remediation of Freshwaters Contaminated by Cyanobacteria Sana Saqrane, Brahim Oudra, and Moulay Abderrahim El Mhammedi

Contents 7.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2 Cyanobacteria and Proliferation of Blooms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.1 Identification of Cyanobacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.2 Mechanisms of Cyanobacteria Bloom Proliferation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3 Cyanobacteria Toxins (Cyanotoxins) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.1 Neurotoxins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.2 Hepatotoxins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4 Toxic Cyanobacteria and Environmental Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.1 Ecological Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.2 Socioeconomic Impact . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4.3 Heath Risks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5 Cyanotoxin Remediation Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.1 Physical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.2 Chemical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5.3 Biological Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6 Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6.1 Role of Aquatic Macrophytes in Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.6.2 Role of Macrophytes in the Bioaccumulation and Elimination of Cyanotoxins 7.6.3 Mechanisms for Detoxification of Cyanotoxins by Macrophytes . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Cyanobacteria, commonly referred to as blue–green algae, are naturally occurring microscopic organisms found in fresh, brackish, or seawater that can release cyanotoxins into freshwater systems. Cyanobacteria also continue to attract attention in part because of well-publicized incidents of animal poisoning. There S. Saqrane (*) · M. A. El Mhammedi Polydisciplinary Faculty, Laboratory of Chemistry, Modeling and Environmental Sciences, Sultan Moulay Slimane University, Khouribga, Morocco B. Oudra Faculty of Sciences Semlalia, Laboratory of Biology and Biotechnology of Microorganisms, Environmental Microbiology and Toxicology Unit, Cadi Ayyad University, Marrakech, Morocco © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 161 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_7

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have been numerous studies on the effects of cyanobacteria and their toxins on a broad range of aquatic organisms, including invertebrates and vertebrates, that have reported acute effects (e.g., reduced survival, food inhibition, paralysis), chronic effects (reduced growth and fertility), biochemical changes (phosphatase activity, proteases), and changes in behavior. Therefore, the central aim of this chapter is to discuss the bioaccumulation and biodegradation of cyanotoxins by aquatic plants and their effects with emphasis on the role of aquatic macrophytes in bioremediation, bioaccumulation, and elimination of cyanotoxins. Keywords Cyanobacteria · Neurotoxins · Hepatotoxins · Phytoremediation · Bioremediation · Freshwater

7.1

Introduction

Water is one of the most important substances on earth. It covers about 70% of our planet. The quantity of freshwater represents only 2.5–2.75% (Gleick 1993). However, the majority of freshwater aquatic ecosystems in the world suffer from ongoing pollution causing the eutrophication problem. This phenomenon is becoming more serious by the massive proliferation of primary producers, such as cyanobacteria green–blue algae. The development of this cyanobacterial efflorescence is often accompanied by the release of secondary metabolites with various biological activities (Corbel et al. 2014). Cyanotoxins can have adverse effects on animals like humans and other mammals including sheep, cattle, and horses (Carmichael 2001; Onodera et al. 1997), fish (El Ghazali et al. 2009), invertebrates (Delaney and Wilkins 1995), and vegetable resources (aquatic and terrestrial plants) (Saqrane et al. 2007, 2008; Saqrane and Oudra 2009). To combat these cyanobacteria and subsequently eliminate cyanotoxins released into aquatic ecosystems, various treatment techniques are used, such as chlorine peroxidation, ozonation, and the use of ultraviolet radiation (Zhang et al. 2016; Liu et al. 2016). However, the application of these methods is limited due to the unsustainability of the treatment and the high cost of the materials used. In addition, the persistence of these chemical compounds can lead to damage and imbalance of aquatic ecosystems, through the risks of toxicity to both aquatic organisms and humans (Guo and Xie 2011). To avoid these issues, the implementation of biological methods could be an effective alternative. Phytoremediation is one of the biological strategies that can also be used; it consists of the use of live green plants for the in situ removal of water contaminants (Sasmaz and Obek 2012; Goswami et al. 2014). The effectiveness of phytoremediation depends on the mechanism of absorption of contaminants based on these genetic, physiological, anatomical, and morphological characteristics (Rahman and Hasegawa 2011). Aquatic plants are a major part of the primary productivity of the ecosystem and provide very important ecological

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functions especially in the biological treatment and self-cleaning of contaminated water. Thus, they play a key role in the balance and functioning of freshwater ecosystems. In the eco-toxicology, several recent scientific studies have shown that some aquatic plants are capable of bioaccumulating and biodegrading cyanotoxins (Romero-Oliva et al. 2014, 2015). These data have aroused the interest of scientists in the treatment of cyanotoxin-contaminated waters by macrophytes. Thus, the effectiveness of treatment would be mainly related to the ability of the macrophyte to accumulate these toxins and to detoxify those (Romero-Oliva et al. 2015; Pflugmacher et al. 2015). In this theme, this chapter is a synthesis of previous research on the bioaccumulation and biodegradation of cyanotoxins by aquatic plants. In fact, we have set the goal of exploiting the bio-accumulative potential of these plants in phytoremediation as a treatment bioprocess that can be applied to ecosystems contaminated by cyanobacterial toxins.

7.2 7.2.1

Cyanobacteria and Proliferation of Blooms Identification of Cyanobacteria

Cyanobacteria, known as blue-green algae, are photosynthetic prokaryotes that reproduce asexually (Komarek and Anagnostides 1999). They are part of an ancient group of microorganisms. Much of their morphological diversity developed over 2 billion years ago. Cyanobacteria are Gram-negative photosynthetic bacteria (Schopf 2000), of 150 genera and 2500 species (Duy et al. 2000). Cyanobacteria have considerable morphological diversity. They can be solitary or in colonies or organized in trichomes or filaments (with sheaths) (Fay and Van Baalen 1987). The different species of cyanobacteria are generally distinguished by their morphological characteristics (Komarek et al. 1999). These microorganisms contribute to the formation of the ozone layer. They are also the origin of all plants (Purkayastha et al. 2010). Cyanobacteria share the same habitats, compete for the same resources, and contribute to the primary production of ecosystems. Due to their unique characteristics, cyanobacteria colonize most terrestrial and aquatic freshwater and saltwater ecosystems. Some species are adapted to extreme conditions, such as in ice, thermal springs, ferruginous waters, and extreme pH (Castenholz and Waterbury 1989). The only pH seems to limit the distribution of cyanobacteria, so they tend to prefer neutral or basic conditions and are less frequent at acidic pH (Sze 1986). Cyanobacteria can be toxic because they secrete a multitude of toxins. These toxins have a wide variety of effects (Beaud et al. 2009). Among 2500 species, more than 100 species of cyanobacteria have been evaluated as toxic. The majority of them are able to form algal blooms (Reynolds 1987).

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Mechanisms of Cyanobacteria Bloom Proliferation

Recently, numerous studies have been published on cyanobacterial efflorescence and their determinism (O’Neil et al. 2012). From these studies, there is a clear consensus on the diversity and complexity of factors controlling cyanobacterial blooms. These factors occur at large scales, such as the global increase in temperature (Funari et al. 2012), or at the water body scale, which includes interactions between biotic and abiotic factors. All of these factors are interrelated (Catherine et al. 2013). Increasing nutrient fluxes such as phosphorus and nitrogen leads to eutrophication. These nutrients are mainly derived from human activities, both rural and urban (Smith et al. 2006). Human activities therefore have favorable conditions for the development of cyanobacteria. Phosphorus is very often the limiting element in freshwater aquatic ecosystems and is therefore generally the main culprit for the development of cyanobacteria (Paerl and Huisman 2008). In addition, many experts point out that climate change has the potential to act as a catalyst for the proliferation of cyanobacteria and will only amplify the problem (Paerl and Huisman 2008; Dupuis and Hann 2009). The abiotic factors that favor the predominance of cyanobacteria in aquatic systems include: • The temperature of the water, the stability of the water column, and the duration of the phytoplankton growing season Toxic cyanobacteria have an optimum growth temperature of 25  C in the North African Basin (Robarts and Zohary 1987; Oudra et al. 2001): • Increasing nutrient levels in waterbodies, generated by accelerated runoff and erosion caused by precipitation The factors that allowed cyanobacteria to colonize and dominate several types of environment are: • A wide range of photosynthetic pigments, to support photosynthesis at a low light intensity • The ability of several species to migrate vertically into the water column, under proportionally unalterable conditions, to receive light during the day on the surface, and to migrate deep during the night to take advantage of an excess of nutrients • The ability of certain species to fix atmospheric nitrogen • The ability to store phosphorus • The power to go dormant when environmental conditions unfavorable to their proliferation In addition, the toxicity of cyanobacterial strains is also influenced by nutrients. Toxic strains had higher nitrogen and phosphorus requirements than non-toxic strains (Conley et al. 2009).

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Cyanobacteria Toxins (Cyanotoxins)

Some species of cyanobacteria can produce water-soluble substances (cyanotoxins) that are potentially neurotoxic and/or hepatotoxic.

7.3.1

Neurotoxins

They are alkaloids and have two main groups: toxoids and saxitoxins. As their name suggests, they have a neurotoxic activity that acts on the nerve synapses by blocking the sodium channels and subsequently stopping the respiratory muscle and killing the individual in minutes to hours (Sivonen and Jones 1999). Neurotoxins don’t present a particular problem in recreational activities or water ingestion because they are not as widely distributed in the environment as hepatotoxins and do not appear to present the same degree of risk of chronic toxicity. Among the species that produce such a type of toxin found Aphanizomenon flos-aquae, Planktothrix, and Anabaena sp.

7.3.2

Hepatotoxins

They are produced by several species belonging mainly to the genera Microcystis, Anabaena, Planktothrix, Oscillatoria, Nodularia, Nostoc, and Cylindrospermopsis. Depending on their chemical nature, hepatotoxins include nodularin, cylindrospermopsis, and especially microcystins. The latter is the first to have been identified for the first time in Microcystis aeruginosa and is considered being the most frequent and most studied (Nishiwaki-Matsushima et al. 1992). Microcystins cause hemorrhagic hepatotoxicosis that can induce general disorders or the death of sufficiently exposed animals. They have an inhibitory activity of certain phosphatase proteins which play an essential role in cellular metabolism. These toxins are considered to be tumor promoters in the liver (NishiwakiMatsushima et al. 1992). On a structural level, they are cyclic heptapeptides whose general chemical structure is a cyclo (-D-Ala-X-DMeAsp-Y-Adda-DGlu-Mdha), where X and Y are variable amino acids whose initials are used to name different variants (Rao et al. 2002). There are now more than 100 variants of microcystins in part because of the different combinations of variable amino acids X and Y. The most common and most studied variant is microcystin-LR, with leucine (L) and arginine (R) as a variable amino acid. Other variants such as microcystinRR, microcystin-YR, M microcystin-FR, and microcystin-LA are also frequently mentioned. After synthesis, the microcystins remain inside the cell, but they can be

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rapidly and massively released after cell lysis, whether due to natural senescence or the use of algaecides.

7.4 7.4.1

Toxic Cyanobacteria and Environmental Impacts Ecological Impacts

Ecological nuisances related to the production of cyanobacterial toxins could touch all the links of the ecosystem, namely, zooplankton, green algae, plants, and fish as well as humans capable of accumulating these toxins. Algal/cyanobacterial allelopathy interactions have been demonstrated by several works. The oldest was made in situ by Lefevre (1950) and Dauta and Feuillade (1995) on antibacterial allelopathy, to explain the elimination of other algae by Aphanizomenon. Allelopathic effects on other competing eukaryotic algae can become more significant because of the importance of cyanobacterial biomasses in the aquatic environment. Ikawa et al. (2001) found that cyanotoxins can induce growth and photosynthesis inhibition of green algae (Papke et al. 1997). Cyanobacteria are much less grazed by zooplankton than other microalgae (Brett and Müller-Navarra 1997). This behavior could be explained by their toxicity, the difficulty of their ingestion and digestion, or their considerably low nutritional value (Elert et al. 2003). Toxicity to zooplankton is mainly related to hepatotoxins and neurotoxins but also other biologically active compounds, such as Daphnia toxic compounds (Jungmann and Benndorf 1994). Thus, increasing the density of cyanobacteria in water harms the growth, fecundity, and survival of zooplankton populations (Ghadouani et al. 2003). Bivalves can also ingest the cyanotoxins directly in the form dissolved by water filtration containing cyanobacteria. Many species like Dreissena polymorpha seems to be able to selectively feed itself by expelling other algae, as the case of some species of Microcystis that have been preferentially ingested, while diatoms were rejected (Baker et al. 2000). Dionisio Pires and Van Donk (2002) observed that the unwanted food of some bivalves is dominated by many more green algae cells than cyanobacteria. In fish, the effect of microcystins has been studied at all stages of their development (embryo, larva, juvenile, and adult) on several parameters such as growth rate, iron regulation, and heart rate (Malbrouck and Kestemont 2006). High fish mortality was recorded following their exposure to cyanobacterial bloom and the effect of certain environmental factors such as pH and oxygen levels influenced by the proliferation of these toxic blooms. The liver and kidneys are the main organs damaged by microcystins with the same symptoms described in mammals, namely, the appearance of necrosis in the liver and the total loss of its structure subsequently affecting its functioning (Råbergh et al. 1991; Carbis et al. 1996). Like most of the aquatic organisms listed above, aquatic plants are also exposed to cyanotoxins. Macrophyte abundance and diversity have been significantly reduced in the presence of cyanobacterial blooms (Casanova et al. 1999). This has

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been explained by changes in environmental conditions related to the massive proliferation of cyanobacteria, such as the reduction of light and oxygen and the acidification of the environment (Romanowska-Duda and Tarczynska 2002). Other studies have shown that the factor more important during the appearance of cyanobacteria is the release of cyanotoxins. However, studies referring to the phytotoxicity of cyanobacteria in the aquatic environment are still in their infancy. In 1986 Kirpenko showed for the first time the inhibition of plant growth Elodea and Lemna aquatic by toxins isolated from a natural bloom. This allelopathic action has been confirmed by (Weiss et al. 2000) following the co-cultivation of the Lemna minor plant with Microcystis aeruginosa cells. Even more recent work has reported the negative effect of cyanotoxins on biology and physiology (growth, photosynthesis, toxin accumulation, oxidative stress, etc.) of different species of the family Lemnaceae (e.g., Jang et al. 2007; Saqrane et al. 2007). Microcystins are also specific inhibitors of protein phosphatases 1 and 2A that play an essential role in several physiological processes, such as photosynthetic production of carbohydrates and maintenance of chlorophyll activity. The metabolic effect of microcystins is also manifested by a change in RNase activity and by a reduction in chlorophyll content (a, and b) in aquatic plant Spirodela oligorrhiza (Romanowska-Duda and Tarczynska 2002). The toxic effect of cyanotoxins in the physiology of an aquatic plant could subsequently have repercussions negative effects on the functioning and balance of the aquatic ecosystem through action on its biodiversity.

7.4.2

Socioeconomic Impact

Water quality is a major environmental issue in the agricultural sector. It is also important for all human agricultural uses, from irrigation to watering and domestic needs. The use of contaminated water by cyanotoxins in irrigation involves negative effects on the germination and growth processes of crop plants (Gehringer et al. 2003; Saqrane et al. 2008). Thus, this agricultural practice could induce very important economic losses which are manifested by a reduction of productivity. In this sense, the toxicology research of cyanobacteria has given special interest to cultivated plants in order to highlight the socioeconomic consequences related to the use of cyanotoxin-contaminated water table. On the other hand, investigations in this field are still very recent. In 1999, Codd et al. (1999) demonstrated for the first time the persistence of Microcystis aeruginosa cells on lettuce leaves (Lactuca sativa) irrigated for 10 days with water containing toxic Microcystis cells. Inhibition of plant growth by microcystins was observed for the first time in white mustard Sinapis alba (Kos et al. 1995). Kurki-Helasmo and Meriluoto (1998) confirmed, by isotopic labeling of the toxin, the inhibition of growth of Sinapis alba by the microcystinLR. In the same plant, M-Hamvas et al. (2003) showed the inhibitory effect of MC-LR on the growth and development of roots exposed to a concentration of 3.0–30 μg/mL. In the same sector, cyanobacterial blooms are also incriminated in many cases of poisoning of livestock (cattle, sheep, pigs, etc.) related to the presence

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of cyanotoxins in drinking water. Ressom et al. (1994) cited in a bibliographic synthesis of more than 80 cases of domestic and livestock poisoning that led to significant economic losses in the agricultural sector. Cattle watering with water containing a sublethal dose of microcystins lead to physiological disturbances related to bioaccumulation of these toxins in both the liver and plasma of the animal (Orr et al. 2003). However, according to Feitz et al. (2002), no detectable microcystin dose was quantified in the milk of exposed cows.

7.4.3

Heath Risks

In addition to their ecological and physiological effects, cyanotoxins can be bioaccumulated by living organisms in different tissues depending on the abundance of cyanobacterial blooms in the natural environment (Amorim and Vasconcelos 1999; Malbrouck et al. 2003). Thus, research in ecotoxicology of cyanobacteria is currently oriented towards the study of their bioaccumulation and their bio-transfer in the food chain following the consumption of aquatic products (fish, seafood, etc.) and terrestrial (forage plants and vegetables, etc.) contaminated. In the literature, studies have shown that microcystins are most commonly accumulated by zooplankton, which is the primary consumer of cyanobacteria (Kotak et al. 1996; Thostrup and Christoffersen 1999). Williams et al. (1997) reported that the larvae of Cancer magister accumulate microcystins from foods containing cyanobacteria. Through the trophic chain, these toxins are introduced into the Atlantic salmon species which accumulates them in the liver (Anderson et al. 1993). Indeed, the larvae can be considered as a vector for the transfer of toxins to other links in the trophic chain (Anderson et al. 1993). In fish, microcystins can be accumulated in various organs, particularly the liver, intestine, kidney, muscle, gallbladder, blood, and brain (Sahin et al. 1996; Xie et al. 2004; Cazenave et al. 2006). They were also detected in bivalves and crustaceans (Lymnaea stagnalis, Helisoma trivolvis, Physagyrina) (Zurawell et al. 1999). The rate detected depends on the abundance of blooms in the fishing and shellfish harvesting area. For this reason and to avoid any health risks related to the bioaccumulation of microcystins, many authors have recommended being cautious in the consumption of fish caught in bodies of water where blooms are abundant. In addition, several studies have shown the accumulation of microcystins by some aquatic macrophytes, such as Phragmites australis and Lemna gibba (Pflugmacher et al. 2015; Saqrane et al. 2007). Recently, research on the plant/cyanobacterial interaction has confirmed that terrestrial plants can also accumulate these toxins in their tissues following the use of cyanotoxin-contaminated water in irrigation (Saqrane et al. 2008; El Khalloufi et al. 2012). The potential effects of plant microcystin-RR accumulation have been studied by Yin et al. (2005); it is differently accumulated by the leaves and roots of Vallisneria natans seedlings. Järvenpää et al. (2007) showed that microcystin-LR

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was detected at the root level at concentrations of 0.9–2.4 ng g 1 f.w. in broccoli, respectively (Brassica oleracea var., Italica) and mustard (Sinapis alba). Peuthert et al. (2007) confirmed that plants irrigated by water containing cyanobacteria can accumulate microcystins at different concentrations. This study concerned the plants most commonly consumed by humans and animals such as Pisum sativum, Cicer arietinum, Phaseolus vulgaris, and Medicago sativa (Peuthert et al. 2007). All these studies have shown potential health risks strongly related to bioaccumulation and bio-transfer of cyanotoxins in the various trophic chains. The frequency of toxic cyanobacterial blooms in aquatic ecosystems is likely to increase in the future, especially with increasing numbers of drinking water and irrigation reservoirs, aquaculture water bodies, and recreation. Thus, this new problem related to cyanobacterial toxic blooms poses a serious threat to the quality of freshwater and a health risk for users and consumers of water contaminated by cyanotoxins. To assess the situation of this risk and to prevent the ecological, socioeconomic, and health nuisances involved, the control and prevention strategy can only be effective if it is based on reliable scientific data. Therefore, an establishment of cyanotoxins monitoring program in the natural water bodies is highly recommended, especially in favorable periods of cyanobacteria bloom proliferations.

7.5

Cyanotoxin Remediation Methods

The fight against the proliferation of cyanotoxins is crucial. It reduces the frequency of exposure to cyanotoxins and thus reduces the risk of cyanotoxins on public health. There are many methods for eliminating cyanobacteria and cyanotoxins, including coagulation, flocculation, clarification, sand filtration, the use of activated carbon and ozonation (Chang et al. 2014), and the oxidation of chlorine dioxide (Zhou et al. 2014). Alternative methods, currently under development, to optimize water treatment systems or to replace conventional technologies are too inefficient (e.g., advanced oxidation processes and membrane filtration). In general, there is no miracle cure for cyanotoxins. It is therefore necessary to determine means adapted to each case. In addition, experiments have shown that some techniques still can be ineffective. According to Rodriguez et al. (2007), more work is needed to make recommendations that take into account interference factors and to improve methods for cyanotoxins removal in water. Since the revitalization of ecosystems and the control of blooms seem to be difficult to apply in the short term, it is necessary to develop efficient water purification systems.

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Physical Methods

Eutrophication of water bodies can lead to excessive growth of cyanobacteria common to many lakes and rivers around the world (Codd 1995). Eutrophic ecosystems with high productivity and large biomass, cyanobacterial blooms are most often associated with eutrophication (Brient et al. 2000) and consequently the production of cyanotoxins. • The fight against the eutrophication of water bodies should first and foremost be through preventive methods, which really tackle the problem of contamination by limiting the diffusion of phosphorus, nitrogen, organic matter, or even pollutants in the environment. • Mechanical cleaning is intended to restore the initial depth to the medium. • Pulling up and collecting aquatic plants allow a certain amount of organic matter to be removed from the water body. The supply of nutrients in the water is thus reduced. The picking can be done by hand, collecting with rakes, or cutting with specific shears.

7.5.2

Chemical Methods

Pre-oxidation with chlorine, potassium permanganate, or ozone is generally used in the surveillance of cyanobacteria proliferation. Recently, the pre-oxidation process has become less used because of the problems of disinfection by-products associated with these chemical products. A study by Zhang et al. (2016) demonstrated that chlorine is an efficient oxidant in the degradation of microcystin-LR. In addition, the application of oxidants in water with cells of cyanobacteria can lead to cell lysis and liberation of cyanotoxins. So, in this case, it is necessary to delay pre-oxidation until the removal of cyanobacterial cells by ordinary treatment. In the case where the pre-oxidation could be applied to intact cells, the process should be optimized to minimize cell lysis and toxin liberation, by testing in the laboratory before proceeding to large-scale application (Ding et al. 2010).

7.5.3

Biological Methods

• Microbial biodegradation Biodegradation has been of great interest to the detoxification of MCs in the aquatic ecosystem. Many detoxifying bacteria have been isolated (Imanishi et al. 2005). Recently, it has been shown that blue algae can both generate renewable energy biogas and lead to the biodegradation of Microcystins by significantly reducing their concentration from 1220.19 to 35.17 μg L 1 (Yuan et al. 2011).

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• Use of plants Plant control is a new technology that seeks to exploit the metabolic and bio-accumulative capacities of plants. However, the establishment of a “remediation” system for contaminated soils and waters is relatively recent (Cunningham et al. 1997; Jasrotia et al. 2015). This technique aims to exploit the adaptability of root systems with high pollutant loads and substrate anoxia or hypoxia conditions, resulting in symbiotic relationships between microorganisms and roots that promote the elimination of pollutants (Kern and Idler 1999). This technique, called phytoremediation, can be used as a complement to physicochemical and biological methods; it has several advantages: reliability, low cost, and respect for the environment (Pilon-Smits 2005).

7.6

Phytoremediation

Phytoremediation has several names: green rehabilitation, botanical restoration, sanitation, and vegetative restoration (Pilon-Smits 2005). It is a term grouping several technologies that are based on exploiting the natural ability of plants to resist and modify the speciation of metals in water, sediments, and soils in order to rehabilitate polluted sites or reduce risks related to contaminants (Prasad et al. 2006). This technique is ecologically acceptable, efficient; it has the advantage of being realized in situ at low cost and allows bio-recovery of metals (Robinson et al. 2006). The idea of using accumulator plants for contaminants was introduced in 1983. Plants used in phytoremediation need to meet requirements such as high biomass, rapid growth, and high capacity for contaminant accumulation (Putra et al. 2015). Previous research has shown that macrophytes can cure water containing excess nutrients, organic pollutants, and heavy metals. Thus, macrophytes are being used more and more for wastewater treatment. The accumulation of metals in aquatic plants is commonly associated with the induction of cellular changes, some of which contribute directly to the plants’ ability to tolerate contaminants (Sasmaz et al. 2015).

7.6.1

Role of Aquatic Macrophytes in Bioremediation

Macrophytes are defined as aquatic photosynthetic plants whose entire life cycle, including reproduction that takes place in water. Macrophytes include phanerogams and macroalgae (Lauret et al. 2011). They play an important role in maintaining the ecological functions of shallow aquatic ecosystems. In addition, macrophytes assimilate and store in their tissues a significant amount of nutrients dissolved in water such as ammonia, nitrates, phosphates, and inorganic carbon (Van Donk and Van de Bund 2002). They can also contribute to increased water transparency by providing a

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Table 7.1 Examples of aquatic macrophytes (floating and immersed), tested for the elimination of different contaminants Plants Lemna gibba Lemna minor

Azolla pinnata Eichhornia crassipes Elodea canadensis Potamogeton lucens Myriophyllum aquaticum

Contaminants Heavy metals Nutriments Sulfadimethoxine Heavy metals Nutriments Heavy metals Nutriments Heavy metals Cadmium Herbicide 137 Casieum, 60cobalt, 54manganese

References Sasmaz et al. (2015) Day and Saunders (2004) Drobniewska et al. (2017) Sasmaz et al. (2015) Zhang et al. (2008) Priya and Selvan (2014) Javed et al. (2013) Knauert et al. (2010) Dhir et al. (2009)

refuge for predators of phytoplankton biomasses such as zooplankton (Hilt and Gross 2008). Plants are generally regarded as major players in the transport and degradation of many well-known contaminants in aquatic environments through their uptake and metabolic detoxification (Dhir et al. 2009). A multitude of researches studies have shown the ability of aquatic plants to absorb, accumulate, and metabolize, directly or indirectly, various contaminants such as heavy metals, nutrients, insecticides, and herbicides (Table 7.1). On the other hand, aquatic plants such as Lemna sp., Myriophyllum sp., and Eichhornia crassipes are particularly important because they are in direct and permanent contact with water contaminants (Tel-Or and Forni 2011).

7.6.2

Role of Macrophytes in the Bioaccumulation and Elimination of Cyanotoxins

Recently, there has been particular attention to the effects of cyanotoxins on aquatic ecosystems (Zanchett and Oliveira-Filho 2013). In order to solve this problem, investigation of the accumulation of cyanotoxins by macrophytes has increased. They described that microcystins can be absorbed and metabolized by several aquatic plants and that toxin transfer along the aquatic food chain may occur (Pflugmacher et al. 2001; Saqrane et al. 2007). In vitro exposures have also shown that macrophytes have the ability to bioaccumulate isolated microcystins from their environment (Table 7.2).

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Table 7.2 Different plants tested for their accumulation capacity of cyanotoxins

Plants Hydrilla verticillata

Egeria densa

Ceratophyllum demersum

Hydrilla verticillata

Vallisneria natans C. demersum, Egeria canadensis Myriophyllum spicatum

MCs MCLR MCRR MCLR MCYR MCRR MCLR MCYR MCRR MCLR MCYR MCRR MCRR MCLR MCRR MCYR

Duration of exposure 7 hours

Initial concentration of MCs (μg/L) 42.2 41.3

Final concentration of MCs (μg/L) 7.23.7

14 hours

104.4

0.7

14 hours

104.4

3.1

14 hours

104.4

40.8

RomeroOliva et al. (2015)

16 hours

100

15.1

10

0

Yin et al. 2005) Pflugmacher et al. (2015)

References RomeroOliva et al. (2014)

MCs microcystins, MC-LR microcystin-LR, MC-RR microcystin-RR, MC-YR microcystin-YR

7.6.3

Mechanisms for Detoxification of Cyanotoxins by Macrophytes

Organisms could protect against the toxicity of microcystins by reducing their concentration in vivo through an intrinsic physiological process (Pflugmacher et al. 1998; Beattie et al. 2003). The first step in the detoxification of cyanotoxins is the formation of a glutathione conjugate (Pflugmacher et al. 1998). After absorption of the toxin by the plant, a small amount will be conjugated directly to the glutathione conjugate. A second amount will be conjugated via the glutathione S transferase system (Pflugmacher et al. 1998). Some of the rest of microcystin will bind to protein phosphatases and inhibit it. These are responsible for regulating important cellular processes such as carbon and nitrogen metabolism, tissue

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development, and photosynthesis (MacKintosh et al. 1990; Hastie et al. 2005). Finally, the rest of the toxin could be absorbed by the chloroplasts, in which three main routes were requested. A non-enzymatic pathway by binding to GSH, enzymatic conjugation to GSH via the GST system, and reactions with proteins or structures of the chloroplast photosynthetic apparatus. To suppress glutathione conjugates, plants transfer these conjugates into the vacuole through multi-resistant proteins (MRPs), which are part of the ABC transporter family, for temporary storage and further processing of GSH conjugates (Walbot et al. 2000; Dixon et al. 2010).

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Chapter 8

Biochemical Methods for Water Purification Gulzar Muhammad, Adeel Mehmood, Munazza Shahid, Raja Shahid Ashraf, Muhammad Altaf, Muhammad Ajaz Hussain, and Muhammad Arshad Raza

Contents 8.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2 Biochemical Treatment Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.1 Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.2 Bioremediation by Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.3 Bioremediation by Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2.4 Bioremediation by Algae . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Prompt amplification in industrialization and population has resulted in the heightened discharge of polluted water containing heavy metals, dyes, pigments, and other organic pollutants into freshwater systems, which upsets the quality of water, causes biotic menaces and harms the health of the human beings. Particularly heavy metals and dyes are harmful to aquatic life and are not biodegradable and must be extracted from wastewater before discharging into freshwater. Treatment of polluted water using various physicochemical methods such as electrolysis, chemical precipitation and flocculation, adsorption, reduction, ion-pair extraction oxidation, and electrochemical treatment is expensive and is associated with larger sludge production. Thus biochemical approaches are a contemporary area of research fascinating the researchers around the globe.

G. Muhammad (*) · A. Mehmood · R. S. Ashraf (*) · M. Altaf (*) · M. A. Raza Department of Chemistry, GC University Lahore, Lahore, Pakistan e-mail: [email protected]; [email protected]; [email protected] M. Shahid Department of Chemistry (SSC), University of Management and Technology, Lahore, Lahore, Pakistan M. A. Hussain Department of Chemistry, University of Sargodha, Sargodha, Pakistan © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 181 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_8

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Bioremediation using bacteria, fungi, algae, and plants is economical and environment-friendly, which focuses on the cost-effective transformation of dissolved contaminants into hefty particles, which can be removed smoothly by the organisms. The chapter provides a comprehensive description of economical, eco-friendly, and efficient biochemical water purification methods with the removal efficiencies. An in-depth look into the difference in the removal efficiencies of different microorganisms and the relation with different parameters such as temperature, pH, and initial dye concentration are also discussed. Keywords Phytoremediation · Biochemical methods · Dyes · Wastewater · Heavy metals · Water purification · Bioremediation · Langmuir model · Freundlich model · Kinetics

8.1

Introduction

Wastewater polluted by effluents such as hazardous dyes and heavy metals from industries is nowadays a severe issue causing various diseases. Water contaminated by human waste, industrial wastes, domestic wastewater, rain runoff, animal wastes, and groundwater infiltration contained 99.9% water and 0.1% contaminants (suspended solids or dissolved materials) by weight. The solid pollutants include detergents, heavy metals, grease, salts, food leftovers, plastics, oils, sands, excrements, and grits (Gray 2000). A small amount of heavy metals in drinking water causes severe diseases to animals and humans. There are some toxic synthetic dyes, which are carcinogenic and mutagenic and create an allergic reaction in the human body. To avoid such problems, domestic, municipal, agricultural, and industrial wastewater are preliminarily treated using different techniques such as flotation, phytoremediation, adsorption, ion exchange, and microbial treatments. Transformation of toxic materials into secure end products is the primary goal of wastewater treatment, leading to safer disposal, public health protection, primary nutrients recycling, cost-effectiveness, and treatment feasibility (Samer 2015). The chapter reviews secondary treatment focusing on the transformation of dissolved materials or particulate matters into larger particles, which are removed by the filtration process. Biological treatment has given preference over chemical treatment because the chemical materials only react with a small number of waste materials and heavy metals, and large portions of waste material remain unaffected. However, chemical materials are very costly and produce a large quantity of chemical sludge. Thus biological methods are appropriate to treat wastewater by producing dense biomass produced from organic matter, which is easily removed by sedimentation. Microbes also feed on dissolved organic matter resulting in small sludge production as compared to the chemical treatment (Samer 2015). The chapter provides a comprehensive description of water purification methods and an in-depth

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look into the difference in the removal efficiencies of different microorganism and the relation with various restrictions like temperature, pH, and initial dye concentration. The chapter also consists of summarized research work of renowned scientists on wastewater treatment, which provides critical discussion to researchers and environmental engineers for wastewater purification.

8.2

Biochemical Treatment Processes

After the removal of solid particles, the water entered into a bioreactor where microorganisms replicate by utilizing organic contents under aerobic (bacteria) or anaerobic conditions (bacteria, algae, and fungi) under bioenvironmental conditions (Gaikwad et al. 2014). Microorganisms remove the organic materials through biological oxidation process depending upon food and oxygen availability. Bacteria are mostly explored for wastewater treatment owing to the formation of basic trophic levels in food chains present in the bioreactor. In addition, fungi are also investigated well for bioremediation of wastewater (Tondee et al. 2008). Microorganisms decompose organic matter present in sewage via biological oxidation and biosynthesis. Biological oxidation usually produces several end products that are released with effluents, and dense biomass is produced from organic matter in biosynthesis which can be separated via the sedimentation process. Various biochemical methods using plants, bacteria, fungi, and algae are discussed in the chapter.

8.2.1

Phytoremediation

The increase in industrialization and overpopulation destroys the environment. The major aqueous pollutants are toxic heavy metal cations which are discharged into the aquatic environment. The available water purification technologies are too expensive and non-environment-friendly. Therefore, researchers are trying to develop economic and eco-friendly tools for water purification. Hydrophytes are considered best to treat wastewater since plants have a superior adsorbing capacity for pollutants. Observations show that surface floating plants and emergent plants adsorb toxic heavy metals through roots; however, submerging plants adsorb heavy metals and other contaminants via roots and leaves (Boyd 1970; Denny 1980; Dhote and Dixit 2009). The aquatic environment has been adversely affected by the heavy metals discharged from metal processing industries. The heavy metals in the aqueous solution are not easily removed via conventional treatments and produce a large amount of poisonous chemical mud. Recently, fungi, algae, bacteria, and plants are widely used for wastewater treatment. In such a study, various biomass fragments of Potamogeton crispus yield turions with a remarkable removal efficiency of total phosphorus, total nitrogen, nitrogen ammonia, and nitrogen nitrates in the water and

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enhance water quality. The biomass fragments depict efficient photosynthesis and growth of the leaves and stem throughout 49 days of evaluation. Moreover, floating particles produce an ecological niche with algae biomass reduction, hence transform the dominated algae community to benthic algae and cyanobacteria (Zhou et al. 2017). In the same way, barley plants cultivated using hydroponic culture technique gain a reasonable height of about 25.0 cm and produce 25–59 t/ha yield, which depends upon seed density when irrigated with wastewater. The hydroponically grown barley reduces wastewater pollutants such as total solids, chemical oxygen demand, NHþ 4 N, NO2-N, PO3 4 -P, and NO3 -N from 52.7% to 60.5%, 72.9% to 83.1%, 76.0% to 76.0%, 97.6% to 99.2%, 87.1% to 95.1%, and 76.9% to 81.6%, respectively. Nevertheless, the effluents formed via the hydroponic system possess a significant value of total solids recommended for aquatic organisms (Snow and Ghaly 2008). Another multipurpose plant, M. oleifera Lam. at an optimal amount of 15.0 mg/L, removes total suspended solids present in the wastewater from about 35.0  1.65 to 3.0  0.3 NTU as compared to aluminum sulfate at a dosage of 55 mg.L1. The plant retains higher treatment capability (91.0%) than aluminum sulfate (79.0%). Furthermore, the plant increases the pH of the water sample from about 5.0 to 6.0, while pH decreases for aluminum sulfate to almost 3.5. The plant exhibits constant conductivity range (113 μS), whereas the electrical conductivity of aluminum sulfate exceeds from the initial reading of 87 to 1120 μS. Similarly, the plant extract is very effective coagulant in removing turbidity for the groundwater, thus proving the plant’s capability as an inexpensive and safer coagulant than aluminum sulfate for water treatment (Tunggolou and Payus 2017). The seed powder is a common material to coagulate water contaminants. The powder is also capable of removing toxic organic dyes and heavy metal ions using catalytic agents. The powder used for water purification consists of copper phosphate and low molecular weight, coagulant protein of the plant to form inorganicprotein nanoflowers. The morphological studies confirm that by increasing the protein concentration in copper phosphate-protein-based nanoflowers, the surface of the tightly packed petals becomes smoother, which adsorbs eosin yellow, Congo red, rhodamine B base, rose bengal, rhodamine 6 G, and methylene blue over the nanoflowers more prominent due to ionic forces between the cation charges present on protein, and is oxidatively removed by Fenton-like mechanism taking place between Cu+2 and OH radicals generated by H2O2. Toxic heavy metal cations (Pb2+, Cd2+, and Hg2+) are also adsorbed by nanoflowers offering 99% specific selectivity for Pb+2. Therefore, the coagulation protein present in the seed powder is essential for scavenging organic dyes and heavy metal cations from aqueous solutions (Fig. 8.1) (Polepalli and Rao 2018). A combination of plants such as L. perenne var. and L. perenne var. Top One can also purify water with the microbial suspension consisting of Bacillus sp. MOE1, Microbacterium sp. MOE2, and two denitrifying species which accumulate polyphosphate. L. perenne var. Top One removes 87% total phosphorus and 40% total nitrogen in 20 days along with phosphorus, ammonium nitrogen, and chemical

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Fig. 8.1 (a) Adsorption of various dyes by seed powder (b) colors of dyes after mixing with seed powder in 12 h (c) diffuse reflectance spectra of seed powder after adsorbing dye: methylene blue (pink); rhodamine B base (cyan); rose bengal (green); eosin yellow (red); rhodamine 6 G (blue); Congo red (black). Inset shows the treatment of seed powder after adsorption of dye with hydrogen peroxide and copper salt [Congo red (black) and methylene blue (pink)]. (d) The adsorption capacity by seed powder as attained from inductively coupled plasma-atomic emission spectroscopy. (Reprinted with permission of American Chemical Society from Polepalli and Rao 2018) Where CR congo red, EY eosin yellow, RB rose bengal, Rbase rhodamine B base, R6G rhodamine 6 G, and MB methylene blue

oxygen demand. Likewise, polyphosphate gathering species such as Bacillus sp. MOE1 and Microbacterium sp. MOE2 succeed in removing 73% total phosphorus, 81% ammonium nitrogen, and 32% total nitrogen that are further heightened via the addition of perennial grasses, particularly L. perenne var. Top One. The unique amalgamation of polyphosphate-gathering species and L. perenne var. Top One positively augments the activities of alkaline phosphatases, invertase, and urease with bioremediation of eutrophic water (Li et al. 2011). Another aquatic plant, Eichhornia crassipes in association with rhizospheric bacteria, upgrades water quality by consuming pollutants as a portion of food by rhizospheric bacteria in batch culture. Naphthalene is entirely (100%) removed by adsorption after the 9th day using Eichhornia crassipes along with rhizospheric bacteria. Plants without rhizospheric bacteria remove 45% naphthalene on the 7th day. Moreover, the usage of naphthalene by Eichhornia crassipes takes place in two phases: speedy first phase (2.5 h) and a sluggish period (2.5–225 h) (NesterenkoMalkovskaya et al. 2012). An economic phytoremediation technique involving plants like water lettuce, water hyacinth, and vetiver grass to treat wastewater minimizes total dissolved solids, electrical conductivity, total suspended solids, hardness, chemical oxygen demand, biological oxygen demand, dissolved phosphorus, oxygen, nitrogen, and heavy metals. The water purification by floating plant system is affected by climate, origin, occurrence, concentrations of contaminants, temperature, and ecological factors (Gupta et al. 2012). Aquatic resources are facing the severe threat of eutrophication for ecosystem stability. Among chemical, physical, and biological methods of wastewater treatment, phytoremediation fascinates the researchers. A hydrophyte water hyacinth (Eichhornia crassipes) is considered as the most exceptional candidate for eutrophic water purification. The plant can remove nitrogen and phosphorous with the

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inhibition of algae growth by involving different functions such as response to strains, biosynthesis, synthesis, and secretion, energy metabolism, photosynthesis, and metabolism of nitrogen and phosphorous (Li et al. 2015). In India, people living in villages purify contaminated water from heavy metals using Strychnos potatorum seeds (cleaning nuts or nirmali). The seeds of the plant have coagulant protein, which is separated by ammonium sulfate fractionation from soluble seed extract. The coagulant protein shows 70% biosorption of Cd (II) ions at pH 5.0 in 6 h following the Langmuir isotherm model and pseudo-second-order kinetics (Saif et al. 2012). Similarly, proteins from Vigna unguiculata and Parkinsonia aculeata seeds possess better coagulation potential than M. oleifera proteins and can act as prominent coagulants for the treatment of water in the future (Marobhe et al. 2007). The literature reveals that the people in developing communities used plants for purification of the water. One such plant, Opuntia species, lowers 98% turbidity by bridging mechanism with the kaolin clay particles at pH 9.8, similar to M. oleifera. Therefore, the plants are more valuable for the water purification technology at low cost and readily available for some developing communities (Miller et al. 2008). Pistia stratiotes (water lettuce) and Eichhornia crassipes (water hyacinth) are beneficial and efficient plants for the treatment of effluents present in water. Eichhornia crassipes (150 g) and Pistia stratiotes (50 g) decrease the dissolved oxygen, pH, biological oxygen demand, turbidity, chemical oxygen demand, PO3 4 ,  NO 3 , NO2 , NH3, and total Kjeldahl nitrogen of wastewater due to production of CO2 which in turn suppresses the microbial growth (Akinbile and Yusoff 2012). M. oleifera is a distinctive plant having a wide range of applications in the global market due to unique adaptive behavior in different farm systems and ecosystems. The plant efficiently removes metal ions, turbidity, organic, and biological species from wastewater even at a low dose, which makes the plant an inexpensive and potential candidate for wastewater purification. The coagulation efficiency of the plant is more considerable for high turbidity wastewater and smaller for low turbidity waters. Besides, the seed extract is not only used as a co-coagulant (primary or secondary) for potable water purification but also for the treatment of polluted water (Kansal and Kumari 2014). Hydrophytes like Salvinia, E. crassipes, Lemna, and Pistia are potential candidates for the biopurification of wastewater by adsorbing pollutants or boosting oxidation processes (Abbasi and Abbasi 2010). One such macrophyte is E. crassipes, which is effectively used for efficient removal of total suspended solids, biological oxygen demand, organic matter, heavy metals, and pathogens. E. crassipes removes 65% of heavy metals and produces biogas on a small scale (Mary Lissy and Madhu 2011). The floating plant bed technology is useful for the remediation of eutrophic water using various species of ryegrasses (L. perenne L). Among 12 species of ryegrasses (Grazer, Rustmaster, Secale cereale, Angus I, Jivet, Energa, Surrey, Abundant, Angus II, Gulf, Barwoltra, and Major) used as floating plant bed systems, three abundant species (Abundant, Angus II, and Major) express greater efficiency toward eutrophic water purification and remove nitrogen and

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phosphorus with efficacy ranging from 52.0% to 74.0% and 75.0% to 85.0%, respectively, in 4 weeks. The higher rate of nitrogen and phosphorus accumulation by different species of ryegrasses is positively correlated with water purification (Ding et al. 2012). Vetiver grass (Vetiveria zizanioides) is another potential grass to treat industrial wastewater. The grass adsorbs maximum contaminants (total dissolved solids, total suspended solids, turbidity, biological oxygen demand, chemical oxygen demand, and nutrients concentration) from wastewater discharged by biogas effluents in Pinora Ltd., Ghana. However, the weed shows low efficiency toward the oil palm industry effluents. A system prepared from constructed wetlands planted with aquatic plants (Eichhornia crassipes and Pistia stratiotes) not only disinfects water but also produces bioethanol. Aforesaid plant species have been selected for the study due to the high growth rate and easy cultivation. Recombinant S. cerevisiae or Escherichia coli strains are also used in combination with plants for simultaneous saccharification and fermentation that yields 0.14–0.17 g ethanol per gram biomass. Moreover, E. coli produces more ethanol with both aquatic plants as compared to S. cerevisiae. The constructed wetlands planted with free-floating aquatic plants and microorganisms are thus proved useful for wastewater treatment and energy production (Soda et al. 2013). In rural areas, readily available sources of water are the primary reason behind the water-borne diseases. A cost-effective source for the treatment of contaminated water is the powder of readily available M. oleifera dried seeds in Africa. The seeds powder at a concentration of 12 g/L and alum (10.0 and 12.0 g/L) purify pond water with acceptable values defined by the World Health Organization (Amagloh and Benang 2009). M. oleifera seeds have an active, positive protein that attracts and captures the suspended particles in the water and eliminates the turbidity during water purification. In one such study, aluminum sulfate (100 mg/L) and M. oleifera coagulants diminish water turbidity within the international guidelines (less than 5 NTU). Further observations reveal that aluminum sulfate energy consumption and carbon dioxide (CO2) emissions are almost 40% and 80% higher than the plant-based coagulant. Moreover, the addition of NaCl can reduce the demanded dosage of the plant (Amante et al. 2016). A cost-effective adsorbent Annona squamosa seeds adsorbs methylene blue and malachite green from aqueous media with the abstraction capability of 24% and 76%, respectively, at an optimum pH 6 and 27  C. The equilibrium data follows the Langmuir model (R2 higher than 0.97) with maximum removal efficiencies of malachite green and methylene blue as about 26 and 8.5 mg/g, respectively. Kinetic model obeys the pseudo-second-order equation (R2 higher than 0.99). Thus, A. squamosa seed can act as a promising inexpensive adsorbent for dye uptake from industrial wastewater (Santhi et al. 2016). Adsorbent from A. squamosa also uptakes dye, orange II via batch process at pH 4 in 40 h. Kinetic modeling showed pseudo-second-order (r2 higher than 0.9854) for all concentrations (5–40 mg/L). Adsorption data fits best to both Freundlich and Langmuir isotherm models showing better removal of orange II dye and certifies the use of A. squamosa as a costeffective adsorbent for water purification (Sonawane and Shrivastava 2011).

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Anionic dyes such as reactive blue 4 and acid blue are removed successfully using chemically modified barley straw. Kinetic study confirms that adsorption takes place following pseudo-second-order, and adsorption data follows Langmuir isotherm at 250  C with maximum removal efficiencies of 51.95 and 31.5 mg/g for acid blue and reactive blue 4, respectively (Ibrahim et al. 2010). Surfactant-modified barley straw also adsorbs acid blue and reactive black 5 with adsorption capability of 1.02  104 and 2.54  105 mol/g, respectively (Oei et al. 2009). Another toxic dye, malachite green, is eliminated from aqueous media by chemically treated Bread skin having removal efficiency of 353.0 mg/g as witnessed by Sips isotherm. The pseudo-second-order kinetics defines well the adsorption processes with a reasonable value of the rate constant (0.32 g/mmol/min) (Chieng et al. 2015). Similarly, breadnut peel (Artocarpus camansi) exhibits a maximum adsorption capability of 409 mg/g according to the Langmuir isotherm model at pH 4.8 for methylene blue. The methylene blue dye shows rapid degradation kinetics following the pseudo-second-order kinetic model, i.e., half of the methylene blue is adsorbed within 5 min at 50 mg/L dye concentration (Lim et al. 2014). As compared to reported fruit biomasses, breadnut (A. camansi Blanco) is found quite useful in the removal of 50% methylene blue. Breadnut reveals optimum extraction efficiency of 369 and 328 mg/g provided by Langmuir and Sips isotherms, respectively. The value of Gibbs free energy is negative, showing exothermic and spontaneous behavior of the adsorption process (Lim et al. 2016). A combination of activated charcoal and coconut husk removes 86% cyanosine from aqueous solutions following first-order kinetics, and adsorption data follows Langmuir and Freundlich isotherm models (Gupta et al. 2010). C. sativus isolated from agricultural wastes exhibits a maximum adsorption capacity of 60.0 mg/g for acid blue 113 dye. Kinetic study indicates pseudo-secondorder kinetics showing quick dye removal by establishing equilibrium within 20 min (Lee et al. 2016). Similarly, C. sativa fruit acts as an efficient biosorbent for malachite green at pH 6.0 following Langmuir, Dubinin-Radushkevich, Tempkin, and Freundlich isotherms models. The fruit removes almost 100% malachite green at 1.0 g/50 mL and 25 mg/L adsorbent dose and initial dye concentration, respectively, at 25  C (Santhi and Manonmani 2011). The same adsorbent uptakes methylene blue from industrial effluents following pseudo-second-order kinetics model, while thermodynamic study discloses exothermic and spontaneous nature of the adsorption process (Shakoor and Nasar 2017). A cost-effective biosorbent C. sativus peel waste quenches effectively harmful crystal violet dye. The adsorption isotherm studies disclosed that the sativus peel system possesses maximum biosorption capability of 149.0 mg/g. The Langmuir adsorption isotherm explains the adsorption data, and adsorption kinetics is described by pseudo-second-order kinetic models (Shakoor and Nasar 2019). An agricultural waste material, Durian rind powder, is also an effective biosorbent to degrade brilliant green and methylene blue dyes with 95.91% and 97.81% efficiency, respectively. Further investigations display that the adsorption data fit well to the Langmuir model than the Freundlich model (Anisuzzaman et al. 2015; Nuithitikul et al. 2010). Another study reveals the potential use of Cempedak

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durian peel as an inexpensive, environment-friendly, and highly proficient adsorbent for methyl violet dye removal from aqueous media. Thermodynamics data explains that the adsorption process is spontaneous, achievable, and endothermic (Dahri et al. 2015a). Papaya seeds are also potential candidates for the uptake of methylene blue from wastewater using the batch process. The adsorption data of papaya seeds follow well the Langmuir model adsorbing 556 mg/g of dye at initial dye concentration and an adsorbent dose of 50–360 mg/L and 0.05–1.00 g, respectively, at pH (3–10) and 30  C (Hameed 2009). The same adsorbent material removes methylene blue via pseudo-second-order adsorption kinetics, and adsorption data harmonize with the Langmuir isotherm model with a monolayer sorption capacity of 512.55 mg/g (Mukhlish et al. 2012; Unuabonah et al. 2009). Papaya seeds also adsorb methylene blue following the abovementioned kinetics and isothermal model with an abstraction efficiency of about 637 mg/g at an optimum pH 12.0 (Paz et al. 2013). Papaya seeds are also employed for the reduction of direct black 38, a tannery dye from aqueous media by keeping in mind the main parameters like pH, contact time, and initial dye concentration, which affect adsorption capacity significantly. The adsorption data is best elucidated by the Langmuir isotherm model with an extraordinary adsorption capability of 440 mg/g. The adsorption kinetics confirms second-order kinetic equation, thus endorsing the usage of papaya seed as an effectual, cheaper, and environment-benign adsorbent for the uptake of direct black 38 from aqueous media (Weber et al. 2013). Similarly, 88% of methylene blue dye is removed using white-rot fungus Phanerochaete chrysosporium in association with rice straw at a primary dye concentration of 400 mg/L (Cheng et al. 2015). Methyl violet 2B is removed from industrial effluents using Casuarina equisetifolia needle, exhibiting a maximum extraction capacity of 164.99 mg/g with experimental data best fitted to the Langmuir isotherm model. The adsorption process follows the pseudo-second-order kinetics. Thermodynamics studies unveil the spontaneity of the process with endothermic nature (Dahri et al. 2013). The potential of C. equisetifolia needle against methylene blue and malachite green by batch adsorption process is also encouraging. The adsorption data suits well to the Langmuir isotherm model and provides confiscation ability of almost 111 and 78 mg/g, respectively. Thermodynamic studies confirm the endothermic and spontaneous nature of the adsorption process (Dahri et al. 2015b). Similarly, C. equisetifolia cone and C. equisetifolia needles remove rhodamine B by batch process providing maximum adsorption capacity of 49.5 and 82.34 mg g1, respectively (Dahri et al. 2016; Kooh et al. 2016). Table 8.1 describes the phytoremediation of various pollutants.

8.2.2

Bioremediation by Bacteria

Currently, water pollution control is a significant challenge around the globe, and different researchers have discovered novel strategies to treat wastewater. In a study, Pseudomonas species were immobilized by sol-gel method to reduce azo dyes such as remazol black (75%), methyl orange (79%), and benzyl orange (83%)

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Table 8.1 Phytoremediation of dyes, metals, and organic and inorganic pollutants Plants Barley M. oleifera Lam L. perenne var. Water hyacinth

Pollutants Chemical oxygen demand Total suspended solids Ammonium nitrogen Naphthalene

% Removal 83.1 32 81 45

M. oleifera K+-cross-linked kappacarrageenan beads S. potatorum seeds Cactus specie Opuntia Water hyacinth Water hyacinth

Pb+2 Crystal violet

99 70

Cd+2 Turbidity Turbidity Heavy metals

70 98 92 65

L. perenne L. A. squamosa seed A. squamosa seed A. camansi Blanco Coconut husk Cucumis sativus fruit

Phosphorous Methylene blue Malachite green Methylene blue Cyanosine Malachite green

85 24 76 50 86 100

References Snow and Ghaly (2008) Tunggolou and Payus (2017) Li et al. (2011) Nesterenko-Malkovskaya et al. (2012) Polepalli and Rao (2018) Mahdavinia et al. (2015) Saif et al. (2012) Miller et al. (2008) Akinbile and Yusoff (2012) Mary Lissy and Madhu (2011) Ding et al. (2012) Santhi et al. (2016) Santhi et al. (2016) Lim et al. (2016) Gupta et al. (2010) Santhi and Manonmani (2011)

Fig. 8.2 Decolorization of 100 mg/mL of (■) benzyl orange, (●) methyl orange, and (▲) remazol black. (Reprinted with permission of Elsevier from Tuttolomondo et al. 2014) Where MO methyl orange, BO benzyl orange, and RB remazol black

enzymatically. The species described above decolorize wastewater owing to the rapid release of extracellular enzymes by immobilized bacterial species (Figs. 8.2 and 8.3). Degradation of dyes by bacteria is advantageous due to high usable cell populations, long-lasting stability, and low cast (Tuttolomondo et al. 2014).

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Fig. 8.3 Decolorization of remazol black, methyl orange, and benzyl orange using immobilized bacteria in aqueous solution after regeneration four times. (Reprinted with permission of Elsevier from Tuttolomondo et al. 2014) Where MO methyl orange, BO benzyl orange, and RB remazol black

Similarly, among ten bacterial strains obtained from wastewater of textile industries, Burkholderia cepacia-TN5 and Comamonas acidovorns-TN1 degrade azo dyes such as acid orange 7 and direct blue 75 after the addition of 1.0 g/L yeast extract under controlled environmental conditions (Alalewi and Jiang 2012). Besides, eight resistant indigenous Gram-negative bacteria such as Delftia sp., Stenotrophomonas sp., Comamonas sp., Chryseobacterium sp., Ochrobactrum sp., two Enterobacter sp., and Providencia sp. from activated sludge successfully remove harmful heavy metals (cadmium, chromium, cobalt, and copper) from industrial effluents. Thus, bacteria along with sludge could be the best choice for heavy metal ions uptake as compared to the activated sludge alone due to inexpensive and environment-friendly nature (Bestawy et al. 2013). The strains of a bacterial consortium including Alcaligenes sp. (DQ779012), Bacillus sp. (DQ779011), and Bacillus licheniformis (DQ79010) also display maximum decolorization (70%) at pH 7 and 37  C in the presence of 0.1% peptone and 1.0% glucose. The relative spectrophotometric and liquid chromatography-mass spectrometric studies reveal that bacterial consortium successfully degrades and transforms sample containing 2-nitroacetophenone, 2,3-dimethylpyrazine, 2,20 -bifuran-5-carboxylic acid, 2-methylhexane, methylbenzene, 3-pyrroline, 2,3-dihydro-5-methylfuran, pchloroanisol, and acetic acid into 2-nitroacetophenone, 2,20 -bifuran, indole, 2-methylhexane, and 2,3-dihydro-5-methylfuran (Bharagava and Chandra 2010). As compared to fungal laccases, bacterial laccases efficiently decolorize textile dyes in wastewater even at basic pH and high temperature. The stability of bacterial laccases is due to the cloning of the thermo-alkali-stable CotA-laccase gene of Bacillus pumilus W3, which is overexpressed in E. coli but successfully regulated in Bacillus subtilis WB600. In a 3-L fermenter, purified CotA-laccase (373.0 U/mL) is used for detoxification and decolorization of colored wastewater with or without

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using mediators. Acetosyringone is the most efficient and potential mediators, which increases the decolorization and decreases the toxicity of textile industry effluents. Thus B. subtilis WB600-5 strain has a vast scope in the production of CotA-laccase, which is a prospective contender for wastewater bioremediation (Guan et al. 2015). On the other hand, Bacillus thuringiensis var. is utilized for the biosorption of noxious metal from the polluted water under different experimental conditions such as pH, initial concentration of metal ions, and temperature. Scatchard and adsorption studies of B. thuringiensis against nickel (II) ions confirm the equilibrium binding, and adsorption data follows Freundlich and Langmuir isotherms. B. thuringiensis’ spore–crystal mixture and vegetative cells adsorbed 16% and 10% Ni (II) ions, respectively, at a rate of 250 mg/L. Thus, B. thuringiensis’ spore–crystal mixture has a better affinity for heavy metal ions than the vegetative cells of the species (Öztürk 2007). Thermophilic bacteria Geobacillus thermodenitrificans is another bacterial species applied for biosorption of heavy metals. After 12 h of biomass introduction, G. thermodenitrificans exhibits maximum affinity toward Pb+2 (37%), Ag+ (43%), Cd+2 (49%), Zn+2 (55%), Cu+2 (57%), Co+2 (80%), Cr+3 (81%), and Fe+3 (91%) ions at optimum pH. The bacterial strain also demonstrates maximum adsorption capacity toward metal ions such as Fe+3 (44%), Cr+3 (39%), Cd+2 (36%), Pb+2 (18%), Cu+2 (13%), Co+2 (11%), Zn+2 (9%), and Ag+ (8%) within 2 h for industrial wastewater (Chatterjee et al. 2010). Likewise, a novel bacterial consortium consisting of Serratia marcescens, Klebsiella oxytoca, Citrobacter sp., and unknown bacterium (DQ817737 (T4)), i.e., MMP1 identified by 16S rDNA-based approach, are isolated from sediments of a waterfall for water purification. The consortium decolorizes waste effluents up to 1.0%, 10%, 7.5%, and 8.0%, after 48 h when used in combination with various media such as caramel (30% w/v), viandox sauce (13.5% v/v), sugarcane (20% v/v), and beet molasses (41.5% v/v), respectively (Jiranuntipon et al. 2008). Another mixture of different dyes (0.5–2.0 g/L) from the textile industry is also removed by bacteria Alishewanella sp. strain KMK6 within 8.0 h. The chemical oxygen demand is reduced to 28% at stable anoxic conditions immediately after decolorization, which is further decreased to 90% by incubation. The bacteria also degrade the dyes to various less toxic products (Kolekar et al. 2013). Furthermore, B. cereus strain HJ-1 along with 0.15 g/L yeast extract and 0.125 g/L glucose decolorizes azo dye reactive black B at pH 8, 50  C, and 1.0 critical micelle concentration of Triton X-100. The results of the denaturing gradient gel electrophoresis technique show that bacterium survives in river sediments even after 12 days of incubation (Liao et al. 2013). Bioflocculant producing bacterium extracted from activated sludge decolorizes 97.0% wastewater dyes such as whale and mediblue in Durban, South Africa, at pH 7 and 35  C, while fawn and mixed dye are efficiently removed at pH 9 (40–45  C) and 10 (35–40  C), respectively. The bioflocculant from bacterium provides cheap, eco-friendly, and cleaner technology for wastewater purification (Buthelezi et al. 2012). In addition, a consortium of Burkholderia sp. and Pseudomonas sp. from coking wastewater sludge removes 93.0% pyrene (100 mg/L) within

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36 h at pH 7 and 37  C facilitated by the addition of glucose or anthraquinone. Degradation of pyrene follows first-order reaction kinetics (Deng et al. 2012). Similarly, three different kinds of bacteria (Aeromonas eucrenophila, Pseudomonas aeruginosa, and Clavibacter michiganensis) immobilized on cellulose acetate nanofibrous web exhibit 95% methylene blue removal from aqueous solution similar to that of unrestricted bacteria. Bacteria immobilized nanofibrous web retains 45% of removal capacity after the end of the fourth cycle suggesting an inexpensive wastewater treatment process for decolorization of methylene blue dye (San et al. 2014). In the same way, sulfonated reactive dye green HE4BD is rapidly decolorized by a bacterial consortium comprising Micrococcus glutamicus NCIM-2168 and Proteus vulgaris NCIM-2027. The decolorization activity of consortium increases after the addition of inexpensive co-substrate, such as extracts of agricultural wastes. Total organic carbon and chemical oxygen demand are reduced to 90% within 24 h by oxidoreductive enzymes released from consortium showing mineralization of sulfonated reactive dye green HE4BD to nontoxic products. The study can be helpful in cheaper bioremediation of many reactive dyes and textile industries effluents (Saratale et al. 2010). Likewise, the bacterial strain RMLRT03 identified as Staphylococcus hominis by 16S rDNA degrades acid orange dye in Bushnell and Haas medium. In static conditions, the addition of yeast extract and glucose as co-substrate further boosts the decolorization performance of bacterial strain. An additional advantage of S. hominis RMLRT03 strain for wastewater treatment is the resistance against acid orange dye at pH 7, dye concentration of 600 mg/L, and 35  C when incubated for 60 h (Singh et al. 2014). The biodegradation of reactive blue 4 (300 mg/L) using bacterial consortium comprising Paenibacillus polymyxa and Salmonella subterranea with various co-substrates such as glucose, propionate, and butyrate at 35  C occurs through acidogenesis at 24%, 12%, and 10%, respectively, after 24 h incubation (Watanapokasin et al. 2009). Another study explores two strains of Pseudomonas sp. (PCT01 and PTS02) which degrade 229–461 mg/L of phenol after 18 h incubation. However, pyridine and other compounds in coking wastewater increase bioremediation time from 20 to 32 h at 160 to 280 mg/L initial concentration of phenol. The degradation kinetics discloses the highest rate constants of 1.25 and 0.75 h1 for PCT01 and PTS02, respectively, in the second-stage model at minimum phenol concentration using a single substrate (Zhu et al. 2012). In another study, two species of Pseudomonas (Sz6 and SDz3) efficiently decolorizes 85% diazo Evans blue dye within 48 h, and complete removal takes place after 120 h (Zabłocka-Godlewska et al. 2012). An innovative way to get rid of effluents is by coupling photosynthetic bacteria with membrane bioreactor separation under optimum conditions such as natural light, micro-aerobic environment, and reasonable food to microbe ratio (2.0). The removal of chemical oxygen demand reaches 99.3%, with 99.5% biomass recovery under optimum conditions for membrane operation such as 0.1/0.2 MPa pressure and 27.5 L/h flow rate. The backwash cycle of photosynthetic bacteria-based membrane bioreactor is extended to 48 h due to operation simplicity, lesser backwash water consumption, and durability of the membrane as compared to traditional

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membrane bioreactor systems (0.5–3.0 h). Moreover, recovered biomass serves as raw materials in fish farming or agriculture that eliminates secondary pollutants. Further, photosynthetic bacteria did not form flocs, and the extracellular polymeric substance is little in the system. The photosynthetic bacteria-based membrane bioreactor system is preferred over a conventional membrane bioreactor system due to lesser contamination, more exceptional durability, and membrane backwash cycle (Lu et al. 2013). The microbial consortium involving Staphylococcus spp., Pseudomonas spp., Streptomyces spp., Bacillus spp., and Actinomycetes spp. diminish biological oxygen demand and chemical oxygen demand up to 94.0% and 90.0%, respectively as compared to individual microbial species ranging from 59.0% to 77.0% and 42.0% to 60.0% for biological oxygen demand and chemical oxygen demand. The decline in total dissolved and suspended solids by consortia are also noted to 74.0% and 80.0%, respectively (Gaikwad et al. 2014). A magnetotactic bacterium uptakes hexavalent chromium (Cr6+) under optimum conditions of temperature, contact time, pH, microorganism concentrations, and initial concentration of Cr6+, Co2+, and Cu2+ ions in a batch-stirred system. At pH 6.00 and temperature 29  C, living cells remove 77% of metal ions in just 10 min. Furthermore, Co2+ and Cu2+ ions are supportive in Cr6+ removal under an applied electric field. The removal data follows the Langmuir isotherm model more appropriately than the Freundlich adsorption model. The novel results of the study support magnetotactic bacteria as a favorable aspirant for the elimination of heavy metal (Qu et al. 2014). A bacterial strain, Shewanella decolorationis (NTOU1), is capable of decolorizing and detoxifying crystal violet dye in the absence of oxygen at pH 8–9 and 30–40  C. Experimental results confirm crystal violet decolorization at a concentration and rate of 1500 mg/L and 298 mg/L/h, respectively, in the presence of formate and ferric citrate in the medium. According to gas chromatography-mass spectrometry analysis, crystal violet degrades into less toxic products such as N,Ń́-bis (dimethylamino) benzophenone (Michler’s ketone), benzophenone, N,Ndimethylaminobenzaldehyde, N,N-dimethylaminophenol, 4-methylaminophenol, and [N,N-dimethylaminophenyl][N-methylaminophenyl] (Fig. 8.4). The inspirational results support the use of the abovementioned bacteria for wastewater treatment (Chen et al. 2008). The Lactobacillus plantarum (PV71-1861) from pickles in Thailand decolorizes melanoidin pigment present in molasses wastewater up to 68.0% from the solution which contains 0.1% KH2PO4, 0.4% yeast extract, 2.0% glucose, and 0.05% MgSO4.7H2O in 7 days under static conditions (pH 6 and 30  C). The gel filtration chromatograms show that large molecular weight fraction of melanoidin pigment solutions remain in the waste effluent while the small molecular weight fraction undergoes degradation by the strain (Tondee and Sirianuntapiboon 2008). Among 100 isolates of Bacillus sp., new laccase gene cloned bacterial species, Bacillus amyloliquefaciens 12B degrades dye reactive blue 52 at elevated temperature, pH 4.0 and 7.0 (Lončar et al. 2013). An Enterobacter strain (GY-1) portrays high decolorization activity against reactive black 5 present in wastewater from textile industries due to extracellular enzymes. The decolorization of the dye improves by

8 Biochemical Methods for Water Purification H3C

H 3C

+ CH3 N

CH3

H3C N

195

N

C

H3C

– e-

CH3 N

O

CH3

CH3

N

H3C

+

HO

O

CH3 H + HO

H3C N,N-Dimethylaminobenzaldehyde

Smaller products

N CH3 N,N-Dimethylaminophenol

CH3 Michler's ketone

C

CH3

Leucocrystal violet

N

N

N

C H

H3C

Crystal violet

H3C

CH3

H C + e- 3

CH3

H3C

N

Smaller products

N CH3 N,N-Dimethylaminophenol

Smaller products

Fig. 8.4 Proposed degradation products of crystal violet dye by Shewanella sp. NTOU1. (Modified after Chen et al. 2008)

enhancing the size of the inoculum and is maximum at 16 mL inoculum size (Fig. 8.5). Moreover, decolorization is also boosted at neutral pH, 35  C and lower concentrations of NaCl (Figs. 8.6 and 8.7) (Chen et al. 2011). Two bacteria, Staphylococcus capitis and Bacillus sp. JDM-2-1, reduces hexavalent chromium into trivalent, which confirms that both the species uptake heavy metal ions from wastewater and are resistant toward Cr(VI) at a concentration of 2800 and 4800 μg/mL, respectively. The species also demonstrates resistance to Cu2+ (200 μg/mL), Cd2+ (50 μg/mL), Hg2+ (50 μg/mL), Pb2+ (800 μg/mL), and Ni2+ (4000 μg/mL). Moreover, the species abridge 81% and 85% hexavalent chromium in 96 h from the medium, while percentage removal from wastewater effluents is up to 89% and 86%, respectively, after 144 h. The uptake and reduction of chromium may be due to the production of a protein with a molecular weight of 25 kDa during bioremediation (Zahoor and Rehman 2009). Another bacterial consortium “Bx” biodegrades Blue Bezaktiv S-GLD 150 dye in inoculated sequencing batch reactor with maximum degradation efficiency (97%) and minimizes chemical oxygen

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Fig. 8.5 Decolorization of reactive black 5 using strain GY-1 at optimum temperature and pH. (Reprinted with permission of Elsevier from Chen et al. 2011)

demand up to 98% in the absence of oxygen at pH 7 and dye loading rate of 15 g dye m3.d1 at room temperature (Figs. 8.8 and 8.9). The percentage of decolorization and chemical oxygen demand exclusion drop to 70% and 90%, respectively, when the dye loading rate exceeds 20 g dye m3.d1 (Khouni et al. 2012). Using enrichment cultivation, a bacterial alkali consortium is obtained which treats textile wastewater in sequencing batch reactor and biocontact oxidation reactor with almost alike treatment productivities against polyvinyl alcohol (75–81.0%) at acidic pH during 3 months. The removal of chemical oxygen demand using bacteria described above ranges from 74.0% to 77.0%. The bacterial species in both the reactors express diverse behavior and contain Bacteroidetes, Actinobacteria, Proteobacteria, Firmicutes, and an unknown cluster. Among the species mentioned above, the isolated Paracoccus sp. effectively degrades polyvinyl alcohol (Yang et al. 2011). Table 8.2 summarizes the decolorization efficiency of various bacterial strains.

8.2.3

Bioremediation by Fungi

The industrial effluents are the primary sources to affect the water clearness and gas solubility along with the carcinogenicity and toxicity in water bodies. The removal of the effluents is necessary before disposal into natural water. The ligninolytic fungi

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Fig. 8.6 Decolorization of reactive black 5 using strain GY-1 at different pH (a) and temperatures (b) in 24 h. (Reprinted with permission of Elsevier from Chen et al. 2011)

produce nonspecific oxidative enzymes that are helpful in removing the colors of such dyes. The immobilized growing cells are found active than free cells. So the future research of developing feasible microbiological processes with immobilized cells should not only focus on the treatment of wastewater but also on the reuse. Thus, there is a need to design special bioreactors with optimized parameters for the decolourization of dyes. Biodegradation of effluents results in toxic intermediates and metabolites that must be evaluated. There are various studies available regarding

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Fig. 8.7 Decolorization of reactive black 5 using different concentrations of common salt in the presence of strain GY-1 at optimum temperature and pH. (Reprinted with permission of Elsevier from Chen et al. 2011)

Fig. 8.8 Photograph showing decolorization of Blue Bezaktiv S-GLD 150 dye using bacterial consortium “Bx” in aqueous mead (A) after treatment and (B) before treatment. (Reprinted with permission of Elsevier from Khouni et al. 2012) Where MLVSS mixed liquor volatile suspended solids

azo dye degradation, but the complete mechanism of dyes degradation using whiterot fungi and ligninolytic enzymes of the fungus is still limited (Couto 2009). Another bacteria-yeast consortium BL-GG decolorizes 98% scarlet RR at pH 9 and 40  C in 18 h from industrial waste effluents. The bacterial yeast consortium also reduces chemical oxygen demand and biological oxygen demand up to 74% and 68%, respectively, within 48 h. The degradation efficiency owes to various enzymes such as laccase, NADH-DCIP reductase, tyrosinase, and veratryl alcohol oxidase, and each enzyme follows a different pathway to biotransform scarlet RR into multiple products (Fig. 8.10) (Kurade et al. 2012). The biochemical and morphological investigations of newly white-rot fungal strain define laccase production in an economical medium consisting of groundnut shell, synthetic dyes, and algal bloom in submerged shaking conditions at 30  C and pH 5.0. The efficient decolorization of synthetic dyes by white-rot fungal strain takes place at pH 5.0 and 30  C. Using sodium dodecyl sulfate-polyacrylamide gel

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Totale decolourization

100

Adsorption

Biodegradation

Colour removal (%)

90 80 70

62.65

60 50

54.68 47.68 40.47

40 30

36.78

22.18

38.98 32.44

17.9

20

27.11

15.24

11.87

10 0 3.33

6.66

13.33

20

Fig. 8.9 Removal efficiency of bacterial consortium “Bx” against Blue Bezaktiv S-GLD 150 dye solutions after incubating for 55 h using sodium azide (total decolorization) and without using sodium azide (adsorption). (Reprinted with permission of Elsevier from Khouni et al. 2012)

electrophoresis method, the molecular mass and isoelectric point value are investigated to be 70 kDa and 4.8, respectively. Strain mentioned above degrades indigo carmine into isatin sulfonic acid and 4-amino-3-methyl-benzenesulphonic acid, whereas p-hydroxybenzene sulfonic acid and p-N,N0 -dimethylamine phenyldiazine are obtained from methyl orange-degraded metabolites. The investigation would provide a road map to produce laccase enzymes from a low-cost substrate on a larger scale (Mishra et al. 2011). In Thailand, molasses wastewater has been successfully decolorized using Issatchenkia orientalis by degrading melanoidin pigment (91.0%) and decreasing chemical oxygen demand and biological oxygen demand to 80.0 and 77.0%, respectively, at pH 5.0 and 30  C in a malt extract-glucose-peptone broth. Gel filtration chromatography reveals that small molecular weight fractions remain in the effluent while large molecular weight fractions are quickly isolated from the melanoidin pigment solutions (Tondee et al. 2008). The significant decolorization of orange G from aqueous medium takes place by Trichoderma sp. and A. niger biomasses attaining 0.45 and 0.48 mg/g monolayer saturation capacity, respectively. Adsorption data follows Langmuir and Freundlich isotherms, and pseudo-second-order kinetic model with rate constants of 10.0–0.80 g/mg/min for A. niger and 8.0–0.40 g/mg/min for Trichoderma sp. is achieved. Additionally, A. niger exhibits better biosorption than Trichoderma sp. (Sivasamy and Sundarabal 2011). Likewise, Candida krusei degrades basic violet 3 in a semi-synthetic medium of sucrose, lactose, maltose, peptone, glucose, yeast extract, urea, and ammonium sulfate. Yeast mentioned above displays 100% and 74% degradation in sugarcane bagasse and sucrose augmented media, respectively. Actually, basic violet 3 undergoes stepwise reduction and demethylation reaction to produce mono-, di-, tri-, tetra-, penta, and hexa-demethylated basic violet 3 products, which then undergoes complete degradation by the enzymes like malachite greenreductase, lacasse, tyrosinase, NADH-DCIP reductase, lignin peroxidase, and

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Table 8.2 Degradation efficiency of various bacterial strains with optimum parameters

Bacteria Pseudomonas sp.

Stenotrophomonas sp. Enterobacter sp. Ochrobactrum sp. Comamonas sp. Bacterial consortium B. thuringiensis G. thermodenitrificans

Bacterial consortium B. cereus strain HJ-1 Bioflocculant-producing bacteria Pseudomonas sp. and Burkholderia sp. Bacterial consortium Consortium GR

S. hominis S. subterranean, P. polymyxa Pseudomonas sp. PCT01 Pseudomonas sp. PTS02 Pseudomonas strains

Bacterial consortia

Dye/heavy metal Benzyl orange Methyl orange Remazol black Cadmium Copper Chromium Cobalt Aromatic compound Nickel (II) Cd+2 Fe+3 Cr+3 Co+2 Cu+2 Zn+2 Ag+ Pb+2 Melanoidins Reactive black B Wastewater dyes mixture Pyrene Methylene blue Sulfonated reactive dye green HE4BD Total organic carbon Chemical oxygen demand Acid orange dye Reactive blue 4

Removal efficiency (%) 83 79 75

Adsorption capacity (mg/L)

320 275 29 140 70

Bestawy et al. (2013)

Bharagava and Chandra (2010) Öztürk (2007) Chatterjee et al. (2010)

16 36 44 39 11 13 9 8 18 10 48 97 93 95 90

Jiranuntipon et al. (2008) Liao et al. (2013) Buthelezi et al. (2012) Deng et al. (2012) San et al. (2014) Saratale et al. (2010)

600 24

Phenol

References Tuttolomondo et al. (2014)

229

Singh et al. (2014) Watanapokasin et al. (2009) Zhu et al. (2012)

461 Diazo Evans blue dye

85

Cr6+

77

ZabłockaGodlewska et al. (2012) Qu et al. (2014) (continued)

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Table 8.2 (continued)

Bacteria Bacterial strain, NTOU1 L. plantarum No. PV71-1861

Dye/heavy metal Crystal violet

S. capitis Bacillus sp. JDM-2-1 Bacterial consortia “Bx”

Removal efficiency (%)

Melanoidin pigment

68

Cr(VI)

81 85 97 98

Blue Bezaktiv S-GLD 150 Chemical oxygen demand

CH2CH3 H N N

N

Adsorption capacity (mg/L) 1500

References Chen et al. (2008) Tondee and Sirianuntapiboon (2008) Zahoor and Rehman (2009) Khouni et al. (2012)

CI

O Scarlet RR

O

NO2

Veratryl alcohol oxidase or Laccase N

Veratryl alcohol oxidase

CHO H2N + CH2CH3

N

O

CI

NO2

Laccase

N

O

NO2

N H 2N

CI

Veratryl alcohol oxidase

Laccase N

CONH2 + CH2CH3

N

CI

N

CHO

CH2CH3 O

Fig. 8.10 Probable products of degradation of scarlet RR by bacterial consortium. (Modified after Kurade et al. 2012)

azoreductase. The study proves the fungal species as an efficient decolorizing agent for basic violet 3 using an economic nutrient source like sugarcane bagasse extract (Deivasigamani and Das 2011). Another fungal consortium comprising of Mucor hiemalis, Galactomyces geotrichum, and A. niger eradicates chemical oxygen demand efficiently (55–75%) of diluted milk, model medium. The consortium also decreases the chemical oxygen

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NH2 SO3Na

O

NH2 SO2Na SO2(CH2)2OSO3Na

O

NH

Remazol brilliant blue R

OH O Sodiuml-amino-9,10-dioxo-9,10dihydroanthracene-2-sulphonate + SO2(CH2)2OSO3Na NH2 Sodium2-((3aminophenyl)sulfonyl)ethylsulfate

Fig. 8.11 Probable products of remazol brilliant blue R dye by Polyporus sp. S133. (Modified after Hadibarata et al. 2012)

demand of the dairy effluent in the industrial biological tank from 100 to 50–70 mg/ L. Moreover, fungal consortium cuts hard and non-biodegradable chemical oxygen demand from 451–1111 to 257–153% (Djelal and Amrane 2013). White-rot fungi Polyporus sp. S133 releases an enzyme, laccase, which biodegrades anthraquinonetype dye (Remazol Brilliant Blue R) into two products (Fig. 8.11). The purified laccase (1.5 U/L) decolorizes completely 200 mg/L Remazol Brilliant Blue R dye by adjusting pH, temperature, and non-ionic surfactant using N-hydroxybenzotriazole mediator at 50  C and pH 5 (Hadibarata et al. 2012). Similarly, purified laccase from Pestalotiopsis sp. offers 88% decolorization efficiency of direct fast blue B2RL azo dye at pH 4.0 in 12 h suggesting enzymatic removal of industrial effluents (Hao et al. 2007). Penicillium sp., a fungal strain QQ isolated from activated sludge, decolorizes 70% brilliant red X-3B dye aerobically through biosorption at pH 5 and 6% salinity. The strain also degrades azo dye entirely under the same anaerobic conditions. The fungal-bacterial consortium consisting of Sphingomonas xenophaga QYY and strain QQ co-cultures displays better performance than individual strain. The sorption of dye by consortium can be enhanced by the addition of a small quantity of surfactant in weak acidic conditions (Gou et al. 2009). Another study reveals the efficient removal of Cd2+ ions by heavy metal-resistant yeast, Candida tropicalis, which adsorbs 80%, 69%, and 57% from the solution after 144, 96, and 48 h, respectively. The yeast also removes 73% and 56% Cd+2 ions from polluted water in 12 and 6 days, respectively. The yeast also resists heavy metals such as Zn2+ (3100 mg/L), Ni2+ (3000 mg/L), Cu2+ (2300 mg/L), Pb2+ (1200 mg/L), Hg2+ (2400 mg/L), and Cr6+ (2000 mg/L) (Rehman and Anjum 2010). The pure culture of recent filamentous fungi strain, Gliomastix indicus in the remodeled Czapeck medium, decolorizes 90% para-cresol in 108 h when dye concentration is 700 mg/L. Degradation data follow kinetics models such as

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Andrews, Webb, Haldane, and Yano with a correlation coefficient greater than 0.99 (Singh et al. 2008). Two microbial cultures with fungal consortia consisting of 21 fungal strains (70% from genus Candida) possess significant decolorization potential against acid red 249, reactive black 5, reactive red M-3BE, and wastewater from the textile industry in 4.5 L continuous bioreactor. The nutrient source for the bioreactor contains glucose (0.5 g/L) and ammonium sulfate (0.10 g/L). Moreover, color removal efficiency enhances with increasing glucose concentrations in the influents (Yang et al. 2009). Apart from high chemical oxygen demand, textile waste also poses severe problems because of the toxic and unmanageable property of dyes. Mucorales fungi strain cultured in two different media yields good biosorption potential, which leads to a decrease in chemical oxygen demand and decolorization up to 94% and 58%, respectively. The Lemna minor toxicity test results in the remarkable elimination of toxic compounds after biosorption and shows decolorization mainly due to detoxification of treated polluted water (Prigione et al. 2008). Five plant species such as Quercus ilex, Cupressus sempervirens, Laurus nobilis, Salix sp., and Pinus mugo treat olive mill wastewater with 50% efficiency due to increase in peroxidase, laccase, and β-glucosidase activities, recovery of bacterial biomass, and transitory occurrence of specialized fungal communities in the presence of Geotrichum candidum. Among rhizosphere fungal strains, Penicillium aurantiogriseum and Penicillium chrysogenum are superior in treating olive mill wastewater (Bodini et al. 2011). A. nidulans, R. arrhizus, and T. viride exhibit maximum acceptance ability of about 26.0 mg/g for Pb, 13.0 mg/g for Cd, and 2.50 mg/g of Cr, respectively. The reported strains prove to be active biosorbent for the uptake of heavy metals from polluted water (Kumar et al. 2014). Various fungi species such as Helminthosporium sp., Aspergillus fumigatus I-II, A. flavus I–V, Mucor rouxii mutant, Cladosporium sp., M. rouxii IM-80, Candida albicans, and Mucor sp1 and sp2 demonstrate biosorption of mercury (II) ions. Among the 14 strains of fungi, M. rouxii IM-80, M. rouxii mutant, Mucor sp1, and Mucor sp2 remove heavy metals up to 95.3%, 88.7%, 80.4%, and 78.3%, respectively, using dithizone at 30  C and pH 5.5 after incubating for 24 h with fungal biomass (1.0 g/ 100 m/L) (Martínez-Juárez et al. 2012). The industrial wastewater containing dye effluent is successfully treated with A. fumigatus XC6 isolated from mildewing rice straw. The fungus degrades dye over the pH range of 3.0–8.0 using 0.20% ammonium chloride or ammonium sulfate or 1.0% potato starch or sucrose as a supplement. The remarkable results prove that A. fumigatus XC6 is an active species for the decolorization of toxic dyes in textile effluents and, hence, could be beneficial for the dyed wastewater (Jin et al. 2007). Table 8.3 summarizes the dyes and heavy metals adsorbed by various species of fungi.

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Table 8.3 Removal efficiency of fungal species Fungi used Bacteria-yeast consortium

I. orientalis

C. krusei Fungal consortium Polyporus sp. S133 Pestalotiopsis sp. Penicillium sp. C. tropicalis G. indicus Fungi consortium

Fungi M. rouxii IM-80 M. rouxii mutant Mucor sp1 Mucor sp2

8.2.4

Dye/heavy metal Dye Scarlet RR Chemical oxygen demand Biological oxygen demand Melanoidin pigment Chemical oxygen demand Biological oxygen demand Basic violet 3

Removal efficiency (%) 98 74

References Kurade et al. (2012)

68 91 80

Tondee et al. (2008)

77 100

98

Deivasigamani and Das (2011) Djelal and Amrane (2013) Hadibarata et al. (2012)

88

Hao et al. (2007)

70 80

Gou et al. (2009) Rehman and Anjum (2010) Singh et al. (2008) Yang et al. (2009)

Chemical oxygen demand Remazol Brilliant Blue R Direct Fast Blue B2RL azo dye Brilliant Red X-3B Dye Cd2+

50

Para-cresol Reactive red M-3BE Reactive black Acid red 249 Real textile wastewater Hg+2

90 65 80 94 89 95.3

Martínez-Juárez et al. (2012)

88.7 80.4 78.3

Bioremediation by Algae

Like plants, bacteria, and fungi, algae have also been extensively evaluated for the decolorization of wastewater from various industries containing heavy metals, dyes, and pigments. The aquatic pollutants directly affect the environment or indirectly damage the balanced biological food web. The aquatic chemical pollutants include herbicides, organochlorine compounds, domestic and municipal wastes, petroleum products and heavy metal cations, which have severe effects on the hydrosphere. Marine algae are used in different ways such as native dried algal biomass, reinforced biomass, immobilized algal biomass, dried algal biomass, or algal polysaccharides for the synthesis of hydrogels and alginate composites.

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In a study, using phase inversion technique, a microalga, Scenedesmus quadricauda, is trapped into beads of calcium alginate/polyvinyl alcohol composite hydrogel for the uptake of copper and cadmium ions. The entrapped microalga in composite adsorbs almost 0.97 and 0.69 mmol/g of copper (II) and cadmium (II) ions, respectively, at pH 6. The biosorption escalates by increasing the initial concentration of heavy metal ions and temperature. The biosorption process is spontaneous due to the negative value of Gibbs free energy (Bayramoglu and Arica 2011). Eight fresh aqua microalgae strains including Chlorella sp. HQ, C. vulgaris, C. emersonii, C. pyrenoidosa, S. quadricauda, Scenedesmus dimorphus, S. obliquus, and Scenedesmus species produce lipids and purify water. The strains C. pyrenoidosa, Chlorella sp. HQ, and S. obliquus show dominancy in the accumulation of biomass, whereas the superior biomass producers are not the top lipid producers. S. dimorphus and Chlorella sp. HQ produced the lowest lipid content (30%), while S. quadricauda achieves the highest lipid content (66.0%). The extracted microalga Chlorella sp. HQ is the plausible participant for the treatment of water associated with lipid production because the species removes total nitrogen and total phosphorous up to 53% and 85%, respectively. The study concludes that Chlorella sp. HQ serves best for the algae-based synchronous and wastewater purification process along with biodiesel production (Zhang et al. 2016). Similarly, Chlorella vulgaris and Chlorella salina have potential to decrease pH, biological oxygen demand, total dissolved solids, chemical oxygen demand, ammonia, nitrate, sulfate, sodium, magnesium, phosphate, calcium, potassium, and other heavy metals (Cu, Zn, Mn, Co, Ni, Cr, and Fe). Both the species show 14–100% efficiency for heavy metals removal. C. vulgaris removes heavy metals more effectively than C. salina (El-Sheekh et al. 2016). Novel pelletization technology produces pallets of microalga, (C. vulgaris UMN235) and two fungal species (Aspergillus sp. UMN F01 and UMN F02) for wastewater treatment. The pellets remove concentrated nutrients such as ammonium, total nitrogen, total phosphorus, and chemical oxygen demand at rate of 100%, 58.85%, 89.83%, and 62.53%, respectively, while the same diluted nutrients are removed at a rate of 23.23%, 44.68%, 84.70%, and 70.34% from wastewater (Zhou et al. 2012). Organic and inorganic contents are released into water bodies causing organic and inorganic pollution. The standard primary and secondary treatment tactics of the wastewaters are delivered in several places, to be able to put off the quickly settled materials and oxidize the organic pollutants in polluted water. The culture of microalgae presents an exciting step for wastewater remediation, mainly because microalga is linked to tertiary treatment with the synthesis of very valuable biomass. The cultures offer a unique strategy for the tertiary and quandary treatments and can grow even in the presence of inorganic nitrogen and phosphorus. The study highlights the functions of micro-algae such as minimization of biological oxygen demand, elimination of nitrogen and phosphorus, prevention of coliforms, and discharge of heavy metals. Moreover, the higher concentrations of inorganic phosphorus and nitrogen contents in many of the wastewater bodies are used for the cheap and easily profitable nutrient source for the assembly of microalgae biomass.

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The biomass is employed for the production of methane liquid fuel like pseudovegetable fuel, fine chemicals, and animal and fish feed (Abdel-Raouf et al. 2012). Adsorption of multi-metals studied using microalgae-treated bark shows an increase in metal adsorption efficiency. The adsorption of higher concentrations of Co, Zn, Ni, and Cd in solution is independent of the water hardness, whereas the uptake of the metals from diluted solution decreases with an increase in water hardness. The algae bark treatment enhances the uptaking of Co, Zn, Ni, and Cd by 50% using a high concentration of metals and Ca2+ solutions. Algae-treated bark uptakes 100% Cu and Pb from low levels of metals and Ca2+ containing water. The 50–60% uptake efficiency is observed for other metals with low concentrations of metals and Ca2+ at pH 3.0. The Pseudokirchneriella subcapitata and Chlorella sp. are the green algae showing better outcomes in metal adsorption and proving efficiency in uptake capacity of the bark (Lourie et al. 2010). Thus, the literature reveals that algae are potential aspirants for the management of polluted water.

8.3

Conclusions

Wastewater discharge from textile industries adds dyes and toxic materials in the freshwater stream. Although several wastewater treatment techniques are available, there is still a dire need to develop a method that can treat a large volume of wastewater in a shorter time efficiently. Biochemical methods using plants, bacteria, fungi, and algae degrade pollutants by the release of enzymes and can be the future of wastewater treatment owing to environment-friendly and inexpensive nature. Biochemical methods are also preferred over physiochemical methods because of the production of a lesser amount of sludge and feeding of microorganisms on the sludge produced. Biochemical processes decolorize dyes and other pollutants; however, the mechanism of degradation is still not completely understood. Moreover, the investigation of intermediates produced during the degradation of contaminants along with the toxicity is always demanding. Microorganisms degrade dyes and pigments by releasing enzymes. However, the mechanism of degradation of dyes by various enzymes is still not adequately investigated. Studies also reveal that the consortium of bacteria produces better results than individual species. However, strenuous efforts are required to explore the success of a combined effect in wastewater treatment for commercial applications. Further challenges are to probe alterations in degradation mechanisms by consortia, intermediates, metabolites, and enzymes. Some enzymes also produced insoluble intermediate products such as phenols that are more toxic than dyes. Therefore, extensive research is mandatory to investigate the combined effect of enzymes to produce less or nontoxic intermediates.

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Chapter 9

Biosorptive Removal of Toxic Pollutants from Contaminated Water A. Saravanan and P. Senthil Kumar

Contents 9.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2 Conventional Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.1 Membrane Separation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.2 Ultrafiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.3 Nanofiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.4 Reverse Osmosis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.5 Electrodialysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.6 Solvent Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.7 Chemical Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.8 Ion Exchange . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3 Biosorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.1 Mechanism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.4 Biosorbent Material . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.4.1 Microbial Adsorbent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.4.2 Agricultural Biomass . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.4.3 Algal Biomass . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.5 Desorption and Regeneration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.6 Future Perspective and Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract The presence of organic and inorganic contaminants in water environment is a challenging issue due to their high toxicity. Biosorption procedure has been utilized for treatment of wastewater because of its higher efficiency and lower costs when contrasted with other conventional methods. Among the

A. Saravanan Department of Biotechnology, Rajalakshmi Engineering College, Chennai, Tamil Nadu, India P. Senthil Kumar (*) Department of Chemical Engineering, SSN College of Engineering, Chennai, Tamil Nadu, India SSN-Centre for Radiation, Environmental Science and Technology (SSN-CREST), SSN College of Engineering, Chennai, Tamil Nadu, India © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 213 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_9

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elective sorbent materials, the utilization of biomass material for the adsorption of toxic pollutants has earned expanding consideration around the world. Since biosorbent have numerous points of interest, such as different sorts of raw materials, recyclability, and low cost, it can accomplish the impact of transforming waste into fortune when utilized for environmental remediation applications. A statistical analysis of the adsorption isotherm, kinetics, and thermodynamics for the biosorption study was revealed in this chapter. In addition, this chapter overview summarizes the source and generation of biomass, calls attention to its sorption study for the removal of contaminants from the water source, and additionally recognizes future research needs and challenges on the adsorption of toxic contaminants by biological material. Keywords Biosorption · Water treatment · Toxic pollutants · Biological material · Removal

9.1

Introduction

Wastewater containing toxic pollutants has been directly discharged into the environment in large quantities. The effluents from different industries such as petroleum, leather, paint, electroplating, tannery, textile, and paper, etc. contain large amount of contaminants. Hence, there is a requisite to remove the pollutants from the wastewater (Kiruba et al. 2014; Anitha et al. 2015; Neeraj et al. 2017; Kumar et al. 2018; Joshiba et al. 2019; Renita et al. 2019). A variety of substances in untreated wastewater effluents are known to be harmful to plants and animals, including people, and posture negative effects on the earth. The major contaminants in wastewater effluents are dyes, heavy metals, and hydrocarbons; organic matters are the real contaminants in wastewater that prompt unfavorable impacts to both human health and the earth ecosystem. Several conventional methods such as membrane separation, ultrafiltration, nanofiltration, reverse osmosis, electrodialysis, solvent extraction, sedimentation, chemical precipitation, and ion exchange have already been employed by many researchers. However, these methodologies have several drawbacks such as high installation cost, lower efficiency when contaminants present at lower concentration, generation of toxic sludge, maintenance cost, disposal problems, and timeconsuming process (Kumar et al. 2010). These disadvantages, together with the requirement for effective and economic techniques for the removal of contaminants from wastewaters, have brought about the improvement of elective sorption technology. Nowadays, biosorption have been effectively used for the removal of toxic pollutants from the water ecosystem (Kumar et al. 2014). Biosorption technique has several advantages when compared to other conventional methods such as low cost, higher efficiency, and reuse of biosorbent material.

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Conventional Methods

A few innovations have been created and utilized for the removal of toxic pollutants from effluent water. Each innovation has its own advantages and disadvantages. Biosorption is a physiochemical process that can be considered as a prominent technique for the effective elimination of toxic pollutants from water sources.

9.2.1

Membrane Separation

Membrane processes are physical dispersion processes of particles in water. They work in light of the fact that specific sorts of films enable particles with specific qualities to go through them, while hindering the section of particles that do not have similar attributes. Usage of films for water treatment has advanced utilizing further developed layers produced using new materials and utilized in different setups. An expanding shortage in new water sources fueled a push toward elective assets, for example, seawater (Le and Nunes 2016). Disadvantages: • Membrane fouling • Films must be supplanted all the time • Production of contaminated water (from discharging)

9.2.2

Ultrafiltration

Ultrafiltration is an assortment of membrane filtration in which forces like concentration gradient or pressure force a solute against a semipermeable membrane. Disadvantages: • Does not remove dissolved salts • Not reduce alkalinity • High cost

9.2.3

Nanofiltration

Nanofiltration is used with low total dissolved oxygen water for example fresh groundwater and surface water. Nanofiltration has pore size in the range of 0–10 nm. Convective transport and diffusion mechanism are the basics in the nanofiltration process.

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Disadvantages: • Membrane fouling • High energy consumption • Membranes are sensitive

9.2.4

Reverse Osmosis

Reverse osmosis is used to demineralize or deionize the water by applying pressure through the semipermeable membrane. The connected weight on reverse osmosis must be sufficient with the goal that water can probably defeat the osmotic weight. The pore structure of reverse osmosis layers is a lot more tightly than ultrafiltration, they convert hard water to delicate water, and they are essentially equipped for expelling all particles, microscopic organisms, and organics; it requires less support. Disadvantages: • Higher amount of water discharge • Need of high pressure in large-scale operation

9.2.5

Electrodialysis

Electrodialysis is a membrane procedure, during which particles are transported through semiporous layer, affected by an electric potential. The membranes are cation- or anion-specific, which fundamentally implies that either positive particles or negative particles will move through. Cation-specific membranes are polyelectrolytes with negatively charged issue, which rejects negatively charged particles and enables decidedly charged particles to course through. Disadvantages: • Not efficient for organic and colloids • Depends of feed water chemistry

9.2.6

Solvent Extraction

Solvent extraction, also called liquid-liquid extraction and apportioning, is a strategy to separate compounds dependent on their relative solubilities in two diverse immiscible fluids. Liquid-liquid extraction is an extraction of a substance from one fluid into another fluid stage. The most common utilization of the conveyance standard is in the extraction of substances by solvents, which are frequently utilized

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in a research center or in enormous scale fabricating. Organic compounds are commonly significantly more dissolvable in natural solvents, similar to benzene, chloroform, and ether, than in water, and these solvents are immiscible with water. Organic compounds are then effectively isolated from the blend with inorganic mixes in the aqueous medium by including benzene, chloroform, etc. Disadvantages: • High cost

9.2.7

Chemical Precipitation

Chemical precipitation is utilized for removal of toxic metals from wastewater. To change over this metal into strong particles, a precipitation reagent is included. A substance response happens where dissolved metals form solid particles. Disadvantages: • High cost • High power consumption • Large amount of waste generation

9.2.8

Ion Exchange

The procedure by which a blend of comparable charged ions can be isolated by utilizing an ion exchange resin. Ion exchange resin exchanges particles as indicated by their relative affinities. There is a reversible trade or comparable charged particles. Mostly comparable charged particles like cations or anions can be isolated by this strategy. Disadvantages: • Properties and nature • Foul the resin

9.3

Biosorption

In perspective on the drawbacks related with conventional methodologies for pollutant removal, there is a requirement for option, cost-effective innovations. Nowadays, biosorption technique is considered as financial and effective elective treatment innovations for the treatment of wastewater (Kumar et al. 2011;

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Ponnusamy and Subramaniam 2013; Senthamarai et al. 2013; Gunasundari and Kumar 2017; Kaveeshwar et al. 2018; Hemavathy et al. 2019; Vardhan et al. 2019). Introduction of the “biosorption” term started many dialogs concerning the definition for this procedure, in principle, of the enormous number of instruments considered during dynamic acting of this procedure, components which depend particularly of the microorganism type, condition type containing the toxin together with the elements, which can influence the conduct and impact certainly the microorganism activity (Adeniyi and Ighalo 2019; Pradhan et al. 2019; Wang et al. 2018). It is a reversible fast procedure engaged with official of contaminations onto the active sites present on the biosorbent in fluid arrangements by methods for different associations. The benefits of this procedure incorporate are basic task, no extra supplement prerequisite, low amount of waste generation, cost-effective, high productivity, and recovery of biosorbent, which are generally the significant confinements for the vast majority of the conventional methods. Biosorption can evacuate contaminants even in lower concentration and has extraordinary importance regarding metal removal attributable to toxic level at concentration levels. • Low cost: the expense of the biosorbent is low since they regularly are produced from abundantly available or waste material. • Metal/dye specific: the contaminants sorbing execution of various kinds of biomass can be more or less selective on various metals. This relies upon different factors, for example, kind of biomass, sort of biomass preparation, and physicosynthetic treatment. • Regenerative: biosorbents can be reused, after the contaminants are reused. • No sludge generation: no auxiliary issues with sludge occur with biosorption, similar to the case with numerous different systems, for instance, precipitation. • Metal recovery possible: in the case of metals, it very well may be recouped in the wake of being sorbed from the arrangement. • Competitive performance: biosorption is equipped for an exhibition comparable to the most comparative method. The sort of utilized biomass, living or dead, directs the kind of delivered bioprocess. Therefore, the dead biomass decides a biosorption procedure described by reversibility, lack of involvement, accomplishment by exemplary components of adsorption with moderately high rates, and without age of secondary toxic items, which does not require expansion of supplements and happens in a single stage not constrained by the creature digestion. Rather, the living biomass initiates a bioaccumulation procedure, process actuated in two phases, somewhat reversible, which is portrayed, in head, by an intracellular collection of toxin, which surpasses the biosorption state on the grounds that the substance species is for the most part utilized by the cell with arrangement of potential toxic compounds. The pace of this procedure is lower and constrained by the cell metabolism. In sorption process, the decision of biosorbent depends on necessities concerning high selectivity, great kinetic features, physical-chemical dependability, large capacity of sorption, mechanical quality, availability, simple recovery, and low cost.

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Mechanism

The key components controlling and describing these instruments are (Wang and Can 2006): • The chemical, stereochemical, and coordination qualities of toxic pollutants like subatomic weight, ionic range, and oxidation condition of the focused on contaminants • Properties of the biosorbent, that is, the structure and nature (live or dead biomass) • Sort of the coupling site (natural ligand) • Accessibility of the binding destinations The biosorption mechanism can be affected by a few factors, for example, kind of biosorbent (living or nonliving), sorts of biomaterials, interaction between the pollutants and the adsorbents, contaminants concentration, pH, contact time, adsorbent measurement, and temperature. The mechanism of biosorption procedure may be characterized by reliance on the cell’s digestion, which is called metabolism subordinate and non-metabolism subordinate. Metabolism subordinate is the procedure of pollutant take-up by physiochemical connection between the solid phase (adsorbent; biosorbent) and liquid phase (dissolvable; water). The liquid phase, which contains the suspended species (adsorbate; metal particles). The cell mass of various microorganism comprises of polysaccharides, proteins, and also lipids, which offers various dynamic destinations, prepared to do binding contaminants. This cell wall composition significantly comprises of a few potential contaminants binding groups, for example, carboxyl, hydroxyl, aldehydes, ketones, liquor, amyl, and ether. These functional groups have exceptional surface science, which attracts the contaminants from the liquid phase. This kind of biosorption is called as non-metabolism subordinate. Assortment of microbial materials, for example, bacterial, fungal, algal mass, and yeast have been utilized as the biosorbent material for the adsorption process. Microbial aggregates have complex surface structure, shaped from miniaturized scale and meso-organisms, and have various shapes, for example, filamentous, oval, round, sheet, and unpredictable. Under the ideal conditions, the microbial aggregates will embrace the thick structure and have some unique highlights, for example, grip and flocculation, which upgrade the microbial adsorbent movement to adsorb the pollutants from the fluid arrangement. The sorts and amount of contaminants binding on biosorbent material will contrast for every microorganism in view of contrasts in the cell wall composition among the diverse functional groups of microorganisms. Cell mass of microbes significantly comprises of peptidoglycan, which is made up of a polysaccharide spine comprising of N-acetyl glucosamine and N-acetyl muramic acid deposits. Peptidoglycan is majorly responsible of the solidness of the bacterial cell wall and for the assurance of the cell shape. Regularly, fungi cell wall is adaptable which is made up of glucosamine polymer chitin, chitosan, inorganic particles, proteins, and polysaccharides. Numerous negatively charged functional groups were available on

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the cell wall of fungi species, for example, carboxyl, hydroxyl, sulfate, phosphate, and amino groups. Strangely, these negative groups potentially attract the positive contaminants, which was available in the water environment.

9.4 9.4.1

Biosorbent Material Microbial Adsorbent

Microbial adsorbent in a liquid phase condition plays a significant job in the removal of contaminations from wastewater. The contaminant removal proficiency of biosorption innovation can be improved by utilizing the microbial biomass. Because of the permeable structure of its surface, adsorb the contaminants from the wastewater can without much of a stretch onto the outside of the microorganism (Saravanan et al. 2017; Chergui et al. 2007). The bacterial cell wall is the central part that is presented to contaminants, where the solutes can be bound superficially or inside the cell divider structure. Since the strategy for solute take-up by dead/dormant cells is extracellular, the compound functional groups of the cell divider accept basic employments in biosorption (Deepa et al. 2006). As they are negatively charged and richly accessible, carboxyl groups effectively take an interest in the binding of pollutants. In addition, amine groups are very compelling at the removal of pollutants yet in addition adsorb anionic pollutants via electrostatic interaction or hydrogen holding. Gram classification isolates microscopic organisms in two general classifications: Gram positive and Gram negative. Gram negative generally comprises pathogens in spite of the fact that pathogens are likewise revealed in Gram positive. Gram-positive microorganisms are incorporated thick peptidoglycan layer related by amino corrosive scaffolds. For the most part, Gram-positive bacterial cell divider included 90% peptidoglycan. Some teichoic acids are associated with lipids of lipid bilayer encircling lipoteichoic destructive. These lipoteichoic acids are associated with lipids of cytoplasmic film. They set up linkage of peptidoglycan to cytoplasmic film. This results in cross-associating of peptidoglycan forming a system-like structure. These teichoic acids are responsible for negative charge on cell divider as a result of nature of phosphodiester bonds between teichoic destructive monomers. Fungi are likewise considered as financial and eco-friendly biosorbents due to trademark highlights, that is, anything but difficult to develop, high return of biomass, and simplicity of modification. The cell wall of fungi indicates astounding restricting properties due to recognizing highlights like lipids, proteins, polyphosphates, and chitins among various types of organisms. The cell wall of fungi is wealthy in glycoproteins and polysaccharides, which contain different metal-restricting groups like carboxyl, amines, hydroxyl, and phosphates. The fungal organism is utilized in a wide assortment of fermentation processes. Subsequently, they can be effectively created at the mechanical level for biosorption of pollutants from a huge volume of polluted water assets. In view

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of their filamentous nature, they are anything but difficult to separate by methods for simple strategies like filtration.

9.4.2

Agricultural Biomass

The disadvantages of utilizing microorganisms can be overcome by utilizing minimal effort adsorbents. In general, a sorbent can be considered as low cost in the event that it requires low quantity (in case of processing) and is abundantly available in nature or is a side effect or waste material from another industry, which has lost its financial or further handling qualities. There have been a few minimal effort adsorbents that have been utilized for the removal of contaminants. These days, the change of agro-wastes into the minimal effort adsorbents is a helpful elective innovation for the removal of contaminants from the water sources. Agro-waste biomass containing the components, for example, cellulose, lignin, proteins, starch, basic sugars, lipids, hydrocarbons, and water conceivably increase the adsorption capacity of agro biomass for the removal of toxins from the effluent water. Chemical composition (the nearness of lignin, hemicellulose, water hydrocarbons, sugars, lipids, and starch having an assortment of active sites), accessibility and the utilization of agro-wastes is by all accounts a reasonable choice for heavy metal remediation. Agricultural wastes have a few points of interest, for example, minimal effort, sustainable nature, simple accessibility, and higher potential for evacuation different toxins. Cost is a significant parameter for contrasting the sorbent materials. Notwithstanding, cost data is only occasionally announced, and the cost of individual sorbents differs relying upon the level of handling required and local accessibility.

9.4.3

Algal Biomass

The usage of algae as a biosorbent has gotten center as a result of the uncommon essential of enhancements, high sorption limit, high surface area, less volume of muck to be masterminded, and the potential for metal recovery. They are considered as both economical and eco-friendly answers for wastewater treatment (Donmez et al. 1999). The cell wall of algae on a very basic level contains three sections: cellulose, polysaccharide, and alginic acid with high substance of carboxyl groups that are engaged in with the methodology of the biosorption of metals (Murphy et al. 2008). Red algae have received thought for biosorption due to the proximity of sulfated polysaccharide made of galactans (having high substance of hydroxyl and carboxyl social occasions). Algae contain cellulose with an abnormal state of protein bound to polysaccharides, which contain various functional groups like hydroxyl, amino, carboxyl, and sulfate.

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Desorption and Regeneration

Biosorption is a procedure of treating contaminants bearing answers which makes it contaminant free. In any case, it is likewise vital to have the option to recover the biosorbent. This is possible just with the guide of appropriate pollutants, which normally brings about a concentrated contamination agreement. Thusly, the general achievement of a biosorption methodology is to think the solute, i.e., sorption sought after by desorption. Desorption is of most extraordinary hugeness when the biomass plan/age is costly, as it is possible to reduce the method cost and besides the dependence of the system on a consistent supply of biosorbent. An effective desorption procedure requires the best possible determination of elutants, which unequivocally relies upon the sort of biosorbent and the system of biosorption. Additionally, the elutant must be: • Less exorbitant • Not harmful to the environment • Successful Despite the fact that some chemical agents perform well in desorption, they might be unfavorable to the biosorbent. Bacterial biomasses present issues during desorption because of their minuscule structure. They will in general be influenced by the nearness of both solid acidic and antacid conditions, which are regularly utilized during desorption forms. Notwithstanding, the majority of the distributed work has expected to assess the coupling capacity of biomass and the parameters influencing the procedure. Less consideration has been paid to the recovery capacity of the biosorbent, which frequently chooses the mechanical materialness of a procedure. Hence, biosorption thinks about ought to underscore the likelihood of biomass recovery to improve the procedure reasonability.

9.6

Future Perspective and Conclusion

Conventional water medications may not totally remove pollutants. Consequently, biosorption might be coordinated downstream of other ordinary water medications. This is particularly significant because of toxins whose impacts are felt even at ppb levels. The proficiency for the removal of explicit pollutants is obstructed by the nearness of different contaminants. This might be significant during the recuperation of explicit metals of financial worth. In such manner, biosorption might be connected to squanders and effluents before it enters the sewage or normal release streams like waterways, oceans, etc. Nonetheless, with the point of treating profluent/remediating water assets of every single/most contaminant, it might be a bit of leeway to have all toxins (metal or contaminants) evacuated all the while utilizing a non-explicit/nonspecific biosorbent and lessening the quantity of tasks/steps. Various biosorbents of various specificities/selectivities can

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likewise be utilized. Nature gives an assorted variety of biomass shifting in restricting explicitness, effectiveness, and roughness. This incorporates the expense and solidness of the biosorbent (layer), the decrease in restricting destinations (fouling), and poor understanding and general hesitance to embrace innovations and so on. Thus, given its eco-accommodating nature and different benefits, it will discover its place as a normal water treatment process.

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kinetics and mechanism. Desalin Water Treat 52:2620–2633. https://doi.org/10.1080/ 19443994.2013.792016 Kumar JA, Amarnath DJ, Kumar PS, Kaushik CS, Varghese ME, Saravanan A (2018) Mass transfer and thermodynamic analysis on the removal of naphthalene from aqueous solution using oleic acid modified palm shell activated carbon. Desalin Water Treat 106:238–250. https://doi.org/10.5004/dwt.2018.22066 Le NL, Nunes SP (2016) Materials and membrane technologies for water and energy sustainability. Sustain Mater Technol 7:1–28. https://doi.org/10.1016/j.susmat.2016.02.001 Murphy V, Hughes H, Mcloughlin P (2008) Comparative study of chromium biosorption by red, green and brown seeweed biomass. Chemosphere 70:1128–1134. https://doi.org/10.1016/j. chemosphere.2007.08.015 Neeraj G, Krishnan S, Kumar PS, Shriaishvarya KR, Kumar VV (2017) Performance study on sequestration of copper ions from contaminated water using newly synthesized high effective chitosan coated magnetic nanoparticles. J Mol Liq 214:335–346. https://doi.org/10.1016/j. molliq.2015.11.051 Ponnusamy SK, Subramaniam R (2013) Process optimization studies of Congo red dye adsorption onto cashew nut shell using response surface methodology. Int J Ind Chem 4:17. https://doi.org/ 10.1186/2228-5547-4-17 Pradhan D, Sukla LB, Mishra BB, Devi N (2019) Biosorption for removal of hexavalent chromium using microalgae Scenedesmus sp. J Clean Prod 209:617–629. https://doi.org/10.1016/j.jclepro. 2018.10.288 Renita AA, Kumar PS, Jabasingh SA (2019) Redemption of acid fuchsin dye from wastewater using de-oiled biomass: kinetics and isotherm analysis. Bioresour Technol Rep 7:100300. https://doi.org/10.1016/j.biteb.2019.100300 Saravanan A, Kumar PS, Yashwanthraj M (2017) Sequestration of toxic Cr(VI) ions from industrial wastewater using waste biomass: a review. Desalin Water Treat 68:245–266. https://doi.org/10. 5004/dwt.2017.20322 Senthamarai C, Kumar PS, Priyadharshini M, Vijayalakshmi P, Kumar VV, Baskaralingam P, Thiruvengadaravi KV, Sivanesan S (2013) Adsorption behavior of methylene blue dye onto surface modified Strychnos potatorum seeds. Environ Prog Sustain Energy 32:624–632. https:// doi.org/10.1002/ep.11673 Vardhan KH, Kumar PS, Panda RC (2019) A review on heavy metal pollution, toxicity and remedial measures: current trends and future perspectives. J Mol Liq 290:111197. https://doi. org/10.1016/j.molliq.2019.111197 Wang J, Can C (2006) Biosorption of heavy metals by Saccharomyces cerevisiae: a review. Biotechnol Adv 24:427–451. https://doi.org/10.1016/j.biotechadv.2006.03.001 Wang L, Liu X, Lee D-J, Tay J-H, Zhang Y, Wan C-L, Chen X-F (2018) Recent advances on biosorption by aerobic granular sludge. J Hazard Mater 357:253–270. https://doi.org/10.1016/j. jhazmat.2018.06.010

Chapter 10

Microbial Exopolymeric Substances for Metal Removal Caleb Cheah and Adeline Su Yien Ting

Contents 10.1 10.2

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxic Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.2.1 Metals in the Industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.2.2 Metal Pollution and Health Hazards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3 Methods for Metal Removal in Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.1 Conventional Metal Removal Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.3.2 Bio-based Metal Removal Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4 Exopolymeric Substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4.2 Composition of Exopolymeric Substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4.3 Metal Biosorption by Exopolymeric Substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4.4 Mechanisms of Metal Biosorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4.5 Cultivation and Extraction of Exopolymeric Substances . . . . . . . . . . . . . . . . . . . . . . 10.4.6 Advantages and Limitations of Exopolymeric Substances as Biosorbents . . . 10.5 Innovative Improvements and Future Trends on the Use of Exopolymeric Substances 10.5.1 Immobilization of Exopolymeric Substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5.2 Media Manipulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5.3 Future Trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.6 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract The increased usage of metals in industries poses a serious threat to water pollution as metals leak from factories into nearby water bodies. Although there are existing methods to remove metals from wastewater, these methods are costly and produce toxic by-products. Thus, bioremediation is viewed as an alternative method. This chapter reviews the usage of exopolymeric substances secreted by microbial cells as a potential metal biosorbent. These exopolymeric substances comprise primarily of macromolecules which contain functional groups for metal biosorption. C. Cheah · A. S. Y. Ting (*) School of Science, Monash University Malaysia, Bandar Sunway, Selangor Darul Ehsan, Malaysia e-mail: [email protected]; [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 225 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_10

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Although innovations such as immobilization have been carried out to improve the metal biosorption capability of exopolymeric substances, several new potential methods such as pretreatments have been suggested in this chapter. The use of exopolymeric substances is one of the green approaches in managing metal pollutants in a sustainable manner. Keywords Bioremediation · Biosorption · Exopolymeric substances · Metal removal · Sustainable approach

10.1

Introduction

Metal pollution in the environment is attributed to rampant metal usage in various industries (Ahmad and Danish 2018). Conventional methods to remove metals from environmental wastewater, e.g., reverse osmosis and chemical precipitation, are costly and tend to produce toxic by-products (Parmar and Thakur 2013). To address these limitations, alternatives are sought, in which bioremediation approach is explored as the way forward. The bioremediation approach includes the utilization of bio-based materials such as agricultural waste or microbial cells to adsorb metal pollutants in environmental wastewater. Metal removal through the use of bio-based materials is typically via the biosorption or biodegradation process, which is aligned to the green approach advocated globally (Crini and Lichtfouse 2018). Of the many approaches in bioremediation, the application of non-cell forms of microorganisms is considered an attractive alternative. In this approach, enzymes or other biomolecules are harnessed and used instead of the microbial cells. One such important molecule is the exopolymeric substances that are produced by microbial cells. In nature, microbial cells typically secrete exopolymeric substances extracellularly as a response to environmental stress such as dehydration or pollution (Sivakumar et al. 2012). In laboratory settings, microbial cells can be stimulated to produce exopolymeric substances when supplied with optimal nutrients (Mahapatra and Banerjee 2013). Exopolymeric substances consist of a mixture of inorganic compounds, nucleic acids, proteins, lipids, and carbohydrates (Jia et al. 2017). Due to the composition of exopolymeric substances, there are many functional groups present, such as carbonyl, carboxyl, hydroxyl, phosphoryl, sulfhydryl, and amino groups (More et al. 2014). The presence of these functional groups allows the adsorption of metal ions onto the exopolymeric substances. There are several advantages of using exopolymeric substances as a metal biosorbent, which includes cost-effectiveness and nontoxic by-products (Crini and Lichtfouse 2018). In this review, a brief introduction on metals is provided, followed by the use of exopolymeric substances as a potential solution for metal removal in wastewater. The metal biosorption capabilities of exopolymeric substances and its mechanisms are also discussed. Lastly, the innovations carried out to enhance exopolymeric

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substances for metal removal and their possible future trends for utilization in various industries are presented.

10.2

Toxic Metals

10.2.1 Metals in the Industry Metals are categorized as a group of elements in the periodic table and includes transition metals, post-transition metals, lanthanides, actinides, alkaline earth metals, and alkali metals (Crini and Lichtfouse 2018). Generally, metals are able to conduct electricity and heat and are typically ductile or malleable (Crini and Lichtfouse 2018). As such, metals are utilized in various industries for the manufacturing of various products, ranging from simple batteries to specialized equipment such as car radiators (Farooq et al. 2010). Furthermore, metals can be combined with each other or other elements to form alloys that can be further utilized in industries such as the manufacturing of stainless steel (Farooq et al. 2010). Table 10.1 summarizes the types of metals and their application in various industries. With the onset of industrialization, there is an increase in the demand for metal products, which inevitably leads to increased metal mining and processing (Ahmad and Danish 2018). This results in the leaching of metals into the environment, which has implications to human health and the environment. Although living organisms utilize metals as well, these are in trace amounts. Therefore, high concentrations of metal will lead to metal toxicity (Jaishankar et al. 2014).

10.2.2 Metal Pollution and Health Hazards Metals are leached into the environment from extensive mining processes and industrial effluents (Ahmad and Danish 2018). Often, water and soil environments are both affected, particularly in sites with close proximity to the activities. A study on 72 mining areas in China by Li et al. (2014) validated that metal pollution levels in neighboring soils often exceed the permissible levels by the Environmental Protection Agency. The metal pollutants located in these neighboring soils will leach easily into nearby water systems due to rainfall (Sud et al. 2008). Unlike organic pollutants, metal pollutants are resistant to biodegradation and can accumulate to high concentrations over time in water bodies (Ahmad and Danish 2018). Some metals such as mercury (Hg) can bioaccumulate in lipid layers of aquatic organisms and, through time, biomagnify in these organisms. This impacts the usual balance of metal concentrations in the organism, leading to metal poisoning in aquatic animals and plants. In certain areas, metal contamination may occur in water used as a drinking source. Aquatic plants and animals, which are contaminated

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Table 10.1 The different metals and their application in various industries Metals Cadmium (Cd) Chromium (Cr) Cobalt (Co) Copper (Cu)

Industry Battery Plastic Leather

Usage Cadmium is used in conjunction with nickel to produce portable and rechargeable nickel-cadmium batteries Cadmium pigments are used as coloring agents in plastic products Chromium is used in the coloring of leather products

Airline

Cobalt is used in alloys to construct jet turbines

Automotive

Copper alloys are used in manufacturing radiators, gears, and brake lining Copper is used in the manufacturing of copper wire, cables, and connectors Copper is used to coat containers, pipe, and pumps due to its high corrosion resistance Lead is used as electrodes in lead-acid battery production Lead is used to manufacture lead plates and other alloys which are used in ship parts Mercury switches are often used in hood lights and anti-lock brakes Methyl mercury is a key component in various pesticides and fungicides Nickel is used to form stainless steel alloy used in making culinary equipment and jewelry Zinc is used to coat the surface of structures to prevent atmospheric corrosion Zinc is used in fertilizers as it is essential in plants for chlorophyll production

Electronics Petrochemical Lead (Pb)

Battery Ship building

Mercury (Hg)

Automobile Pesticides

Nickel (Ni)

Stainless steel

Zinc (Zn)

Construction Fertilizer

Data sourced from Farooq et al. (2010)

by metals, may also end up being ingested by humans, causing metal toxicity (Zakhem and Hafez 2015; Azimi et al. 2017). The mechanism of metal toxicity is attributed to the ability of metal ions to displace other essential cations required by living organisms, such as Na+, Ca2+, and Mg2+ (Jaishankar et al. 2014). The displacement of these cations leads to the interference of various biological processes such as protein folding, enzyme regulation and production, and processes involving intra- and intercellular signaling (Jaishankar et al. 2014). The disruption of these processes in turn leads to various health hazards as summarized in Table 10.2.

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Table 10.2 Health hazards by metal pollutants and their maximum contaminant level Metals Arsenic Cadmium Chromium Cobalt Copper

Maximum contaminant level (mg/l) 0.050 0.010 0.050 0.050 0.250

Lead

0.015

Mercury Nickel Zinc

0.002 0.200 0.800

Health hazards Vascular disease, visceral cancers Kidney damage, lung fibrosis, cancer, weight loss Allergic dermatitis, cancer Diarrhea, coma, sterility Kidney or liver damage, stomach ache, headache, irritation to eyes and nose High blood pressure, kidney damage, muscle pain, cancer, anemia Muscle pain, kidney damage, dermatitis Coughing, nausea, dermatitis, chronic bronchitis Increased thirst, depression, restlessness

Maximum contaminant level indicates the highest contaminant concentration, which is allowed in drinking water as regulated by the Environmental Protection Agency Data sourced from Sud et al. (2008); Nguyen et al. (2013); and Parmar and Thakur (2013)

10.3

Methods for Metal Removal in Wastewater

10.3.1 Conventional Metal Removal Techniques Removal of metals from the environment can be achieved through the following methods: physicochemical and bio-based approach. The physicochemical approach is the conventional approach, utilizing techniques such as chemical coagulation and ion exchange (Table 10.3). These techniques utilize chemicals and membranes to precipitate out or trap metal pollutants present in wastewater. However, these techniques are costly and often create toxic waste products. Thus, the cheaper and more environment-friendly bio-based approach is preferred.

10.3.2 Bio-based Metal Removal Techniques Bio-based removal techniques rely on the use of bio-organic sources, which includes industrial and agricultural wastes as well as microorganisms. With the use of bio-organic sources, metals are typically removed via the biosorption mechanism. Biosorption is a process where metal ions (adsorbate) attach to a biomass (biosorbent), which results in efficient metal removal (Ngah and Hanafiah 2008). The first study on metal biosorption was in 1902, in which two strains of fungi (Tilletia tritici and Ustilago crameri) were used to remove Cu in aqueous solutions (Park et al. 2010). However, it was only in the last half of the century that research in biosorption began to truly expand (Crini and Lichtfouse 2018). The two main groups of biosorbents, waste products and microorganisms, are summarized in Table 10.4. Of the two groups, microorganisms present a more attractive alternative, as they can

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Table 10.3 Physicochemical techniques for metal removal from wastewaters Technique Chemical precipitation

Chemical coagulation

Electrodialysis

Electrochemical methods Ion exchange

Reverse osmosis

Description Addition of chemicals such as iron salts or lime to metal solutions to precipitate out metal hydroxides M2++ 2OH
¼ M(OH)
Similar to chemical precipitation, whereby coagulants such as ferric salts are added so that pollutants form sediments that can be easily removed Metal ions from the solution are separated out when filtered through an ion exchange membrane via application of an electrical current Combination of electrodialysis and chemical coagulation techniques Metal ions in water will be exchanged with an ion of similar charge which is pre-attached to an immobile surface Water is forced to move through a semipermeable membrane, which will block the passage of metal ions

Limitations Generates a large amount of toxic sludge which creates a disposal problem Expensive and consumes large amount of chemicals

High operational cost due to energy consumption and membrane fouling Extremely high operational cost Expensive and has low metal removal efficiency High operational cost

Data sourced from Nguyen et al. (2013) and Parmar and Thakur (2013) Table 10.4 Types of biosorbents commonly utilized for metal biosorption Biosorbents Waste products

Category Industrial waste Agricultural waste Others

Microorganisms

Fungi

Bacteria

Algae

Examples Sludge (sewage sludge) Fruit waste (watermelon rinds) Plantation waste (rice husk) Chitin-based materials (crab shells) Mushrooms (Lepiota hystrix) Yeast (Saccharomyces cerevisiae) Molds (Aspergillus niger) Gram-positive bacteria (Bacillus cereus) Gram-negative bacteria (Pseudomonas fluorescens) Cyanobacteria (Cyanospira capsulata) Seaweed (Sargassum baccularia)

References Phuengprasop et al. (2011) Lakshmipathy and Sarada (2013) Sheveleva et al. (2009) Vijayaraghavan et al. (2011) Kariuki et al. (2017) Zhao et al. (2015) Vale et al. (2016) Costa and Duta (2001) Hussein et al. (2004) De Philippis et al. (2003) Hashim and Chu (2004)

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be easily cultivated, thus minimizing the demands for labor and resources. Furthermore, these bio-based resources are deemed more beneficial than the use of commercially available activated carbon adsorbents, as activated carbons are extremely expensive and require regeneration after each use (Ngah and Hanafiah 2008). Therefore, the use of bio-based materials, particularly with the recycling of agricultural wastes and the upscaling of microorganisms, is expected to reduce costs. Wastes are a form of organic source that has potential for metal removal. Industries or sewage treatment plants often produce a semisolid viscous waste known as sludge. Sludge comprises of mixtures of proteins, fats, cellulose, silica, calcium oxide, magnesium oxide, phosphoric acid, and alumina (Smith et al. 2009). The presence of these components presents a variety of functional groups that can adsorb metal ions (Smith et al. 2009). Phuengprasop et al. (2011) proved that sewage sludge was capable of adsorbing metals and reported a biosorption capability of 7.8, 17.3, 42.4, and 14.7 mg/g for Ni, Cu, Pb, and Cd, respectively. Agricultural wastes are another source of biosorbents for metal removal. Unwanted or unusable materials produced from agricultural operations, which ranges from fruit waste to shells and seeds, all have potential for application for metal removal. Agricultural waste consists primarily of celluloses, lignins, and hemicelluloses (Demirbas 2009). They are made up of various sugar chains and aromatic groups, thus presenting a large number of functional groups which include amine, carbonyl, methoxyl, hydroxyl, and carboxyl groups that can adsorb metal ions (Demirbas 2009). For instance, Lakshmipathy and Sarada (2013) conducted a biosorption test on watermelon rinds, the inedible part of watermelons. They discovered that the rinds were capable of removing up to 60% of Co in 20 ml metal solutions (50 mg/l). In addition, Sheveleva et al. (2009) discovered that rice husks effectively removed Fe, Cu, Cd, and Pb. Other than agricultural waste, wastes from animals such as seafood wastes (shells of crabs, shrimp) can also be used for metal removal. This is attributed to the high chitin content in the shells (Dotto et al. 2017). The linear biopolymer chitin consists of 2–amino–2–deoxy–D–glucopyranose and 2–acetamido–2–deoxy–D– glucopyranose units (Crini and Lichtfouse 2018). The structure of this biopolymer provides several unique functional groups (e.g., ether and acetamido groups), which metal ions may attach to (Dotto et al. 2017). Vijayaraghavan et al. (2011) used crab (Portunus sanguinolentus) shells to remove Zn and Mn at maximum biosorption capability of 123.7 and 69.9 mg/g, respectively. These industrial and agricultural wastes offer not only a cheap and environmentfriendly solution to remove metal pollutants, but utilizing them as biosorbents also reduces the amount of wastes discarded in landfills around the world (Bhatnagar et al. 2015). This in turn reduces the production of methane gas in landfills and slows down the threat of global warming (Bhatnagar et al. 2015). However, wastes tend to require further processing to increase its effectiveness as biosorbents (Farooq et al. 2010). There are also better uses of agricultural waste such as converting it to animal feed or biofuel (Yevich and Logan 2003). As such, there is an inevitable shift to explore microbial cells as biosorbents, as microbial cells can be easily cultivated in abundance, thus minimizing the demands for labor and resources.

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Ever since Antonie van Leeuwenhoek discovered the existence of unseen miniscule organisms in 1676 (O’Malley 2009), there have been extensive researches on microbial cells over the years. Microbial cells have been proven to be involved in various aspects of the environment and in organisms, such as recycling key nutrients in the carbon and nitrogen cycle, decomposing organic matter, or causing infectious diseases (Maier et al. 2009). Although microbial cells were known to be able to break down or adsorb environmental pollutants since the 1800s (Gupta and Diwan 2017), it was only during the 1970s that bioremediation was truly put into practice in real-world applications. Early bioremediation applications focused on cleaning up oil spills by utilizing a mixture of hydrocarbon degrading bacteria such as Halobacterium spp., Desulfosarcina spp., and Bacillus spp. (Rekadwad and Khobragade 2017). Since then, the field of bioremediation has expanded to include the removal of dyes and metals by fungal, bacterial, and algae cells (Crini and Lichtfouse 2018). Fungal cells are effective biosorbents as their cell walls are made up of complex structures consisting primarily of proteins, polysaccharides, lipids, pigments, and chitin (Wang and Chen 2009). Biosorption is demonstrated in macrofungi, unicellular fungi, and filamentous fungi. Mushrooms, which are a type of macro-fungi, were utilized by Kariuki and his colleagues in a biosorption study. They reported that Lepiota hystrix was capable of removing 3.8 and 8.5 mg/g of Cu and Pb, respectively (Kariuki et al. 2017). Besides that, Saccharomyces cerevisiae, a well-known yeast species commonly utilized in the brewing and baking industry, removed approximately 93% of Ag with the application of only 0.1 g of the unicellular yeast (Zhao et al. 2015). Filamentous fungi or more commonly known as molds have also been proven to be effective biosorbents. Vale et al. (2016), in their study on Aspergillus niger, revealed the mold had an adsorption capacity of 4.997 and 3.833 mg/g for Cr and Zn, respectively. Bacteria, a class of single cell microorganisms with simple internal structures, are also effective metal biosorbents. Bacteria are classified into Gram-positive or Gramnegative groups based on composition of their cell wall (Seltmann and Holst 2013). Gram-negative bacteria have an outer membrane rich in lipopolysaccharides as well as a thin inner layer of peptidoglycan, while Gram-positive bacteria consists of several layers of peptidoglycan but does not have an outer membrane (Seltmann and Holst 2013). The cell wall composition has a major influence in the metal adsorption capability of bacteria. Costa and Duta (2001) proved that several types of Gram-positive bacteria from the genus Bacillus (B. cereus, B. sphaericus, and B. subtilis) were effective in adsorbing Cu, Pb, Zn, and Cd. On the other hand, Hussein et al. (2004) used different strains of Gram-negative bacteria from the genus Pseudomonas (P. fluorescens and P. putida) and reported removal rates of 93 and 38% for Cu and Cr, respectively. The metal biosorption capability of bacteria is not limited to just Gram-positive and Gram-negative bacteria, but it has also been demonstrated in cyanobacteria. Cyanobacteria are a class of bacteria, which produce oxygen and obtain energy through photosynthesis. In a study comparing two cyanobacteria (C. capsulata and Nostoc PCC7936), it was discovered that

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C. capsulata was more effective at removing Cu as it displayed a higher biosorption capacity of 96 mg/g (De Philippis et al. 2003). Algae are a diverse group of aquatic organisms capable of conducting photosynthesis. They have also been found to be capable of metal biosorption. Their cell wall contains polysaccharides and glycoproteins such as xylan, mannan, and alginic acid (Abbas et al. 2014). As such, metal ions bind easily to the negatively charged surfaces of algae, enabling biosorption of metals such as Cd when brown and red seaweed (S. baccularia and Gracilaria salicornia) were used as biosorbents (Hashim and Chu 2004). The utilization of microbial cells for metal biosorption has several advantages. Firstly, they can be easily cultivated in abundance, thus making it a cost-effective technique (Crini and Lichtfouse 2018). Furthermore, this technique does not produce toxic by-products; hence further treatment of toxic sludge is not required, making the process environment-friendly (Crini and Lichtfouse 2018). Besides that, microbial cells are able to adsorb a wide range of metal pollutants, allowing it to be used for the treatment of various metals (Crini and Lichtfouse 2018). Lastly, microbial cells can be applied for in situ bioremediation treatments as microbial cells can be directly inoculated into polluted wastewater, thus allowing the metal removal process to take place faster compared to any conventional methods (Groudev et al. 2001). However, there are also several challenges in using microbial cells. One of the main challenges is to maintain the viability of live cells when used for environments with high metal concentrations (Ahluwalia and Goyal 2007). Besides that, the introduction of either viable or dead microbial cells into environmental wastewater would cause a potential health threat due to the pathogenicity of microbial cells (Ochman et al. 2000). To address these limitations, researchers explore the potential use of compounds produced from microbial cells for bioremediation. Compounds such as exopolymeric substances are useful as they have been shown to play a major role in the adsorption process of microbial cells (Gupta and Diwan 2017). By extracting these exopolymeric substances and using them as biosorbents independently from its microbial host, it will help overcome the limitations of using microbial cells as biosorbents.

10.4

Exopolymeric Substances

10.4.1 Introduction Exopolymeric substances are typically made up of macromolecules consisting of carbohydrates, lipids, proteins, nucleic acids, and some inorganic compounds (Sheng et al. 2010; Jia et al. 2017). The exopolymeric substances often exist as a matrix surrounding the exterior of microbial cells (Fig. 10.1). Exopolymeric substances may be formed due to the hydrolysis of macromolecules or cellular lysis but are mainly secreted by microbial cells as a response to environmental stress (Sivakumar et al. 2012). Exopolymeric substances plays a vital

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Fig. 10.1 Sketch of spatial distribution of exopolymeric substances matrix surrounding the microbial cell. (Modified after Sheng et al. 2010)

role in microbial cells as they assist in many functions such as adherence to surfaces, formation of protective barriers, water retention, adsorption of organic and inorganic compounds, cell-cell recognition, aggregation of cells, and formation of biofilms (More et al. 2014; Sutherland 2016). Exopolymeric substances are produced extracellularly in two main forms, which are soluble exopolymeric substances (e.g., colloids and slimes) or bound exopolymeric substances (e.g., capsules, sheaths, loosely bound polymer, and condensed gels) (Sheng et al. 2010). Soluble exopolymeric substances are generally weakly bound to cells or dissolved in the surrounding solution, while bound exopolymeric substances remain closely bound to cells. A wide range of microbial cells, which include bacteria, fungi, and algae, produces exopolymeric substances. Among these, exopolymeric substances production by cyanobacteria (Anabaena spp.), Gram-positive (Bacillus spp.), and Gramnegative bacteria (Flavobacterium spp.) is more well studied (Messner 1997; De Philippis et al. 2011). On the contrary, exopolymeric substances production by fungi is less commonly studied. Few species which are known to produce exopolymeric substances include Penicillium spp., Phellinus spp., and Pleurotus spp. (Mahapatra and Banerjee 2013). Similarly, exopolymeric substances production by algae is also not extensively studied, with examples such as Dunaliella salina and Chlamydomonas reinhardtii being reported (Mishra and Jha 2009; Bafana 2013). The majority of current research on exopolymeric substances is focused on bacterial cells. One of the reasons is that bacterial cells have a shorter incubation time and can grow more rapidly (Mahapatra and Banerjee 2013). Furthermore, the optimal conditions required for maximal exopolymeric substances production is easier to replicate for bacterial cells (Mahapatra and Banerjee 2013). Lastly, bacterial cells have a better capability to produce a wider range of exopolymeric substances (i.e., xanthan, gellan, alginate, hyaluronan), that are utilized in different industrial applications (Freitas et al. 2011). The early studies on exopolymeric substances originate from the need to remove or eradicate exopolymeric substances as it aids in biofilm formation, which allowed pathogenic microbial cells to attach to surfaces and contaminate nearby food or

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water (Zottola and Sasahara 1994). It is only in recent decades that studies on exopolymeric substances have explored the use of exopolymeric substances for beneficial applications. The chemical composition and properties of exopolymeric substances were examined, and it allowed researchers to apply exopolymeric substances in the food, pharmaceutical, and cosmetic industries (Freitas et al. 2011). For instance, dextran (exopolymeric substances from Leuconostoc spp.) is used in the pharmaceutical industry as a blood volume expander, while xanthan (exopolymeric substances from Xanthomonas spp.) is used as food additive in the food industry (Freitas et al. 2011). Exopolymeric substances are also studied as a biosorbent, which can be utilized for bioremediation of oil spills, dyes, and metals (Zhang et al. 2009; Gutierrez et al. 2013; More et al. 2014).

10.4.2 Composition of Exopolymeric Substances The exact composition of the exopolymeric substances matrix varies among microbial cells, but generally the primary content is carbohydrate, which constitutes approximately 40–95% of the total composition (More et al. 2014). The sugar molecules of exopolymeric substances are either the neutral sugar molecules such as hexose or negatively charged sugar molecules such as uronic acids (Flemming and Wingender 2010). Uronic acids and other similar compounds such as succinates, pyruvate, and acetate ester determine the charges (neutral, cationic, or anionic) of the exopolymeric substances (Gupta and Diwan 2017). For instance, uronic acid contains a carboxyl functional group (-COOH), which when ionized will release the hydrogen ion, resulting in the negatively charged -COO
group and contributing to the overall charge of the exopolymeric substances. Exopolymeric substances may also exist in the form of either homopolysaccharides or heteropolysaccharides (Gupta and Diwan 2017). Homopolysaccharides are made up of a singular monosaccharide (e.g., L-fructose or D-glucose) (Flemming and Wingender 2010). A common homopolysaccharide produced by bacteria is cellulose, which is used in the formation of biofilms for attachment to surfaces (Sutherland 2016). Cellulose is composed of repeating D-glucose units with β-(1,4)-linkage (Fig. 10.2a). Another example of homopolysaccharide is dextran, which is produced by Streptococcus spp. and Leuconostoc spp. (Leathers 2002; Naessens et al. 2005). Dextran is composed of repeating D-glucose units with α-(1,6)-linkage and also forms branches from α-(1,3)-linkages (Fig. 10.2b). In contrast, heteropolysaccharides are formed by a combination of two or more monosaccharides such as D-galactose, D-glucose, D-mannuronic acid, D-glucuronic acid, L-fructose, and L-guluronic acid (More et al. 2014). P. aeruginosa, a pathogenic bacterium, secretes a heteropolysaccharide known as alginate, which is used in the formation of biofilm (Ryder et al. 2007). Alginate consists of mannuronic acid and guluronic acid units that are joined together by the β-(1,4)-linkage (Fig. 10.2c). Xanthan is another example of heteropolysaccharide and is produced by Xanthomonas campestris and is composed of mostly mannose, glucuronic acid,

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Fig. 10.2 Chemical structure of (a) cellulose (b) dextran (c) alginate. (Modified after More et al. 2014)

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and pyruvate (Silva et al. 2009). Xanthan is commonly utilized as a thickening agent and food additive (Silva et al. 2009). Thus, the presence of these monosaccharide molecules in the exopolymeric substances matrix provides abundant negatively charged carboxyl and carbonyl functional groups, capable of adsorbing metal cations (More et al. 2014). The second most abundant component of the exopolymeric substances matrix are proteins, which constitutes approximately 1–60% of the matrix (More et al. 2014). The majority of proteins exist as enzymes, responsible for the enzymatic breakdown of substrates such as polysaccharides, proteins, and lipids that are trapped in the exopolymeric substances. These enzymes can also degrade the components of the exopolymeric substances matrix when the microbial cell is starving, to release essential nutrients to sustain growth (Flemming and Wingender 2010). Other than enzymes, the exopolymeric substances matrix also contains lectins. Lectins form a linkage between the extracellular surface and the microbial surface, which assist in microbial aggregations (Flemming and Wingender 2010). In addition, a wide variety of glycoproteins (membrane associated, flagella associated, surface layer) have also been discovered in the exopolymeric substances matrix of both Gram-positive (Bacillus spp.) and Gram-negative bacteria (Flavobacterium spp.) (Messner 1997). These glycoproteins are created when a protein molecule attaches to a sugar molecule via covalent bonds. Glycoproteins are responsible for intercellular signaling as well as maintaining the stability of the microbial structure (More et al. 2014). These protein molecules contain functional groups (e.g., amino, thiol, carboxyl, and hydroxyl) that can potentially bind metal ions (More et al. 2014), thus enhancing the biosorption capability of the exopolymeric substances matrix. The exopolymeric substances matrix also consists of small amounts of lipids (1–10%) and nucleic acids (1–10%) (More et al. 2014). Lipids may interact with carbohydrates to form glycolipids, which have similar functions with glycoproteins (More et al. 2014). The interaction of lipids with proteins results in lipopeptides that serves as biosurfactants (Chen et al. 2015). For example, B. subtilis secretes surfactin (a type of surfactant) to decrease surface tension of the bacterial cell, thus improving its ability to attach to surfaces and form colonies (Chen et al. 2015). Nucleic acids in exopolymeric substances generally exist as extracellular DNA (Czaczyk and Myszka 2007). Extracellular DNA is typically coupled with signaling peptides in the exopolymeric substances matrix so that it may be utilized in horizontal gene transfer between microbial cells (Czaczyk and Myszka 2007). Lipid molecules tend to contain carboxyl group in their fatty acid monomers and hydroxyl groups within their glycerol monomers, whereas DNA molecules contain a mixture of functional groups (e.g., carbonyl, sulfhydryl, phosphate, amino, carboxyl, and hydroxyl). These groups further improve the biosorption capability of exopolymeric substances (More et al. 2014). In recent years, the exopolymeric substances matrix has also been discovered to harbor humic substances such as peat, sewage, water sediments, soil, and sludge (Sheng et al. 2010). Microbial cells do not secrete these substances but adsorb them from their surroundings instead (Sheng et al. 2010). The amount (~20%) and composition of humic substances vary greatly between microbes due to their

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different surroundings; thus they would also contain a wide variety of functional groups, which influences the biosorption capability exopolymeric substances matrix (Sheng et al. 2010).

10.4.3 Metal Biosorption by Exopolymeric Substances The many various constituents of the exopolymeric substances matrix contribute to the presence of functional groups present, such as carbonyl, carboxyl, hydroxyl, phosphoryl, sulfhydryl, thiol, and amino groups (More et al. 2014). These functional groups offer cation exchange potential, and thus they allow microbial cells to bind with metal ions. Pradhan et al. (2007) discovered the presence of amino and carboxyl groups in the exopolymeric substances of Microcystis aeruginosa using infrared spectroscopy (IR), which they reported as responsible for the biosorption of Cr, Ni, and Fe. Chojnacka et al. (2005) further validated the role of the functional groups in the exopolymeric substances. In their study, they esterified phosphate, and carboxyl groups located in the exopolymeric substances of Spirulina sp. resulted in a significant decrease of Cr biosorption. In another study, Wei et al. (2011) used Fourier transform-infrared spectroscopy to prove that phosphate and carboxyl groups are responsible for Cd binding in B. subtilis, with shifts in their spectra detected pre- and posttreatment. Although other functional groups such as phosphoryl, sulfhydryl, thiol, and amino groups contribute to metal biosorption; the primary groups responsible for metal binding are carboxyl, carbonyl, and hydroxyl groups (De Philippis et al. 2011; More et al. 2014). It was also revealed that functional groups in exopolymeric substances have a higher preference for metal cations with lower electropositivity (e.g., Cu, Pb, Hg), as these metal cations have a higher affinity to bind to the functional groups (Brown et al. 1999). This is confirmed in a biosorption experiment carried out by Sim and Ting (2017), whereby Pb and Cu metals, which had lower electropositivity, were more easily adsorbed compared to Zn and Cd metals. The composition of the exopolymeric substances matrix for microbial cells varies widely between different microbes. For instance, a pathogenic bacterium such as P. aeruginosa would have more carbohydrate molecules (carboxyl and carbonyl groups) in its exopolymeric substances matrix as it secretes alginate, a heteropolysaccharide to form biofilms (Ryder et al. 2007). In contrast, a fungus known as Ganoderma lucidum would contain more protein molecules (hydroxyl and amino groups) in its exopolymeric substances matrix as it secretes glycoproteins (Mahapatra and Banerjee 2013). As such, it is generally understood that the functional groups of exopolymeric substances involved in metal adsorption is species dependent, as each species and their exopolymeric substances matrix is compositionally different (Pal and Paul 2008; De Philippis et al. 2011; More et al. 2014).

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10.4.4 Mechanisms of Metal Biosorption Metals are removed in microbial cells via active and passive mechanisms. The active mechanisms are bioaccumulation, bioleaching, and reductive biotransformation, all which involves energy-consuming cell activities to overcome metal toxicity (Gupta and Diwan 2017). The only passive mechanism is the biosorption mechanism (Gupta and Diwan 2017). Biosorption occurs on the surface of the cells where metal ions (adsorbate) in a solution binds to the surface of the cells. When exopolymeric substances are used in place of microbial cells, the biosorption process occurs when metal ions attach to the functional groups present among the exopolymeric substances matrix. The exopolymeric substances matrix typically contains an abundance of functional groups, which causes the matrix to be negatively charged; thus it will attract positively charged metal cations (Gupta and Diwan 2017). The attracted metal ions will be bound to the exopolymeric substances matrix via several mechanisms (e.g., complexation, ion exchange, chelation, physisorption, and precipitation) (Fig. 10.3). It is theorized that these mechanisms act simultaneously at different degrees (Sud et al. 2008). Complexation involves the formation of a coordination complex, whereby the complex consists of a central metal ion surrounded by an array of functional group ligands (Srivastava and Goyal 2010). For instance, Fig. 10.3 shows a central zinc metal ion surrounded by hydroxyl group ligands. This complex may be negatively or positively charged or neutral, and the bonding can be either via electrostatic interactions or covalent bonding (Srivastava and Goyal 2010). Although covalent bonding is stronger compared to electrostatic interactions (Srivastava and Goyal 2010), electrostatic interactions are more common in exopolymeric substances biosorption. Chelation is similar to complexation, whereby a metal ion is attached to two or more functional group ligands (Escudero et al. 2018).

Fig. 10.3 The mechanisms utilized in exopolymeric substances metal adsorption. (Modified after Crini and Lichtfouse 2018)

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Ion exchange involves the formation of electrostatic interactions between the negatively charged functional groups in exopolymeric substances and metal cations (Gadd 2009). Ion exchange is theorized to have a major influence in the overall metal removal procedure (Sud et al. 2008). In contrast, physisorption or also known as surface adsorption has been determined to have only minor influence on the overall metal removal process (Sud et al. 2008). Physisorption is a reversible process where weak bonds (van der Waals force) form between the surface of the exopolymeric substances surface and the metal ions (Demirbas 2009). Metal ions bonded via this mechanism can be easily desorbed if the exopolymeric substances are soaked in distilled water (Demirbas 2009). Furthermore, precipitation of metals occurs when the solubility of metal ions reaches its saturation point, thus reacting with hydroxide ions in water to form metal hydroxides (Gadd 2009). The metal hydroxides are then also capable of binding to the surface of exopolymeric substances matrix via van der Waals force (Gadd 2009).

10.4.5 Cultivation and Extraction of Exopolymeric Substances To obtain sufficient exopolymeric substances to perform biosorption experiments or for potential application in industrial settings, researchers require an effective method to cultivate microbial cells and to extract the exopolymeric substances produced.

Cultivation of Exopolymeric Substances The growth conditions of the microbial cells have significant effect on their rate of exopolymeric substances production. Microbial cells cultivated in optimal growth conditions can increase in numbers exponentially and produce abundant exopolymeric substances (Sheng et al. 2010). One of the main factors influencing microbial growth is the nutrient level. Janga et al. (2007) reported that increased nutrient content in culture conditions facilitated the growth of exopolymeric substances production of microbial sludge in a bioreactor. This was observed by Liu et al. (2010) who reported reduced rates of exopolymeric substances production in Chryseobacterium daeguense, when cultivated in conditions low in carbon and nitrogen content. They further concluded that carbon and nitrogen sources are essential in order for microbial cells to synthesize exopolymeric substances. The importance of nitrogen in supporting microbial growth is also highlighted by Mahapatra and Banerjee (2013), whereby they discovered that yeast extract and corn steep powder were the most effective nitrogen supplements for the production of exopolymeric substances in fungal cells (e.g., Alternaria alternata, Cordyceps militaris, Ganoderma applanatum, Rhodotorula glutinis).

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Other than nutrients, growth conditions such as pH, temperature, and oxygen supply also influence the production of exopolymeric substances. Microbial cells such as Bacillus spp. or Escherichia spp. typically produce high amounts of exopolymeric substances when incubated at near neutral pH (pH 5–7) and at temperature range close to room temperature (25–30  C) (Sheng et al. 2010; More et al. 2014). Few exceptions exist though, such as for psychrophiles (Listeria monocytogenes) and thermophiles (Alicyclobacillus acidocaldarius), whereby these bacteria thrive and have the highest rate of exopolymeric substances production at extreme temperatures (e.g.,
1.5  C and 60  C, respectively) (Georlette et al. 2003; Sardari et al. 2017). Besides that, the growth of microbes and their exopolymeric substances is also dependent on the amount of oxygen in the surrounding environment. For instance, Bayer et al. (1990) reported that P. aeruginosa had increased exopolymeric substances production when oxygen levels were increased from 40 to 80 mm Hg. In contrast, anaerobic bacteria (e.g., Clostridium difficile, Clostridium baratii) would produce more exopolymeric substances in an oxygen-free environment (Meyer-Reil 1994; Donelli et al. 2012). It is also suggested that microbial cells ought to be mixed together during cultivation as their interactions may improve exopolymeric substances production. This is because microbial cells may secrete certain compounds or enzyme that assists the growth of other microbes (Sheng et al. 2010). Zhang et al. (2007) discovered a significantly higher quantity (15 g/l) of pure exopolymeric substances that can be extracted from Pseudomonas sp. and Staphylococcus sp. when they are cultured together compared to if they were cultured individually. Furthermore, a combination of microbial cells from different phylum may lead to the synthesis of novel exopolymeric substances, which has increased metal biosorption capability (Seneviratne et al. 2008). For instance, a combined cultivation of Penicillium frequentans and Bacillus mycoides lead to the synthesis of an improved biofilm with increased bioremediation potential (Seneviratne et al. 2008). Overall, it is generally understood that the optimal growth conditions vary for different microbial species, and researchers will need to optimize these conditions to maximize their production of exopolymeric substances (Sheng et al. 2010).

Extraction Methods of Exopolymeric Substances There are two main methods to extract exopolymeric substances bound to microbial cells, which are via physical or chemical means. Physical methods include heating, centrifugation, or sonication, whereas chemical methods include treatment with acid, alkali, enzymes, alcohol, and organic solvents such as ethylenediaminetetraacetic acid (Sheng et al. 2010). Physical methods generally create pressure which causes the exopolymeric substances to detach from the microbial cells, whereas chemical methods disrupt the binding interaction between exopolymeric substances and the microbial cell surface, causing the exopolymeric substances to dissolve in water (Sheng et al. 2010). Both extraction methods are, however, found to cause disruption to the components of exopolymeric substances (More et al. 2014). For instance,

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heating or utilizing chemicals such as ethylenediaminetetraacetic acid may induce changes in the functional groups present in exopolymeric substances, thus reducing its biosorption capability (More et al. 2014). Currently, most researchers favor a mixture of these two methods, whereby chemicals such as ethylenediaminetetraacetic acid are used to increase the solubility of exopolymeric substances in the surrounding solution, followed by centrifugation (physical) to fully separate the microbial cells and their exopolymeric substances (Sheng et al. 2010; Pan et al. 2010). This is proven by Zhang et al. (1999), which reported that centrifugation with formaldehyde was the most effective extraction method among five different extraction methods to obtain exopolymeric substances from microbial sludge. These methods include regular centrifugation, ethylenediaminetetraacetic acid, ultracentrifugation, steaming, and centrifugation with formaldehyde. This is further supported by another study carried out by Sheng et al. (2005), in which ethylenediaminetetraacetic acid with centrifugation was reported as the most effective method to extract exopolymeric substances from Rhodopseudomonas acidophila compared to NaOH, H2SO4, and heating/ centrifugation. Overall, a combination of chemical and physical (centrifugation) method is preferred to maximize the extraction yield and purity of exopolymeric substances. However, the specific type of chemicals (e.g., NaOH, H2SO4, ethylenediaminetetraacetic acid) is dependent on the type of microbial cells being utilized, as microbial cells which have tightly bound exopolymeric substances have to undergo harsher chemical treatments (e.g., strong H2SO4) (Sheng et al. 2010).

10.4.6 Advantages and Limitations of Exopolymeric Substances as Biosorbents The utilization of exopolymeric substances for metal biosorption has several advantages. Firstly, microbes producing exopolymeric substances can be easily cultivated; thus it is considered to be cost-effective (Crini and Lichtfouse 2018). Furthermore, biosorption of metals onto exopolymeric substances does not produce toxic by-products, which makes the process environment-friendly (Crini and Lichtfouse 2018). In addition, the sensitivity of exopolymeric substances is high, allowing metals in low concentrations in aqueous solutions to be removed (Gupta and Diwan 2017). Exopolymeric substances can also be reused multiple times after metals are desorbed by washing with weak acid or distilled water (Gupta and Diwan 2017). However, there are also several limitations in utilizing exopolymeric substances as a biosorbent. Although exopolymeric substances show effective metal biosorption potential in lab tests, their large-scale application is hindered by the lack of optimal adsorption parameters in actual environmental settings (Gupta and Diwan 2017). The functional groups of exopolymeric substances are also not specific to metals;

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thus they may bind various metal ions and saturate the exopolymeric substances, which prevent the binding of the targeted toxic metal ions (Gupta and Diwan 2017). Thus, based on the understanding of these limitations, researchers have made several innovative improvements on exopolymeric substances to potentially enhance its biosorption capability in industrial settings.

10.5

Innovative Improvements and Future Trends on the Use of Exopolymeric Substances

10.5.1 Immobilization of Exopolymeric Substances Immobilization is a technique that fixates a particular biomolecule onto the surface of an object to increase its efficacy and stability (Datta et al. 2013). This technique is typically applied for enzymes in industrial settings (e.g., immobilization of lipase for biosynthesis of polyester) (Datta et al. 2013). In recent years, microbial cells have also been immobilized. For instance, BIOCLAIM™, which is made up by Bacillus spp., was immobilized in alginate beads, and AlgaSORB™, an algal biosorbent from Chlorella vulgaris, is immobilized in silica gel (Park et al. 2010; De Philippis et al. 2011). Immobilization of microbial cells aids in their enzymatic reactions and speeds up the desired industrial process. On the contrary, the immobilization of exopolymeric substances has not been extensively explored and no commercialized product of immobilized exopolymeric substances is currently available. In the few existing laboratory tests, immobilization of exopolymeric substances is typically achieved via encapsulation or cross-linking (De Philippis et al. 2011). In encapsulation, exopolymeric substances molecules are caged via noncovalent or covalent bonding within fibers or gels (Datta et al. 2013). In contrast, cross-linking involves the formation of covalent bonds between exopolymeric substances molecules with the use of reagents (e.g., glutaraldehyde), thus leading to the formation of cross-linked aggregates (Datta et al. 2013). Encapsulation or cross-linking of exopolymeric substances is generally achieved with the use of alginate beads, glass beads, foam blocks, cellulose acetate, polyacrylamide, or agar (De Philippis et al. 2011). Immobilized exopolymeric substances would naturally confer a better rate of metal removal. Ozdemir et al. (2005) validated this when exopolymeric substances from Chryseomonas luteola immobilized in alginate beads revealed a better adsorption capacity for Cu (1.989 mg/g) and Ni (1.284 mg/g) compared to adsorption of 1.505 and 0.996 mg/g, respectively, by free (non-immobilized) exopolymeric substances. The immobilization of exopolymeric substances is essential to ensure that exopolymeric substances are not washed away by the flow of water (De Philippis et al. 2011). Furthermore, the immobilization process will prevent clogging of exopolymeric substances in the water systems and helps to concentrate exopolymeric substances into higher loading potential. Immobilization also allows

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for easy separation of the adsorbed metal ions from exopolymeric substances, facilitating regeneration and reuse of exopolymeric substances for subsequent cycles of metal biosorption (Park et al. 2010).

10.5.2 Media Manipulation One method to influence the amount and type of exopolymeric substances produced by microbial cells is to manipulate the growth media used for the cultivation of the microbes. This is because the presence or absence of certain types of nutrients elicits varying response from the microbes (Sheng et al. 2010). For example, the lack of carbon sources will cause a decrease in carbohydrate content, while a lack of nitrogen will cause a decrease of protein content in the exopolymeric substances matrix (Durmaz and Sanin 2001). Furthermore, carbohydrate content in exopolymeric substances for microbial sludge is reported to increase when there is an absence of phosphorus (Sheng et al. 2010). In addition, the presence of toxic substances (mimicking pollutant molecules) has also been reported to induce greater production of exopolymeric substances in bacterial cells (Sheng et al. 2010). Sheng et al. (2010) suggests that the protein and carbohydrate content in the exopolymeric substances matrix typically increases significantly in conditions where pollutants are present, as structural proteins are often produced in the exopolymeric substances matrix to degrade or remove these toxic substances. This is supported by Priester et al. (2006) who observed enhanced production of exopolymeric substances (rich in carbohydrate, e.g., N-acetyl-glucosamine, rhamnose, glucose, mannose) in P. putida, when Cr was included into the cultivation media.

10.5.3 Future Trends To date, the use of exopolymeric substances for metal removal has not been widely implemented, and the existing studies are mostly at experimental level. For microbial exopolymeric substances to be feasible in industrial settings, it has to compete with conventional wastewater treatment techniques that are already in place (Abbas et al. 2014). Thus, the metal removal efficacy and cost-effectiveness of microbial exopolymeric substances has to be further improved. The following are several possible methods, which can be potentially used to promote exopolymeric substances as an industrial metal biosorbent.

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Pretreatment of Exopolymeric Substances Pretreatment is the act of pre-applying a particular technique to biomolecules in order to increase its effectiveness and increase its feasibility at industrial scale. There are multiple studies showing the improvement of pretreatments for bio-based metal biosorbents such as agricultural waste or microbial cells. The purpose for pretreating metal biosorbents is to increase the number of functional groups present, thus improving its metal removal efficacy. For instance, it was discovered that HNO3 pretreated banana peels had a higher Zn adsorption capacity (2.75 mg/g) as compared to the original untreated banana peels (2.15 mg/g) (Annadurai et al. 2003). Furthermore, Yetis et al. (2000) pretreated cells of Phanerochaete chrysosporium with NaOH and discovered that the pretreated cells (20.1 mg/g) had a higher Pb adsorption capacity compared to the untreated cells (10.5 mg/g). Therefore, although exopolymeric substances typically have good biosorption potential, pretreatments are strongly recommended to further enhance its metal removal efficacy (Park et al. 2010; Gupta and Diwan 2017). Pretreatments involve several methods such as chemical treatments (acid or base), grafting, or chemical modification. Chemical treatments would involve soaking exopolymeric substances molecules in acid (e.g., HCl, H2SO4) or base (e.g. NaOH, CaCI2) solutions for a period of time prior to using them for metal biosorption. It is hypothesized that acids may be able to produce more functional groups for metal bonding by dissolving existing carbohydrate structures in the exopolymeric substances molecules (Farooq et al. 2010). In contrast, bases would convert methyl ester groups present in exopolymeric substances molecules into carboxylate groups, which in turn would produce more binding sites for metal ions. The equation would be listed as R‐COOCH3 þ NaOH ! R‐COO
þ CH3 OH þ Naþ Grafting involves introducing long polymer chains (e.g., acrylonitrile, hydroxylamine, and ethylenediamine) onto the surface of exopolymeric substances molecules. These polymer chains have a variety of functional groups; thus they would provide each exopolymeric substances molecule with additional metal-binding sites (Park et al. 2010). Besides that, chemical modifications such as acetylation and esterification can also be carried out onto exopolymeric substances molecules. These modifications involve the introduction of acetyl groups or ester groups to exopolymeric substances molecules to further increase the number of functional groups present on exopolymeric substances, thus improving its metal biosorption capability (Gupta and Diwan 2017). However, it is noted that these modifications might increase the cost required to manufacture exopolymeric substances products, thus bringing it closer to the price of man-made metal ion exchange resins (Park et al. 2010).

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Reusability of Exopolymeric Substances Future prospects in using exopolymeric substances, which can be explored, include effective desorption of metals from exopolymeric substances. This allows for easier detachment of metals from exopolymeric substances after the biosorption process, thus renewing the exopolymeric substances for subsequent biosorption cycles. Furthermore, valuable metal ions such as Ag or Au can be precipitated out for industrial use (Park et al. 2010). As such, exopolymeric substances will become a much more attractive prospect for industrial application with improved costeffectiveness (Park et al. 2010). One possible technique is through the usage of magnetic nanoparticles coupled with exopolymeric substances, as these nanoparticles are capable of giving magnetic properties to exopolymeric substances once bound to them (Tian et al. 2010). After metal biosorption is completed, exopolymeric substances could be separated out of solutions via an electromagnetic field, desorbed, and reused once again (Tian et al. 2010).

Genetic Engineering of Microbial Cells Genetic engineering involves the modification of an organism’s genes, which can increase or decrease expression of a certain product. Thus, genetic engineering could potentially be employed upon exopolymeric substances producing microbial cells in order to manipulate them to increase expression of specific carbohydrate or protein molecules (e.g., cysteine-rich peptides) vital in the formation of the exopolymeric substances matrix (More et al. 2014; Crini and Lichtfouse 2018). The increased production of microbial exopolymeric substances may potentially resolve issues pertaining to large-scale metal removal demands in the industry.

10.6

Conclusion

This chapter has revealed that exopolymeric substances can be created by different microbes (bacteria, fungi, and algae). Exopolymeric substances are a valuable biomolecule for use in metal removal, embodying the sustainable green approach. Exopolymeric substances is effective in binding metal cations, attributed to the many functional groups present (e.g., carbonyl, carboxyl, hydroxyl, phosphoryl, sulfhydryl, thiol, and amino). Future research of exopolymeric substances should focus upon pretreatments and reusability of exopolymeric substances as well as genetic engineering of microbial cells to increase production of exopolymeric substances. This would increase the cost-effectiveness and metal removal efficacy of exopolymeric substances, thus making it more feasible to be implemented in industrial settings. The implementation of a cost-effective biosorbent such as exopolymeric substances would therefore enable metal removal from wastewater

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on a massive scale, which in turn guarantees a cleaner environment and removes the threat of metal pollutants to human health. Acknowledgments The authors acknowledge the funding by Malaysian Ministry of Education under the FRGS grant (FRGS/1/2018/STG03/MUSM/02/1). The authors also thank Monash University Malaysia for the facilities and resources for studies related to exopolymeric substances for metal removal.

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Chapter 11

Bioremediation of Bisphenols and Phthalates from Industrial Effluents: A Review Meghana Ganta, Anuradha Shilli, Soukhya Channapatana Adishesh, Bhanu Revathi Kurella, and Shinomol George Kunnel

Contents 11.1 11.2

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioremediation of Bisphenols . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.1 Bacterial Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.2 Physical Membrane Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2.3 Myco-remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3 Bioremediation of Phthalates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3.1 Nano-bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3.2 Myco-remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3.3 Bacterial Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.3.4 Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.4 Work in Progress . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.5 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Human exposure to various plasticizers has been increasing in an alarming pace. Bisphenols and phthalates are two such known plasticizers that are found to cause varied health impairments in various organs and also in fetuses and newborns. These compounds have been found leaching into food and water through routine sources such as plastic containers and canned foods along with hospital equipment such as tubings, catheters, and dialysis bags. It has been found to be prevailing in most of the body tissues as observed in various animal models and also in breast milk. The extent of damage varies from developmental, behavioral, and cardiovascular systems to even as endocrine disruptors. Bioremediation aims at removal of these toxins from the site of pollution for their detoxification and disposal. Bioremediation has been found to be a low cost and eco-friendly solution that can accumulate and concentrate the toxin to the point of

M. Ganta · A. Shilli · S. C. Adishesh · B. R. Kurella · S. G. Kunnel (*) Department of Biotechnology, Dayananda Sagar College of Engineering, Bengaluru, Karnataka, India © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 253 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_11

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easy disposal. Many living systems or their components can be used as agents for toxin removal among which we have discussed a few bacteria, fungi, and plant biomass for the removal of bisphenols and phthalates. The scope of their applications can be further enhanced by techniques such as nanoencapsulation and immobilization as indicated in this chapter. Keywords Bioremediation · Bisphenols · Endocrine disruptors · Mycoremediation · Phthalates · Phytoremediation

11.1

Introduction

Rapid growth in population and industrialization has resulted disposing various harmful compounds into the environment. The major sources of environmental contamination include industries, agrochemicals, mining activities, and waste disposals (Malla et al. 2018). Treating industrial effluent is very important as it may contain heavy metals, for instance, cadmium, mercury, chromium, cobalt, and copper, and organic chemicals like phenols, benzene, chloroform, naphthalene, and radioactive wastes (Cleary and DeVantier 2011). These contaminants effect the human health and surroundings (Govil and Krishna 2018). Measurement of total organic carbon, biochemical oxygen demand (also called biological oxygen demand), chemical oxygen demand, and also the concentrations of various chemicals in the effluents released by industries has become vital to prevent bioaccumulation and hazard management. The above values are used to measure organic pollution by measuring the extent of organic compounds in waste. Effluents from textile industries contain higher biological and chemical oxygen demand value compared to other industries (Vandevivere et al. 1998). The ratio of these two designates degradation of organic matter and value is expected to be less than 0.8. When the value is equal or above 0.8, it indicates water pollution (Yang et al. 2014). According to regulations of the Ministry of Environment, Forest and Climate Change, total organic carbon limit in underground water is 10 mg/L near waste management sites (Koda et al. 2017). Bisphenols belong to a cluster of compounds analogous to 4,40 -(propane-2,2diyl)diphenol (2,2-Bis(4-hydroxyphenyl)propane, bisphenol A, BISPHENOL A) (Danzl et al. 2009). Among them, bisphenol A, an artificial estrogen whose occurrence was detected in many consumer products, is widespread in exposure to humans and wildlife. In particular, developmental stages of animals could be more susceptible in most of the endocrine disruptors such as BISPHENOL A, affecting various changes in the offspring, such as changes in the development and function of reproductive organs, postnatal growth, and behavior of the organism (Miyagawa et al. 2016). Other than BISPHENOL A, 4,40 -dihydroxydiphenylmethane, commonly known as bisphenol F, and also another form 4,4’-sulfonyldiphenol,

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bisphenol S, are other monomers commonly used in the resin industry and various forms of plastics and canned goods and exist even in household dust (Danzl et al. 2009). Phthalic acid esters are ubiquitous, man-made compounds serving as plasticizers and as additives in paints, insecticides, lubricants, and adhesives and in cosmetics industry (Fan et al. 2018). Phthalates are xenobiotic compounds obtained from catabolism of phthalic acid esters through a series of sequential hydrolysis by esterases by biological agents (Ebenau-Jehle et al. 2017). Bisphenols and phthalates are the two most common endocrine disruptors that are held responsible for parental and social behaviors and developmental disorders (Rosenfeld 2015). Diethylhexyl phthalate is an ubiquitous compound which along with its metabolite monoethyl hexyl phthalate can cause major neuroendocrine disruption and thus affect the levels of various estrogens and other steroid hormones (Leon-Olea et al. 2014). Bioremediation is a natural course of detoxification of pollutants through biological action that act as catalyst at several stages of ecosystem (Mohanakrishna et al. 2016). It is a microbial degradation process by which bacteria, fungi, or protists convert toxins into harmless product in the form of water and carbon dioxide (Samantaray et al. 2014). Bioremediation strategies range from encouraging natural processes of biodegradation known as biostimulation to supplement the existing system with microorganisms capable of contamination degradation known as bioaugmentation, monitoring, and verification of natural processes also known as natural attenuation (Devi et al. 2014). Recently it has been proved reliable and most effective due to its environmental aspects in nature and profitable method (Azubuike et al. 2016). Based on the site of application, bioremediation includes two major types: in situ and ex situ. The selection of the type of bioremediation depends on the nature of pollutant, extent and grade of pollution, and type of surroundings, location, and cost. The main types of bioremediation include bacterial bioremediation, mycoremediation, nano-bioremediation, and phytoremediation. The main concerns of human and wildlife exposure to bisphenol cause high body weight, elevated risk of breast and prostate cancers, reformed reproductive function, and other prolonged health effects. Bisphenol A is also a prostate, ovarian, and uterine toxicant, and the potential of bisphenol A observed below the lowest level of adverse reaction causes polycystic ovary syndrome (Marciniak et al.2018). After prenatal or adult exposure, bisphenol A decreases the synthesis and secretion of insulin and diminishes the action of insulin following prenatal, perinatal, and adult exposure (Battistoni et al. 2018). Exposure to phthalic acid esters disrupts endocrine pollutants, effects male and female fertility and early onset of puberty, and induces hepatic peroxisome proliferation, reproductive toxicity, hormonal disorders, and carcinogenicity (Zhang et al. 2016, Konieczna et al. 2015). Frequent usage of plastic and other industrial products has led to increase in concentration of toxic chemicals in the human body and environment which has led to dangerous effect on human health and wildlife and environmental pollution. Hence, bioremediation plays a major role in minimizing these concentrations of toxic chemicals in ecological life (Table 11.1).

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Table 11.1 Deleterious effects of bisphenols and phthalates Sl. No. 1.

Pollutant Bisphenols

2.

Phthalates

11.2

Mode of action Contaminated drinking water, soil

Contaminated drinking water, soil

Effects Heart disease Diabetes and elevated body weight Development of fetal brain Prostate and breast cancer Asthma Ovarian toxicity Infertility Premature ovarian failure Nonreproductive disorders Asthma Cardiovascular diseases Hormonal disruption Breast cancer Neuron degenerative diseases

References Gao and Wang (2014) and Gao et al. (2015)

Hannon and Flaws (2015), Bølling et al. (2013), Philips et al. (2017), Chen et al. (2018), and Tseng et al. (2013)

Bioremediation of Bisphenols

Bisphenol is a chemical complex utilized in plastic industries for the synthesis of polycarbonate plastics and epoxy resins. Polycarbonate plastics are extensively used in manufacturing packaging materials for foods and beverages. Epoxy pitches are used to provide finishing to cover metal substances such as bottle caps, sustenance jars, and water supplying channels. The presence of bisphenol A in nourishment foods and drinks represents most of day-by-day human exposure (Lehmler et al. 2018). The essential wellspring of introduction to bisphenol A for the vast majority is through the eating routine. Bisphenol A is emptied into sustenance out of the cautious internal epoxy gum coatings of canned sustenance and from purchaser things, for instance, polycarbonate flatware, sustenance amassing holders, water containers, and tyke bottles (Konieczna et al. 2015). The amount of bisphenol A eluting from polycarbonate bottles into liquids is determined to a greater extent by the temperature of the liquid or container, than the age of the compartment (Lehmler et al. 2018). Bisphenol S has been gradually used as a standard alternative of bisphenol A due to well-being deportations with regard to prior effects. In any event, increasing

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evidence proposes that bisphenol S producse relative endocrine-disrupting impacts as bisphenol A, and likewise its amount on land can pose significant hazards to biological processes and natural well-being (Qiu et al. 2018). Bisphenol AF, a fluorinated organic compound, is associated with bisphenol A, consisting of two methyl groups that are replaced with trifluoromethyl groups. The neurotoxicity of bisphenol AF and hidden components of activity were done on mural essential neuronal cells and hippocampal cell line (HT-22 (Pang et al. 2019). Bisphenol AF expanded the dimension of intracellular calcium, trailed along with the age of responsive oxygen species (ROS). Bisphenol AF was found to upregulate the phosphorylation of mitogen-actuated protein kinase: the extracellular flag-directed kinase, the c-Jun N-terminal kinase, and the protein p38, and also the atomic translocation of atomic factor-κ B. Bisphenol AF interfered with the typical physiologic elements of microglia at non-harmful dimensions. Therefore, it shows that bisphenol AF has neurotoxic properties (Lee et al. 2013). The compounds bisphenol A diglycidyl ether and bisphenol F diglycidyl ether are utilized as a glazing material for inner covering for nourishment cans. The androgenic and antagonistic impacts of the test synthetics by correspondent quality test with these abovementioned cell lines were determined (Satoh et al. 2004). Results recommended that the chlorohydroxy mixtures of the above two compounds go about as androgenic opponents across the way toward authoritative binding to androgen receptor (Satoh et al. 2004). Based on the molecule mass recovered from the soils connected with the mass of soil, the half-lives of bisphenol S were found to be less than 1 day in both bisphenol A-like soils, and bisphenol AF was found to be much more ferocious with recorded half-lives of 32.6 and 24.5 days successively in forest and farm crops (Choi and Lee 2017). The metabolites known for all the three bisphenols undergo meta-cleavage and ortho-cleavage degradation pathways (Choi and Lee 2017).

11.2.1 Bacterial Bioremediation The endophytic organism Dracaena sanderiana was examined to assess the impact of bacterial immunization on BISPHENOL A deportation in hydroponic circumstances. Two promoting bacterial species of plant development, Bacillus thuringiensis and Pantoea dispersa, which have elevated bisphenol A strength, can use bisphenol A for growth which have been used in plant inoculation. The studies on P. dispersa-immunized plants by Suyamud et al. (2018) demonstrated the most astounding bisphenol A evacuation productivity at 92.32  1.23% contrasted with other vaccinated and non-vaccinated plants. Bacterial consortia assemble a wide range of catabolic pathways to enable bisphenol A to be efficiently degraded, converting bisphenol A to mainly bacterial biomass and low- or non-estrogenic metabolites (Eio et al. 2014). Based on the 16 s RNA gene cluster analysis, a bacterium extracted from the waste products of the thermal paper industry was recognized as Pseudomonas aeruginosa, which could live on a basal mineral salt

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medium modeled on bisphenol A, which operates as the sole source of carbon (Vijayalakshmi et al. 2018).

11.2.2 Physical Membrane Bioremediation One of the studies showed that immobilization of the Pavlova cells in sodium alginate enhanced the product yield by 17%. In a similar research, benzophenone was converted to diphenylmethanol with a conversion rate of 49% and diphenylmethyl β-D-glucopyranoside by 6% in Pavlova cell cultures. Incubation in medium containing benzophenone with immobilized Pavlova cells yielded higher amounts of products. The yields of the products diphenylmethanol and diphenylmethyl β-D-glucopyranoside were estimated to be 85% and 15%, respectively (Shimoda and Hamada 2009). In a similar research, Amphidinium crassum, a marine microalga, immobilized to form a new compound by which bisphenol A was glycosylated into glucoside, and immobilized Catharanthus roseus plant cells degrade bisphenol A to its glucoside, diglycoside, gentiobioside, and gentiobiosylglucoside (Shimoda et al. 2011). A biopolymer, made of coral-based scaffold from Hippospongia communis poison hemlock, has been used as support for laccase immobilization acquired from Trametes versicolor. To improve the removal of bisphenols, the techniques of immobilization and anaerobic decomposition have also been repeatedly optimized (Zdarta et al. 2018). Although bisphenol A degradation is small, the ground microbiological society has the capacity to erode bisphenol A at or below 1.0 mg/g land. The microbial community not only degraded bisphenol A in more than 10 mg/g soil but also reduced its diversity, signifying that bisphenol A can be accumulated by various microorganisms (Matsumura et al. 2015).

11.2.3 Myco-remediation Myco-remediation is a parasite-based innovation which is utilized to sterilize the earth. Growths have been turned out to be a shoddy, viable, and ecologically solid path for expelling a wide exhibit of poisons from harmed conditions or wastewater. A cell surface enzyme formed by the fungus, Trametes versicolor, genetically engineered in tobacco crops, produces lignin peroxidases. Six genetically engineered crops called FLP-1, FLP-2, FLP-3, FLP-4, FLP-5, and FLP-8 displayed lignin peroxidase in natural plant form. The enzymes FLP-1, FLP-2, and FLP-8 could efficiently produce 10 μmol of bisphenol A per gram of soil weight of aquaponic plants. Thus, the transgene enhances the environmental decontamination capacity of plants by introducing enzymes to microorganisms that possess degradative capacity (Sonoki et al. 2011).

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Table 11.2 Summary of bioremediation strategies conducted on bisphenols Sl. no. 1. 2.

3.

Types of bisphenols Bisphenol A Bisphenol S Bisphenol AF

Bioremediation Adsorption by Spanish broom (SB) surface-modified cellulose fibers Adsorption by dissolved organic matter (DOM) with modified biochar

Nature of sample Wastewater

Biotransformation of BPF to BPF-G

Urine of SpragueDawley (SD) rats

Ferralsol

References Tursi et al. (2018) Zhang et al. (2019) Li et al. (2013)

A ternary isooctane system: AOT [Bis (2-ethylhexyl) sulphosuccinate sulfosuccinate sodium salt) ]: water was used to remove impurities of lacquers integrated in inverted micelles with a humidity ratio (Wo) of 30, and proteins with a volume of 43.5 μg/ml was observed to counterbalance 91.43% of 200 ppm of Bbisphenol A at pH-6.0, 50ºC when implanted with laccase/Rreverse Llipid bilayers system implanted for a period of 75 minutes (Chhaya and Gupte 2013). Inverse Mmicellar systems function as strength for protein purification processes to detoxify multiple poisonous and intransigent substances by fungal enzymes, catalases, varnishes, peroxidases, and monooxygen cyrochrome P450 (Deshmukh et al. 2016) (Table 11.2).

11.3

Bioremediation of Phthalates

Phthalic acid esters are refractory compounds, i.e., they are resistant to degradation by heat, pressure, and temperature. According to United States Environmental Protection Agency and China National Environmental Monitoring Center, the compounds dimethyl phthalate, diethyl phthalate, di-n-butyl phthalate, di-n-octyl phthalate, butyl benzyl phthalate, and di(2-ethylhexyl) phthalate are the phthalic acid esters that are listed in priority as environmental pollutants (Fan et al. 2018). Factors that affect biodegradation of these esters are temperature, pH, salinity, organic chemical and nutrient content, and oxygen. Phthalates are classified into low-molecularweight phthalates those are derived from three to six carbon alcohols such as dibutyl phthalate, di(2-ethylhexyl) phthalate, and phthalates possessing high molecular weight and six carbon atoms in their back bone such as diisononyl phthalate, diisodecyl phthalate, and dipropylheptyl phthalate. Phthalates with short ester chain degrade faster than long ester chain such as dibutyl phthalate degrade faster than diethyl hexyl phthalate (Ahuactzin-Pérez et al. 2018a). Table 11.3 shows the different types of phthalates and suitable bioremediation methods used.

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Table 11.3 Summary of bioremediation strategies for the removal of phthalates Sl. No. 1. 2.

4. 5.

Types of Phthalates Di-n-butyl phthalate Di(2-ethylhexyl) phthalate Dimethyl phthalate Diethyl phthalate

Bioremediation Degradation by Achromobacter, halotolerant Sphingobium Degradation by Rhodococcus ruber YC-YT1, reduction by sunflower, leaf mustard, Ceylon spinach Degradation by unicellular cyanobacter Anaerobic biofilm-centered bioreactors, such as anaerobic fixed film baffle based reactor and upward flow anaerobic fixed film fixed bed reactor

Nature of sample Rural domestic water Contaminated water and soil Freshwater Synthetic wastewater

References Jin et al. (2013, 2015) Yang et al. (2018) and Wu et al. (2019) Zhang et al. (2016) Ahmadi et al. (2017)

11.3.1 Nano-bioremediation Nano-bioremediation is the new emerging technique used to remove contaminants from environment where synthetic nanoparticles are used. It is a combination of both bioremediation and nanotechnology (Yadav et al. 2017). These nanoparticles are synthesized from plants, fungi, and bacteria. Biological method of synthesizing nanoparticle is simple, cost-effective, and non-toxic to human health and environment (Kulkarni and Muddapur 2014). Nowadays, nano-bioremediation of radioactive element such as uranium is widely used. Complex organic compounds are degraded into simpler molecule with the help of nano-capsulated enzymes. These simpler molecules further degraded by action of microbes and plant. By conventional technologies, it is very difficult to treat wastewater, whereas nanoparticles have ability to penetrate deeper due to its unique features such as high interaction reaction and absorption capacity (Yadav et al. 2017).

11.3.2 Myco-remediation Myco-remediation plays a major role in degrading the wastes from industrial effluents. Fungal growth is more stable than bacterial growth since bacteria are sensitive to harsh environment (Deshmukh et al. 2016). A fungal plant pathogen, Fusarium culmorum, converts dibutyl phthalates into fumaric and malic acid, enters in to Krebs cycle, and helps in production of citric acid. There are many white rot fungal species such as Bjerkandera adusta, Phanerochaete chrysosporium, Trametes versicolor, Pleurotus sp., etc. that have been used widely to degrade endocrine disruptors from different environmental pollutants. They synthesize enzymes such as laccases and peroxidases which convert complex organic molecule into simpler molecule (Ahuactzin-Pérez et al. 2018b).

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11.3.3 Bacterial Bioremediation Recent studies have shown that bacteria are used to degrade phthalic acid esters. Cyanobacteria are prokaryotes those have high photosynthesis regulation than the terrestrial plants. Phthalic acid esters at a concentration of 200 milligrams per liter promotes growth of cyanobacteria (Zhang et al. 2016). The enzymes that participate in the bioconversion of phthalic acid esters are esterases, peroxidases, phthalate hydrolases, and the enzymes belonging to phthalate metabolic gene cluster. The pathway followed in degradation of phthalates by bacteria is the conversion of phthalic acid esters to phthalates and corresponding alcohols. Further, by the action of two enzymes phthalate 4, 5-dioxygenase and the enzyme 3, 4-dioxygenase, the phthalates are transformed into protocatechuates. By a series of reactions, protocatechuates are broken down into carbon dioxide and water (Jin et al. 2015). In one of the case studies by Yang et al. (2018), a liquid sample containing elevated levels of the toxin di(2-ethylhexyl) phthalate was considered and proved that Rhodococcus ruber YC-YT1 had potential to degrade 60% of it. Other species of bacteria were also used to degrade phthalic acid esters such as Bacillus, Pseudomonas, Rhodococcus, and Flavobacterium (Yang et al. 2018). One of the case studies had shown that immobilized microorganisms degrade di-n-butyl phthalate at a faster rate compared to free suspended microorganisms (Hu and Yang 2015).

11.3.4 Phytoremediation Phytoremediation refers to changing natural and inorganic contaminants that are brought with water to unpredictable vaporous species inside the plant pursued by their possible discharge into the climate at relatively low fixations. Phytoremediation plays a major role in degrading phthalates. Cultivating plants in contaminated soil provides proper environment for growth of microorganisms, and plants release the enzymes that help in degradation of phthalates and also improves quality of soil. One of the reports showed that they cultivated three plants, leaf mustard, sunflower, and Ceylon spinach in order to degrade di-(2-ethylhexyl) phthalate. After 45 days of cultivation in contaminated soil, they observed there was increased shoot and root biomass of plants and decreased level of degrade di-(2-ethylhexyl) phthalate (Wu et al. 2019).

11.4

Work in Progress

Bisphenol A and di-2-ethylhexyl phthalate that are widely used in chemical water bottles are associated with diabetes, obesity, high blood pressure, migraine, infertility, cancer, and neurobehavioral and immune system effects. The ongoing study

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aims at understand the effects of chemicals on plastic water bottles and to assess the level of bisphenol A and diethyl hexyl phthalate leaching. Timeline and temperaturebased studies are being conducted for various plastics grades. High-performance liquid chromatographic analysis and timeline-based studies are being conducted followed by compilation of data. Preparation and characterization of biosorbent for the mitigation of bisphenol A and diethyl hexyl phthalate from water samples are tested using Fourier transform infrared spectroscopy and scanning electron microscopy for surface adsorption characteristics. Assessment of disruption acetylcholine esterase activity in zebra fish embryos in the presence and absence of biosorbent by bisphenol A and di-(2-ethylhexyl) phthalate. Activated charcoal from Citrullus lanatus rind powder acts as biosorbent that helps to reduce the bisphenol A and diethyl hexyl phthalate contents from bottled water samples.

11.5

Conclusion

Over the past few years, the expansion of population and industries has led to gradual pollutant increments by xenobiotics and other toxic chemicals. This review deals with suitable bioremediation methods used to degrade bisphenols and phthalates and its ill effects on human and environment. Recent researches on in vitro exposure studies indicate that bisphenol A and phthalates may be arrhythmogenic, prominently in female subjects (Gao 2015). It is important to improve novel methods for bioremediation, and further bioremediation helps in tracking organisms that can mitigate and are efficient in eliminating the pollutants from the site of persistence. Nowadays, nanotechnology-based methods are prevalent solutions in the environment. It is a necessary point of action to develop new methods and better perception of nanotechnology-based innovations for the development of cost-effective alternatives for bioremediation. Acknowledgment The authors would like to thank the management of Dayananda Sagar College of Engineering, Bangalore, for their support and Karnataka State Council for Science and Technology for their financial support for the project 42S_BE_2609 under SPP Stream C. Note: The graphical abstract was created in PowerPoint using ChemDraw and google images. It is a completely modified image created by the authors.

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Chapter 12

Phytoextraction of Heavy Metals A. N. Anoopkumar, Sharrel Rebello, Elsa Devassy, K. Kavya Raj, Sreedev Puthur, Embalil Mathachan Aneesh, Raveendran Sindhu, Parameswaran Binod, and Ashok Pandey

Contents 12.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.2 Heavy Metals: Sources, Contamination, and Effects in the Environment . . . . . . . . . . . . . . 12.3 Deposition of Toxic Elements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.4 Metal Accumulating Plants: The Notable Natural Resource for Phytoextraction . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Heavy metals are considered as the major classes of a contaminant in nature. Heavy metal contamination from fertilizers, metal mining, and industrial activities leads to toxic effects on humans and other organisms. Although the toxic effects of these elements have been recognized for a long time, exposure to these elements continues. The toxic effects induced by them can lead to death in humans. Several advanced strategies have primarily employed to tidy up the surrounding from toxicants; however, most of the strategies are considered as problematic when getting results. The current concerns regarding the contamination due to heavy metal deposition have introduced the novel advanced technologies to detect the presence of them from soil and wastewater. Several plants have been recognized as potent herbs since they are able to absorb these toxicants. The phytoremediation techniques A. N. Anoopkumar Department of Zoology, Christ College, University of Calicut, Irinjalakuda, Kerala, India Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India S. Rebello · E. Devassy · K. Kavya Raj · S. Puthur · E. M. Aneesh Communicable Disease Research Laboratory (CDRL), St. Joseph’s College, Irinjalakuda, Kerala, India R. Sindhu (*) · P. Binod Microbial Processes and Technology Division, CSIR-National Institute of Interdisciplinary Science and Technology (CSIR-NIIST), Trivandrum, Kerala, India A. Pandey CSIR-Indian Institute for Toxicology Research (CSIR-IITR), Lucknow, Uttar Pradesh, India © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 267 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_12

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usually uptake the heavy metals from the soil and wastewater and recognized as a well-established approach to remediate the toxic effects induced by them. The plantbased techniques have some advantages over the conventional strategies. Therefore, the present research has focused on the various technologies used to remove the metal pollutants from natural resources especially from water and soil. These affordable and effective technologies are potentially cost-effective and environmental friendly. Keywords Contaminants · Heavy metals · Phytoextraction · Remediation

12.1

Introduction

Environmental pollution by the deposition of toxic heavy metals in natural resources including water and soil has become a serious threat over the globe in recent years. The disturbance of natural biogeochemical cycles and industrialization together with urbanization has instigated the toxic impacts (Srivastava 2016). Industrial actions including metal fabrication shops, textile factories, service stations, waste disposal areas, and chemical works followed by intensive cultivation are specifically awkward of contaminating the environment (Freitas et al. 2004; Wong 2003). The enhancing reliance on synthetic fertilizers in the soil for the agricultural purpose has enforced a long-term threat over the environment (Barla et al. 2017; McLaughlin et al. 1999; Rebello et al. 2019). Emission of toxic elements has principally regulated in industrialized countries, whereas in the case of developing countries, hasty population explosion and industrial development coupled with worst pollution control strategies that have ensued in wide heavy metal pollution over natural resources (Ji 2000). The entry of heavy metals into the food chain consequently develops a severe impact to animal and human health (Sarwar et al. 2017). The heavy metals usually have a high atomic number, high density, and mass (Alloway 2012). These elements also accumulated in the living organisms, referred to as bioaccumulation, thereby increasing their concentrations in higher trophic levels by biomagnification. The major heavy metals include copper (Cu), manganese (Mn), iron (Fe), chromium (Cr), and nickel (Ni). The excessive accumulation of these elements in plant cells has induced toxic effects even at low concentrations by adversely affecting photosynthetic and respiratory processes, plant growth, membrane integrity, deoxyribonucleic acid structure, and functionality and enzymatic activities (Lajayer et al. 2017b). In addition, certain redundant compounds including lead (Pb), mercury (Hg), cadmium (Cd), and arsenic (As) are considered as highly toxic towards various biochemical and physiological processes take place in plants (Lajayer et al. 2017a). The conventional strategies for indemnification of heavy metals tainted water and other natural resources might not be feasible in most of all the situations since they are expensive and inefficient (Lajayer et al. 2017a). This chapter thereby determines

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the various strategies and concepts and risks and effectiveness involved in the phytoextraction of heavy metals associated with environmental contamination and natural resources.

12.2

Heavy Metals: Sources, Contamination, and Effects in the Environment

Current disquiets concerning the heavy metal-induced environmental damage have extended the need of developing appropriate novel and versatile technologies to find out its presence in natural resources. Instead of numerous contaminants, the effluent from various industrial firms including paper mill is recognized as a rich source of nitrogen, calcium, phosphorus, and magnesium together with trace elements (Jain et al. 2005). Pollution due to the accumulation of metal elements cadmium, chromium, copper, silver, lead, and zinc induces a hazardous effect on biological systems since most of them do not endure biodegradation (Table 12.1) (Pehlivan et al. 2009). Table 12.1 Major plants those are able to accumulate heavy metals Sl. no. 1.

Metal Nickel

2. 3.

Cobalt Zinc

4.

Copper

5.

Lead

6. 7.

Selenium Cadmium

8.

Manganese

9.

Chromium

Plant species that accumulate heavy metals Alyssum serpyllifolium Bornmuellera kiyakii Sebertia acuminata Berkheya coddii Crotalaria cobalticola Picris divaricata Arabis gemmifera Sedum alfredii Arabidopsis halleri Crassula helmsii Ipomea alpina Hemidesmus indicus Plantago orbignyana Sesbania drummondii Stanleya pinnata Arabidopsis halleri Bidens pilosa Virotianeurophylla Austromyrtus bidwillii Maytenus founieri Phytolacca americana Leersia hexandra Gynura pseudochina Salsola kali

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Hence, these toxic heavy metals need to be transformed into non-toxic compounds using various physic-chemical strategies (Gaur and Adholeya 2004; Rebello et al. 2018). The anthropogenic sources such as pesticides, energy, power, construction, steel manufacturing, waste incineration, food processing waste disposal, smelting, and mining followed by coal combustion and military operations are primarily considered as important sources of environmental contamination. This has indicated that the enhanced anthropogenic activities have influenced the environment directly through food chain since heavy metal contamination in vegetables causes potential health threats and safety issues (Ugya et al. 2019). Irrigation of agricultural fields using wastewater frequently has resulted elevated level of heavy metals in several crops (Pan et al. 2016; Rana et al. 2014). In addition, the poor agricultural practices, misuse of the soil, and urban waste disposal, followed by the regular use of the chemical have resulted in soil pollution (Li et al. 2016; Mahmoud and Ghoneim 2016). It is well known that the frequent consumption of food sources tainted with the aforesaid elements has cause severe risk to human health. One of the reasons for this condition is the accumulation of heavy metals at an elevated level. The frequent intake of a hazardous range of heavy metals through food resources has resulted in chronic accumulation in the liver and kidney of humans followed by triggering disruption of several biological and physiological processes including kidney, cardiovascular, bone, and nervous-associated diseases (Mahmood and Malik 2014). Several cleanup inventions are developed for the detoxification of contaminated natural resources; however, few technologies stand efficient to decontaminate the resources. This has enhanced the need for using plants and plants allied microorganisms to degrade, inactive, or remove the toxic environmental contaminants and to invigorate contaminated regions which are receiving much more care and consideration (Vaikosen and Alade 2017).

12.3

Deposition of Toxic Elements

Several toxic elements including nickel, manganese, copper, and zinc are important micronutrients required for the completion of plants’ life cycle. Hence, the concept concerning the use of the plant as a phytoremediation technique has gained much more interest in in the past few eons (Number et al. 1997; Shiowatana et al. 2001). The important heavy metals influence the economic balances by reducing the crop production and induce the risk over and groundwater contamination. Research based on basic chemistry, ecological impacts, and linked health disorders caused by heavy metals are essential to find out their bioavailability and remedial preferences together with speciation. The speciation and chemical nature of the element has also influenced the fate and spread. The initial fast reactions of heavy metal contamination normally require minutes to hours, whereas its slow adsorption necessitates days and years. This indicates their diverse chemical form with toxicity, and

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bioavailability has principally involved in the redistribution of toxic elements (Buekers 2007). The distribution of the aforesaid metals is supposed to be controlled by various processes induced by metals from the resources. They include (i) dissolution and precipitation of minerals; (ii) ion exchange; (iii) adsorption, desorption, and aqueous complexation; and (iv) biological immobilization followed by plant uptake. The accumulation of lead in the kidneys, central nervous system, and gastrointestinal tract may lead to severe threat including poisoning, hyperactivity, decreased reaction time, weakness of the joints, anorexia, loss of memory, lower intelligence quotient, mental deterioration, insomnia, shortened attention span, and impaired development even death. It is well reported that its toxic effects have been much more broadly studied than any other heavy metals in trace amounts since it can generate severe bruise to the kidneys, neuron system, and brain together with a vast range of toxic effects in biological systems. Direct ingestion is primarily renowned as the one of the major sources of lead exposure from the soil. Previous studies have reported its toxic effects; however, they also revealed that it is never accumulated in the fruit crops (strawberries, beans, tomatoes, corn, apples, and squash) in large amounts. Its high concentrations are principal to be found on the exterior of leafy vegetables (Wuana and Okieimen 2011). Arsenic is one of the heavy metals responsible for environmental contamination which always induced notorious toxic effects to humans and other living organisms. The various species of these heavy metals determine its level of toxicity to the living organisms. The inorganic and organic form of arsenic is strongly influenced by the following factors such composition of minerals in surrounding environment, the pH, and activities executed by microbial communities. The two mineral species of arsenic, namely, arsenite and arsenate, are recognized as the foremost one in the mainstream of the environment (Yusof and Malek 2009). These two aforementioned elements are known as highly toxic since the other forms are less toxic. The naturally occurring mercury is present silver–white, odorless liquid and shiny, and is present in various forms. It usually combines with various other elements including oxygen, chlorine, and sulfur and results in the formation of white powders or crystals containing inorganic mercury compounds. Due to its easy vaporization and low boiling point, mercury is always gained much more significance in numerous industrial products. Like any other element, mercury has been found as several forms in soil. Mainly three forms of silver have been found to be present in the soils, and they include (1) the most reduced Hg0 (2), ionic of mercurous ion Hg2 (3), and mercuric ion Hg2+. The major binding force for the adsorption of mercury includes electrostatic forces, precipitated as sulfide, hydroxide, phosphate, and hydroxide. Conventionally, certain anaerobic bacteria have been used for the methylation, a major way for the decontamination of mercury from the contaminated resource (Rodriguez et al. 2005). The aforementioned concepts and statements have verified that the organomercury compounds, especially the mercury salts, are among the major poisonous elements in our surroundings. In addition, the degree of toxicity and mechanisms of their action over the surroundings and

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organisms has strongly interlinked with the redox state and type of compound (Wagner-Döbler). The phosphate fertilizers, refined petroleum products, and detergents are the principal source of cadmium in environmental pollution. The acidification of soils by acid rain has enhanced the geochemical transport of cadmium. It is a biopersistent; however, it has exhibited only a few toxicological concerns, once ingested in the body of an organism, leftovers resident for a long period of time. Within the body, it seriously affects many enzymes and proteins and causes renal damage followed by proteinuria. Precisely, they generate harmful effects towards various important enzymes that are essential for the resorption of proteins and hormones. It is as well capable of diminishing the action of alcohol dehydrogenase, delta-aminolevulinic acid synthetase, and lipoamide dehydrogenase, followed by arylsulfatase; simultaneously it augments the activity of pyruvate decarboxylase, delta-aminolevulinic acid dehydratase, and pyruvate dehydrogenase. Copper is an important element indispensable for both plants (as a micronutrient) and hominids (essential for the production of blood hemoglobin) in several ways. At a minimal level, copper plays an important role in disease resistance and seed production together with regulation of water whereas its high concentration has resulted in kidney damage, anemia, stomach, and intestinal disorders and liver damage. It normally ingested into the human body through the drinking water from copper pipes. The contaminated soils have exhibited certain effects that are either direct or indirect. The direct effects comprise the negative effects on crop growth and profit, whereas the indirect concerns include the entry of toxic components into the food chain with potential adverse impact on the human population. The drastic reduction in crop production could also lead to economic loss for the long term. Nickel is a heavy metal element; it can generate toxic effects when the concentration has reached above the threshold level. It can cause various genetic disorders including cancer in animals. The major source of nickel is the steel manufacturing companies, plating industries, and industries. It may also be released into the atmosphere by trash incinerators and power plants. Most of all nickel compounds will adsorb to sediment when they are released into the environment. Nickel becomes leaches down to the neighboring groundwater in acidic soils and also induces adverse effects to microorganisms and thereby diminishes their population size. However, the microorganisms generally grow resistance against nickel.

12.4

Metal Accumulating Plants: The Notable Natural Resource for Phytoextraction

Most of the plants are able to accrue high levels of heavy metals in their tissues. In this regard, mainly three technologies such as phytoextraction and phytostabilization, followed by phytovolatilization, have been used for

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Fig. 12.1 Phytoextraction of heavy metals – a graphical representation. The heavy metals such as cadmium, nickel, copper, zinc, and lead from soil have been transferred to the shoots of the plants through roots. After the successful uptake of heavy metals by the root, it will penetrate through the different parts inside the plant and finally stored these compounds at the leaves

phytoremediation. The phytoextraction uses plants to decontaminate the elements from soil, and the phytostabilization reduces the mobility of metals, while the phytovolatilization converts metals to volatile chemical species. By virtue of this notable property, phytoextraction could be considered as economically viable and is alternative to the high cost of conventional remediation strategies (Wither and Brooks 1977). The process of heavy metals accumulation in plants involves a series of steps comprising the mobilization of soil, its uptake by roots, transport to shoot through the xylem, and distribution followed by storage in the tissues (Clemens et al. 2002). The uptake of elements by plant species strongly depends on diverse contributing aspects such as the total quantity of the available elements in soil, usually referred as quantity factor, the reaction kinetics, and the intensity factor (Brümmer 1986). Phytoextraction in a natural way is principally troubled by low biomass of heavy metals, and it has required a long period of time to reach environmentally permissible levels. Taking into the applications of phytoextraction, several authors have verified the highly specific and efficient mechanisms of plants to get vital micronutrients, even at low ppm levels. The phytoextraction feats the capability of herbs to uptake the metals from natural resources (Fig. 12.1) (Raskin et al. 1994). However, the phytoextraction has several limits that are arising from threatened metal obtainability and troubles in uptake by roots; and needs great energy as well as xylem loading and symplastic mobility (Meagher 2000). Plants exhibit diverse responses to metal contamination, and most of the morphophysiological responses are sensitive to very minute concentrations. The metal accumulation is usually expressed by “plant to-soil metal concentration ratio.” In addition, both the translocation factor and bioconcentration factor clearly influence the phytoextraction. Tolerant herb species have a habit of less accumulation of heavy metals since they restrict transfer between soil to root together with root to shoot. Usually, the plants exhibit greater than 1 BAC (biological absorption coefficient) can be used for

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phytoextraction; for phytostabilization, plants with higher than 1 bioconcentration factor and lower than 1 transcription factor value can be used (Brümmer 1986). The effectiveness of phytoextractions depends on several desirable properties such as (1) high biomass and rapid growth, (2) well-developed rot system for uptaking large volumes of elements, (3) effective tolerance to metals with high concentrations, (4) great translocation factor, and (5) adaptability against specific area. The phytoextraction of heavy metal-contaminated natural resources especially soil follows on the utilization of herbs to translocate the elements to their specific harvestable regions. Principal focus of phytoextraction is diminishing the concentration if heavy metals in contaminating resources within a short time frame (Nascimento and Xing 2006). The engineering calculations regarding the concept of phytoextraction suggested that effective plant-based decontamination of natural resources would normally necessitate crops capable to concentrate toxic elements in high. Deposition of heavy metals at high concentration would definitely destruct the non-accumulator plant within short time. However, biotechnology has now been productively employed to assess heavy metal uptake through plants. For instance, the expression of binding proteins that are specific against heavy metals and mammalian metallothionein has resulted in increased metal tolerance in Nicotiana tabacum (Lefebvre et al. 1987). Accumulation of heavy metals in natural resources is considered as an interesting area of research that, excepting to major commercial uses, should be responsible to provide answers to certain important questions of phytochemistry, nutrition, and environmental hazards followed by heavy metal accumulation. Phytoextraction of heavy metals at the industrial level, although still in its early stages, may 1 day grow into a well-known cleanup strategy. An integrated multidisciplinary study that links plant biochemistry, agricultural applications, soil microbiology, soil chemistry, and environmental engineering is essential for the development of phytoremediation techniques. It is imperative to note that plants that deposit toxic elements can be grown intended for economic benefits, leaving the water or soil including various natural resources with a diminished level of heavy metal contamination. Environmental cleanup using plants may guarantee a cleaner and greener earth for all of us. Acknowledgment Raveendran Sindhu acknowledges DST for sanctioning a project under DST WOS-B scheme.

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Chapter 13

Tree Barks for Bioremediation of Heavy Metals from Polluted Waters Puneet P. Jain, Zufeshan Nahar Ali, Srishti J. Sisodiya, and Shinomol George Kunnel

Contents 13.1 13.2 13.3

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Heavy Metal Contamination and the Need to Reduce These Contaminants . . . . . . . . . . . Methodologies That Exists for the Expulsion of Metal Ions from Effluents . . . . . . . . . . . 13.3.1 Physical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3.2 Chemical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3.3 Biological Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3.4 Problems Associated with the Existing Technologies . . . . . . . . . . . . . . . . . . . . . . . . . 13.4 Purification Using Natural Remedies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.4.1 Tree Barks: An Alternate Adsorbent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.4.2 Various Barks Used for Heavy Metal Purification . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.4.3 Biosorption: The Probable Mechanism of Heavy Metal Reduction via Tree Barks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5 Ongoing Work (Unpublished Data) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.6 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Water scarcity along with a decrease in the quality of water due to various pollutants is adding to the already existing problems related to water. Due to urbanization and industrialization, there has been an increase in the level of heavy metals present in water resources which is decreasing the quality of water and causing serious health issues, and therefore heavy metal levels need to be monitored, and their concentrations must be reduced in water resources using different techniques. Tree barks are known widely for their availability as well as low-cost metal chelating property for the adsorption of metals in an aqueous environment which acts as a biosorbent. It is known that the barks have a clarifying property and certain heavy metal uptake values analyzed and attained reduction using different species of tree barks. The numbers obtained are compared with the activated carbon sources which were P. P. Jain · Z. Nahar Ali · S. J. Sisodiya · S. G. Kunnel (*) Department of Biotechnology, Dayananda Sagar College of Engineering, Bengaluru, Karnataka, India © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 277 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_13

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commercially available. Bark being easily available and low cost over activated carbon has the possibility to be used without any pretreatment. Selection of specific bark species are usually suggested due to their specific activity with the heavy metals. Natural adsorbents are gaining importance due to their low cost and eco-friendly nature compared to other chemicals, where in the removal of heavy metals by certain tree bark species is a new technique in itself. Keywords Water scarcity · Industrialization · Tree bark · Adsorption · Metal chelating property · Heavy metal uptake · Natural adsorbent · Low cost

13.1

Introduction

Industrialization has led to the invention of various products such as pesticides, preservatives, medicines, surfactants, disinfectants, laundry detergents, dyes, paints, food additives, and personal care products which were essentially developed to serve mankind, but these products when disposed in the environment have found to be hazardous to humans as well as the surroundings. Apart from the products, there is a great amount of waste (rich in toxic heavy metals) generated in various industries which are discarded into the environment without treatment which leads to ecological imbalance thereby affects the life on earth. (Kümmerer 2009; Jaishankar et al. 2014). Today, apart from the conventional pollutants, there is an increasing amount of heavy metal pollutants which are being discharged into the environment which is causing lot of serious issues. The sources for these heavy metals are from combustion of fossil fuels like carbon mono oxide from vehicles, industries, and production of energy (Khan et al. 2008; Zhang et al. 2010). The major natural resource which is affected by such pollution is water, wherein the pollutants from industries along with decaying organic matter provide essential resources such as nitrates and phosphates for eutrophication which promotes excess plant growth thereby causing decrement in the dissolved oxygen level present in the water source (Wastewater Management – A UN-Water Analytical Brief). Additionally, there is rapid growth of pathogenic microorganisms such as bacteria and viruses which makes the water unfit for drinking and also pollute the water (Wang and Chen 2006).

13.2

Heavy Metal Contamination and the Need to Reduce These Contaminants

The dissemination of toxic metal compounds into the nature by various anthropogenic activities leads to spoilage of the surface bodies of water bodies like lakes, reservoirs, and ponds. Sometimes, even the groundwater is affected after they are

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Table 13.1 Health issues due to exposure to high heavy metal concentration Heavy metal 1. Cadmium 2.

Copper

3.

Chromium

4. 5.

Zinc Lead

6.

Mercury

Harmful effects due to high exposure Causes cancer or toxicity to various organ systems. Cadmium affects cell proliferation, differentiation, and apoptosis Causes stomach cramps, vomiting, diarrhea, headache, nausea, irritation to eyes, nose, and ears Causes allergic and eczematous skin reactions, dermatitis, skin and mucous membrane ulcerations, bronchial carcinomas, gastroenteritis, perforation of the nasal septum hepatic cellular deficiency, allergic asthmatic reactions, and renal oligo-anuric deficiency Affects growth, immunity, and neuronal development Affects central nervous system and causes convulsions, coma, and even death. It also causes renal impairment, anemia, hypertension, immune toxicity, and toxicity to the reproductive organs Affects the nervous, digestive and immune systems, and lungs and kidneys and may be fatal

being swept into water (Rhind 2009). Results of various industrial and domestic wastes are the accumulation of many heavy metals. Namely, arsenic, copper, lead, manganese, mercury, cadmium, chromium, nickel, and zinc. The metals are emancipated from anthropogenic sources like ceramics, wood preservatives, pesticides, dyes, mineral ores, metallurgical industries, and pesticide manufacturing industries (Tchounwou et al. 2012). Chromium (Cr) being mainly a major contaminant is comparatively released more into groundwater, air, and water also in the form of a particulate. Damage of gastrointestinal tract, liver, and lungs may be caused due to high level of chromium (VI) which is more reactive than chromium (III). The chromium (VI) concentrations in major rivers of India are way above the tolerable limits which is 50 μg/L in drinking water (Velma et al. 2009). Nickel (Ni) is emancipated from various smelting operations, thermal power plants, battery industries, and others. Cadmium (Cd) in water resources is discharged through waste batteries, paint sludge, fuel combustion, zinc smelting, and incinerations (Wilson 1988). The priority pollutants are considered to be metal ions due to their scattered property in water ecosystems and also because of their toxic levels. The main reason why the metals are considered as pollutants are because they are highly persistent and their biodegradable levels are very poor as they have the ability to accumulate in tissues and cells leading to the cause of various diseases and disorders. The detrimental effects of heavy metal toxicity are reduction in the functioning of mental and central nervous system; reduction of energy damages composition which indirectly affects vital organs such as the liver, lungs, and kidneys (Jaishankar et al. 2014). Table 13.1 illustrates the various health issues due to exposure to high concentrations of heavy metals. Various countries have introduced rules and regulation for the strict control of water purification. Therefore, the permissible limits for toxic heavy metals discharge into aquatic systems are given below. “These permissibility limits for heavy metals are given by the World Health Organization (WHO)

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which are 0.1 (Cadmium), 0.05–1.5 (Copper), 0.1–1 (Iron), 0.02 (Nickel), 0.1 (lead), 5–15 (Zinc), 0.05–0.5 (Manganese) (all these ranges are in ppm) (Environmental Protection Agency, 2013).” The traditional methods followed for the expulsion of metal ions from water have certain restrictions such as technical and financial barriers.

13.3

Methodologies That Exists for the Expulsion of Metal Ions from Effluents

Metal ions should be removed from the effluents using different technologies in various ways. The methods are categorized into three types: physical, biological, and chemical.

13.3.1 Physical Methods Physical methods include adsorption and membrane filtration. The fouling of the membrane leads to the major disadvantage of this process. Hence, conventional design of the adsorption procedure is helpful in forming better and high-quality effluents. Adsorption is considered to be one of the most reliable methods for removal of heavy metals from contaminated water (Bharakat 2011).

13.3.2 Chemical Methods The chemical methods are usually used to obtain better quality of water in very reliable economic rate, series of combinations/techniques like sedimentation, coagulation, electroflotation and electrofiltration, oxidization of water by using chemical agents, electrochemical process, and irradiation. They are comparatively expensive and are very hazardous in nature as the disposal of these chemicals is very dangerous. The common problems arising due to this method are utilization of these chemical reagents in an inappropriate way by consuming and leading to health problems and intake of higher than usual power supply (Matthew et al. 2016).

13.3.3 Biological Methods Different methods are required to lower the concentrations of heavy metals such that it reaches the permissible levels at most convenient rate and by lesser cost; usually

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the conventional methods are neither economical nor effective. An effective alternative could be the use of biological substances that will act as an adsorbent which will adsorb the heavy metals onto its surface thereby reducing its concentration in the environment. Many waste organic components have the efficiency to act as a good sorbent. The adsorbent which is used during the process should be economical, affordable, easy to use, shouldn’t cause any disturbance to the water other than cleaning it up, and should be easily available. Also, there shouldn’t be excess pretreatment or posttreatment steps (Kanamarlapudi et al. 2018).

13.3.4 Problems Associated with the Existing Technologies The methods opted for removal of effluents at present are very expensive. Few methods include solvent extraction, electrodialysis, ion resin exchange, membrane filtration, precipitation, and electrolytic processes. Wide ranges of conventional methods are present from reverse osmosis to activated carbon. The methods mentioned above are not reasonable for small scale. One easy technique for reduction of heavy metals is by precipitation, but it has a drawback as it leads to the formation of sludge, which has higher content of metal ions from heavy metals. The contaminants/effluents are required extra step treatments like adsorption, ion exchange, and reverse osmosis method (Azimi et al. 2017).

13.4

Purification Using Natural Remedies

Certain plants and seeds have the ability to purify water, like the seeds of Moringa which has an adsorbent property and clumps the debris together further causing the settling of particles, Coconut husks suspends the dirt from water, and then it is passed through burnt rice husks to remove debris, and xylem of plants also have the property to sieve the bacteria out. From past so many centuries, usually natural plant extracts were used for the purification of water. The plant extracts were derived from roots, leaves, barks, seeds, extracts of fruits, and vegetables. As an example, Moringa oleifera, commonly known as drumstick, is one of the fast-growing plants which is resistant of drought and belongs to the family Moringaceae (Abd El Hak et al. 2018). It was utilized as a water purifier back in the fourteenth and fifteenth centuries. Proof and evidences were noticed in certain Sanskrit shastras called Sushruta Samhitha. Strychnos potatorum (nirmalī) is also one of the plants which was used in those days along with Moringa oleifera. The moringa seeds can purify and also clarify the water which can be used to lower the concentration of bacteria in water and make it safe for drinking, basically potable water. It also replaces the chemical like aluminum sulfate which is not only expensive but also dangerous for life and environment. It not only destroys bacteria but also filters polluted water (Choy et al. 2014).

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According to the shastras, there was a combination of commonly available nine plants used to clarify the water and reduce contaminants. Usage of certain medicinal plant’s ash which was mixed with alum acts as a cleanser and also retains its original benefits. These low-cost old and ancient methodologies can act as an alternative and can replace the expensive and very complicated water purifiers (Fahiminia et al. 2014)

13.4.1 Tree Barks: An Alternate Adsorbent There is a significant increase in the metal discharge into the environment caused because of industrialization and urbanization. The discharge of toxic heavy metals like lead, cadmium, copper, and zinc into potable and wastewater is causing lots of serious issues with regard to the consumption and utilization of clean and pure water. For the effective removal of these heavy metals, scientists and researchers have identified the use of biological materials which are not only an effective adsorbent for removal of metals but are also cost-effective. Tree barks can be one of those biological adsorbents, as its structural porosity helps in effective collection, accumulation, and retention of aerosol particles, making the substrate an indicator in pollution with special emphasis on heavy metals (Akpor and Muchie 2010). The usage of bark for removal of heavy metal adsorption is a very different field of research as different bark species vary and depend on metal type as well. It has been seen that the amount of heavy metals that can be removed from water solution using a bark is comparatively high. Also, when compared to activated carbon or ion exchange resins, tree barks are a better option as it is reasonable and renewable resource for industrial use. Other advantages of it can adsorb heavy metals at low concentrations as well (below 100 ppm) and reduce those heavy metals (Vazquez et al. 2002) such as chromium reduction (Fiol et al. 2003; Aoyama et al. 2005). The floatability and weight/density of bark are considered in adsorption. After adsorption the recovery of metals can be done with acid washing (Horsfall et al. 2006). Therefore tree barks are biosorbents, and this method is easy to access; it has the potential to detect long-term contamination and is cost-effective (Norouzi et al. 2015). According to recent studies in central Italy, Pinus species barks and its ability to take up extremely high levels of mercury from the atmosphere was carried out. Through the study, it was observed that there was higher mercury concentration in the bark samples of the said area when compared to other zones of Firenze which provided evidences which support the ability of Pinus species to monitor mercury levels (Garty 2001). By modifying the bark, an effective adsorbent can be developed for removal of heavy metals and dyes from water. The phenolic moieties have relatively high capacities to reduce heavy metals (Randall et al. 1974). Formaldehyde was used for the pretreatment of the bark because of the problem of color leaching into the

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water solution (Saliba et al. 2002). The reduction of chromium (VI) can be done by synthesizing material and a relative rate. (Gloaguen and Morvan 1997). The reduction of contaminants from water effluents can be carried out by the biosorption technique which depicts positive results. Natural organic wooden materials are effective and affordable metal ion adsorbents from agro waste materials or by-products (timber industries), and the higher sorption efficiency was achieved with sawdust. It could be due to their higher specific surface area and smaller fraction. For using bark as adsorption materials, it is certainly appropriate to mill it to a smaller fraction and thereby to increase the surface area of the absorbent (Ghaedi and Mosallanejad 2018).

13.4.2 Various Barks Used for Heavy Metal Purification (i) The comparison of better rate of adsorption was checked in which bark showed the maximum capacity for adsorption of heavy metals followed by needles, cones, and wood. For example, variety of cadmium uptake rates g/mg for pine bark; needles and cones, respectively, were compared in which it showed that the cadmium uptakes of juniper wood had lesser adsorption rate than the adsorption rates of barks. Based on various adsorbents being tested, it was found that bark was more effective in filtration media than the wood. (ii) Based on case studies, it depicts that the adsorption levels of Pinus radiata bark improvised so much that it helped in the removal of uranium with acidified formaldehyde treatments (Freer et al. 1989). Formaldehyde treatment is given to Pinus pinaster barks for the removal of cadmium and mercury, after optimization of the treatment. (iii) Eucalyptus globulus bark for the chromium absorption showed the high capacity of encouragement toward the cation adsorption. Pretreatment of bark with amino-containing groups and ammonia solutions as urea is another way of formaldehyde treatments. These treatments lead to decrease in the release of tannins and increment in the metal adsorption (Sarin and Pant 2006). In a case study, ammonia-treated Pinus sylvestris bark was compared with the adsorption results of untreated barks, and higher values of heavy metals were obtained in the case of treated barks (Khokhotva 2010). Development of amino groups on the surface of the bark due to the urea interaction with carbonyl and carboxyl groups adds to the formation of additional active centers of metal binding (Yadav et al. 2014). (iv) Eucalyptus barks were used as an adsorbent for pollution control check like dyes and heavy metals ions. The adsorption rate of eucalyptus bark was checked toward chromium, copper, and nickel in a step-by-step process for certain parameters. The obtained results showed that the retention capacity increased with contact pH, time, and initial metal ion concentration, but there was a decrement with temperature (Saliba et al. 2002).

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(v) Guava bark was collected from garden and locally available in Nagpur. The concentrations of nickel were obtained by atomic absorption spectrophotometer (AAS) (Salem et al. 2001). (vi) The moringa bark has bioremedial and healing properties which are useful in chemical modification enhancement. The plant has clarifying and nutritional property as the fruits and leaves are taken up for health purposes and the bark has healing and bioremedial functions which help in the enhancement of chemical modification. Using bark powders and these dried and modified leaves, the phytoremediation tests were performed to identify the major polluted lakes. The polluted lake samples were subjected to biosorption. Atomic absorption spectroscopy (AAS) values and thin layer chromatography (TLC) analytical techniques were used to detect the reduction in heavy metals (Jiraungkoorskul and Jiraungkoorskul 2016). Table 13.2 shows the details of various barks and their metal quenching ability. The most studied metals were cadmium and copper followed by chromium and zinc. Nickel and lead were also studied to some extent (Nurchi and Villaescusa, 2008). Mercury and iron were studied rarely, and there is only one study with vanadium and uranium (Alloway 2012). Table 13.2 Details of different tree barks and their metal quenching properties Sl. no. 1.

Various tree barks Pinus radiata

2.

Eucalyptus

3.

Guava

It has high content of tannins, starch, and proteins. It is used in treating diarrhea in children and is also used medicinally as a powerful astringent

4.

Moringa oleifera

Ancient medicine shows it has healing characteristics

5.

Cassia abbreviata

6.

Acacia nilotica

The stem bark contains histamine or chemically related substance responsible for lowering high blood pressure It contains a lot of tannins contributing to its many medicinal uses, also acts as a powerful astringent

Functions It is a versatile timber. When treated properly, it can be used for many exposed structural and non-structural applications It is used for its antiseptic and astringent effect

Metal chelation property Removal of uranium with acidified formaldehyde treatments

Adsorption of pollutants such as heavy metal ions and dyes. Removal of copper, chromium, cadmium, and nickel Nickel concentrations were analyzed using atomic mass spectrometry It was observed to have good metal chelation properties Biosorption of heavy metals was found for the bark-treated water samples. Reduction in heavy metals such as nickel and cadmium It is found to have good metal quenching properties. Reduction in metals such as lead and mercury Among various species, Acacia nilotica has the highest metal quenching properties

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13.4.3 Biosorption: The Probable Mechanism of Heavy Metal Reduction via Tree Barks The use of biological content for the removal of metal and nonmetal from aqueous solution is known as biosorption. Basically biosorption is an adsorption process where a biological component is used as an adsorbent. There are three factors on which biosorption is dependent on, i.e.: 1. Adsorbent system Here temperature, pH, and the time of adsorption are the most essential aspects wherein they can change the binding site shape and increase the adsorption rate. 2. Metal factors Certain heavy metal can bind with biosorbent with no specificity, where for some metals it is specific. Also larger ions imply easy to polarize, and therefore the metal atom can easily be separated and be adsorbed by the bark. 3. Adsorbent-related factors These are due to chemical and anatomical properties of the bark. The adsorption capacity of the bark is affected by the amount of absorbent, but the varied components of the bark are hardly considered into the account (Sen et al. 2015). So as a heavy metal comes in contact with the bark, there are some changes in the binding sites which help in better adsorption and thereby removal of heavy metal from the original source. Therefore, for effective adsorption, cellular products and living and nonliving biomass are often used, but their issue still remains unanswerable (Bajpai et al. 2017).

13.5

Ongoing Work (Unpublished Data)

In the current study, a combined system of activated carbon and bionanoparticles are used for water purification. In the current study silver bionanoparticles were prepared from Moringa bark and Bacopa whole plant extracts and were characterized via ultraviolet-visible spectroscopy, Fourier-transform infrared spectroscopy, scanning electron microscopy, dynamic light scattering, energy dispersive X-ray analysis, X-ray diffraction, gas chromatography, mass spectroscopy, and atomic force microscopy. Further cellulose acetate bionanocomposite membranes were prepared and characterized. Later a small-scale water purifier was fabricated in which activated carbon particles was the first line of purification and cellulose acetate bionanocomposite membrane acted as the second line of purification. The water sample testing after purification revealed removal of coliforms and heavy metals to a greater extent as of in permissible limits. Other water quality parameters were also improved depicting the efficacy of the water purifier.

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Conclusions

For pollution monitoring studies, barks act as promising substrate as they provide large specific area for contact, and due to their physical and chemical structure, they are able to retain the pollutants. Among the parts of a plant, barks have the best potential to retain chemicals when compared to other parts. However, there is a need to study and develop new knowledge regarding the uptake and deposition on barks and, in addition, in standardizing sampling and analytical procedures. The use of barks and biomass for the removal of heavy metals has gained interest over the last decades. The adsorption capacities of various barks are comparable to that activated carbon for that of heavy metals such as zinc, cadmium, copper, and chromium. Barks of Camellia japonica, Pinus ponderosa, Quercus velutina, and Moringa oleifera had the maximum potential in the removal of chromium, copper, and mercury, respectively. Further focus should be on structural complexity, heterogeneity, and associated composition of the bark, and further studies on mechanistic properties of coordination chemistry, binding sites, oxidation states, and speciation of metals should be done to maximize the efficiency of heavy metal reduction from barks of a plant (Sen et al. 2015). Acknowledgments The authors thank the management of Dayananda Sagar College of Engineering for the encouragement and Mr. Yashas D for taking out time reading and giving certain suggestions. Note: The figures used in the chapter were created and modified by the authors.

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Saliba R, Gauthier H, Gauthier R, Petit-Ramel M (2002) The use of eucalyptus barks for the adsorption of heavy metal ions and dyes. Adsorp Sci Technol 20(2):119–129. https://doi.org/10. 1260/026361702320360540 Sarin V, Pant KK (2006) Removal of chromium from industrial waste by using eucalyptus bark. Bioresour Technol 97:15–12. https://doi.org/10.1016/j.biortech.2005.02.010 Şen A, Pereira H, Olivella MA, Villaescusa I (2015) Heavy metals removal in aqueous environments using bark as a biosorbent. Int J Environ Sci Technol 12(1):391–404. https://doi.org/10. 1007/s13762-014-0525-z Tchounwou PB, Yedjou CG, Patlolla AK, Sutton DJ (2012) Heavy metals toxicity and the environment. EXS 101:133–164. https://doi.org/10.1007/978-3-7643-8340-4_6 Vazquez G, Gonza’lez-A’ lvarez J, Freire S, Lo’pez-Lorenzo M, Antorrena G (2002) Removal of cadmium and mercury ions from aqueous solution by sorption on treated Pinus pinaster bark: kinetics and isotherms. Bioresour Technol 82:247–251. https://doi.org/10.1016/S0960-8524 (01)00186-9 Velma V, Vutukuru SS, Tchounwou PB (2009) Ecotoxicology of hexavalent chromium in freshwater fish: a critical review. Rev Environ Health 24(2):129–145. https://doi.org/10.1515/ REVEH.2009.24.2.129 Wang J, Chen C (2006) Bio sorption of heavy metals by Saccharomyces cerevisiae: a review. Biotechnol Adv 24(5):427–451. https://doi.org/10.1016/j.biotechadv.2006.03.001 Wastewater Management – A UN-Water Analytical Brief Wilson DN (1988) Association cadmium. Cadmium-market trends and influences; London. In: Cadmium 87 proceedings of the 6th international cadmium conference, 1988, 9–16 Yadav JSS, More TT, Yan S, Tyagi RD, Surampalli RY (2014) Extracellular polymeric substances of bacteria and their potential environmental applications. J Environ Manage 144. https://doi. org/10.1016/j.jenvman.2014.05.010 Zhang J, Mauzerall DL, Zhu T, Liang S, Ezzati M, Remais JV (2010) Environmental health in China: progress towards clean air and safe water. Lancet 375:1110–1119. https://doi.org/10. 1016/S0140-6736(10)60062-1

Chapter 14

Environmental Effects and Microbial Detoxification of Textile Dyes Zahid Maqbool, Habibullah Nadeem, Faisal Mahmood, Muhammad Hussnain Siddique, Tanvir Shahzad, Farrukh Azeem, Muhammad Shahid, Saima Muzammil, and Sabir Hussain

Contents 14.1 14.2 14.3 14.4

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . History, General Structure, Classification, and Properties of Textile Dyes . . . . . . . . . . . . Entry and Presence of Dyes in Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effects of Textile Dyes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.1 Effects on Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.2 Effects on Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.3 Effects on Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.4 Effects on Microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.4.5 Effects on Human Health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5 Microbial Detoxification of Dye Contaminated Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.1 Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.2 Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.5.3 Consortia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

290 291 294 296 296 297 298 299 300 301 301 306 310 312

Abstract Textile dyes are discharged into environment as colored wastewater along with other recalcitrant compounds. Dyes are resistant towards breakdown and exist in environment for longer period due to their fused aromatic structure. The presence of azo dyes in environment causes aesthetic complications and also poses severe threats to aquatic life. Microbial communities and plants are also badly affected by the presence of dyes in environment. Moreover, long-term exposure of dyes causes

Z. Maqbool · F. Mahmood · T. Shahzad · S. Hussain (*) Department of Environmental Sciences and Engineering, Government College University, Faisalabad, Pakistan H. Nadeem · M. H. Siddique · F. Azeem · M. Shahid Department of Bioinformatics and Biotechnology, Government College University, Faisalabad, Pakistan S. Muzammil Department of Microbiology, Government College University, Faisalabad, Pakistan © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 289 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_14

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detrimental diseases in human beings. This chapter discusses the structure, classification, properties, and ecotoxicology of textile dyes extensively consumed for dyeing purposes in textile industries. Keeping the toxicology of dyes in view, removal of these dyes from environment is a worldwide burning issue nowadays. Several physical, chemical, and biological approaches are conventionally used for treatment of textile wastewater prior to their discharge. However, this chapter only focuses on microbiological detoxification of azo dyes from textile wastewater. Several microbial strains and plants have the ability to degrade various azo dyes in aqueous media. However, here we will discuss potential of bacterial and fungal species as well as their consortia for detoxification of azo dyes. Keywords Textile dyes · Environmental pollution · Microbial detoxification · Azo dyes · Aromatic compounds · Contamination · Wastewater treatment

14.1

Introduction

Industrial development is an important factor for economic growth of any country (Sarayu and Sandhya 2012). Among different industries, textile industry is an important sector which includes the manufacturing of threads and yarns by god-gifted resources, e.g., cotton, wool, jute, and silk. The textile processing involves different steps including scouring, desizing, washing, bleaching, dyeing, and printing (Marcos et al. 2019). The dyeing step involves the use of natural or synthetic colors to impart color to fabrics. The dyes adhere to the substrate fabrics through covalent bond formation, adsorption, or by forming complexes (Garfield 2002). It has been observed that the synthetic dyes are more commonly used for dyeing purposes due to different advantages over natural dyes. For instance, synthetic dyes are more stable, easy to apply, form covalent bond with fiber, and have variety in color and shades. However, natural dyes are difficult to apply and have less variation in color and shades and more importantly less resistant towards washing. Among different synthetic dyes, azo dyes constitute the biggest class of dyes and are frequently used in fabric industry. According to some estimates, about 10,000 different azo dyes are commercially available with 7 x 105 metric tons annual production globally (Campos et al. 2001). It has been observed that a considerable quantity (15–50%) of the applied dye is vanished in coloring process and introduced in environment in form of wastewaters (Husain 2010). Textile mills use a large quantity of water which is added into the surroundings as wastewater polluted with dyes (Andleeb et al. 2010). Dyeing of 1 kilogram fabric releases approximately 40–65 liter wastewater into the environment (Manu and Chaudhari 2002). According to O’Neill et al. (1999), dye contents ranged in between 10 and 250 mg L
1 in some tested textile wastewaters. However, Pierce (1994) detected

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about 1500 mg L
1 of dyes in wastewater discharged from textile industries. Discharge of such colored wastewater containing large amounts of dyes is a serious risk to the environment because it causes water pollution, thereby rendering it unhealthy for organisms living in water (Chanyal and Agrawal 2017). Addition of azo dyes in water not merely causes aesthetic problems but also reduces sunlight interception and ultimately reduces the biomass production capacity of aquatic plants (Roy et al. 2010). Several dyes and their by-products are cancer causing and mutagenic which are detrimental for living organisms (Carneiro et al. 2010). Using textile wastewater for agricultural production is common in developing countries which might change biochemistry of soil and activities of microorganisms and their enzymes (Topac et al. 2009; Rehman et al. 2018). Therefore, due to common use and hazardous effects of azo dyes, the detoxification of textile wastewater is a one of the important issue worldwide. Although various physical and chemical techniques are being adopted to detoxify textile wastewater, they are either costly or not effective for complete management of textile wastewater (Maqbool et al. 2018). Moreover, these techniques have low adaptability and cause waste disposal problems and are, therefore, considered unsuitable for handling of textile wastewater (Asgher et al. 2008). Alternatively, biotic approaches are considered more environmental friendly, efficient in elimination of dyes from wastewater along economical operation and maintenance cost (Yang et al. 2011; Mishra and Maiti 2018). Although fungi are also able to detoxify dyes, bacteria are preferred because of easiness in handling fast growth and shorter life cycles (Elisangela et al. 2009). Several microbial strains including bacteria and fungi are identified and used for treatment of synthetic and real textile wastewater (Imran et al. 2014, 2016; Abbas et al. 2016; Maqbool et al. 2016, 2018). This chapter highlights the environmental effects of azo dyes on abiotic and biotic components of ecosystem together with human beings, animals, plants, and microorganisms. Moreover, the detoxification of the azo dyes loaded textile wastewaters using microbe-based processes has also been focused in this chapter.

14.2

History, General Structure, Classification, and Properties of Textile Dyes

Textile dyes are generally divided in two major classes, i.e., natural and synthetic dyes. The use of natural dyes started in prehistoric times to make surrounding world colorful and attractive. Up to 15,000 BC, a naturally occurring material named ochre was utilized by people to paint caves in ancient times. This pigment contained ferric oxide and ranged in color from yellow to deep orange or brown. By 2600 BC, madder plants (which contained a dye alizarin) were being used to color the clothes that were used to wrap Egyptian mummies (Garfield 2002). Products of different living organisms including plants, insects, and fungi particularly Monascus were

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utilized as natural dyeing products. Few natural dyes which are still in use are carmine, hematoxylin, and orcein. The production of dyes and their exploitation to color textile products was initiated around 7000–2000 BC (Grierson 1989; Van der zee 2002). Due to advancement in textile industries and limited range of colors in natural dyes, the chemists moved towards the synthesis of artificial dyes with variety of shades and wide range in bright colors. Synthetic dyes are derived from different organic or inorganic substances. The primary artificial dye, Mauve, was manufactured by William Henry Perkin in 1856. It was used for coloring silk products. Additionally, it was resistant towards color fading against washing and light effects. This finding paved the way for manufacturing of dyes with variety of colors. This trend was gradually amplified in 1865 because of Kekule’s discovery. Therefore, during the twentieth century, synthetic dyes had almost entirely replaced the natural dyes (Welham 2000). The synthetic dyes are very frequently used nowadays. The synthetic textile dyes are mostly aromatic and diverse in their structure. The dye molecules are generally comprised of two groups: (1) chromophore, i.e., electron acceptor which gives specific color to the fabric through absorption or reflection of light, and (2) auxochrome is an electron donor and gives intensity to the color and binding affinity to the applied fiber. Methine group (–CH¼), carbon–sulfur (¼C¼S; CS–S–C), azo group (–N¼N–), ethylene group (¼C¼C¼), carbon–nitrogen group (¼C¼NH; –CH¼N–), nitroso group (–N¼O; ¼N–OH), ethylene group (¼C¼C¼), and nitro group (–NO2; –NO–OH) are few examples of chromophore group, and sulfonate (SO3H), amino (NH2), hydroxyl (OH), and carboxyl (COOH) are examples of auxochrome (Suteu et al. 2012). The presence of these diverse types of groups improves the structure of dye molecules and gives characteristics to a specific dye (Zollinger 1987; Al-Ahmed 2014). For instance, replacement of a group with sulfonic acid makes the dye molecule more soluble in water (Khan et al. 2014). On the basis of chemical structure, synthetic dyes are categorized into several classes such as azo, triarylmethane, anthraquinone, phthalocyanine, and sulfur. Azo and anthraquinone dyes are well known and make up almost 65–75% of the dyes used in textile industry. The dye compounds having azo (N¼N) group in their structure are named as azo dyes. Existence of aromatic ring and azo group in dyes enables them to bind with fiber and show resistance against their degradation (Xu et al. 2006). Based on numerous azo groups, azo dyes are categorized into mono (containing one azo group)-, di (two azo groups)-, tri (three azo groups)-, and poly (more than three azo groups)-azo dyes. It has been observed that mono-, di-, and tri-azo dyes are preferably used over poly-azo dyes. Benzene and naphthalene rings are generally attached with azo group of dye and can be the part of analyzable aliphatic or aromatic heterocycles. These groups give color to the applied material along dissimilar shades and strength. Azo dyes give visible spectrum of red, orange, and yellow color. Anthraquinone dyes are considered as second biggest class of textile dyes that give visible spectrum of blue, green, and violet colors (Yaseen and Scholz 2019).

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Based on dye alienation in aqueous media, dyes are classified into basic, disperse (non-ionic), acidic, direct, and reactive dyes (Sivasakthivelan 2013). Basic dyes are cationic and have positive charge. Therefore, they are applied to those materials which are negatively charged, e.g., acrylic fibers. Reactive, direct, and acidic dyes are anionic in nature. They are applied to cellulose fiber, for example, cotton. It has been reported that, due to higher solubility of these dyes, conventional treatment technologies have no impact on them and they can easily pass from the water (Mathew et al. 2019). The general properties of few types of synthetic dyes have been described below. Acid dyes are anionic in nature and highly water soluble and have molecular weights lower than direct dyes. These dyes are primarily applied for dyeing of different materials like modified acrylic and textile fibers, wool, and nylon. The fixation of acid dyes includes the fixation of sulfonic group in acid dyes with the amide groups of the fiber (Ollgaard et al. 1998). Direct dyes are also anionic and have higher solubility than acid dyes in aqueous media. This class of dyes has thin and horizontal molecules able to align with cellulose through van der Waals forces. Furthermore, their binding to the fiber occurs due to deposition in cavities. The chief use of direct dyes includes dyeing of papers, leather, and cellulosic fibers (Ollgaard et al. 1998). Reactive dyes are alike in structure to direct dyes excluding the existence of reactive groups. These reactive groups fix via covalent bond with the hydroxyl, sulfhydryl, and amino groups of the fibers. These dyeing products show excellent resistance against washing. The color of the reactive dyes is firmly fixed and has a long life due to its chemical reaction with other compounds. The key usage of reactive dyes includes dyeing of cellulosic materials, for example, yarn and rayon (Hunger 2002). Basic dyes are cationic, water soluble, and frequently utilized for dyeing of polymers bearing -ve charges, for example, modified polyester, modacrylic, and acrylic. However, basic dyes are principally applied for dyeing of papers and polyacrylonitrile fibers (Ollgaard et al. 1998). Disperse dyes are non-ionic, insoluble in water, and generally used for dyeing purposes in slightly acidic conditions via aqueous dispersions. They are synthesized by grinding until a relatively fine powder is obtained along the presence of dispersing agents. The key use of disperse dyes is in dyeing of cellulose acetate polyester, nylon, and polyester blends (ETAD 1995; Hunger 2002). Mordant dyes contain mordant which is a metal compound used for fixation on fabrics. They normally have no charge and are used to synthesize metal complex dyes that are water insoluble and show precipitation on applied fiber. The principal use of these dyes includes coloring of furs, leather, and wool (Ollgaard et al. 1998). Solvent dyes are non-ionic with low water solubility and have molecular weights higher than disperse dyes. They are mainly utilize for coloring mineral oil products, waxes, plastic, inks, and fat (Ollgaard et al. 1998). Azoic dyes are a widely used class of synthetic dyes comprising of azo groups linked to at least one or more aromatic rings. Azoic dyes are commonly utilized to give bright red shades in printing and dyeing. Due to the presence of aromatic group,

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azo dyes are resistant towards degradation and retain themselves to the fiber for long period of time (Xu et al. 2006). Nitro dyes are products of phenols comprising of one nitro group with the hydroxyl group. This type of dyes consists of two or more aromatic rings (naphthalene and benzene). The nitro dyes are mostly used to dye wool. Vat dyes involve polycyclic particles and depend on the quinone structure rendering them mostly insoluble. The source of these dyes is carbazole, indigo, and anthraquinone. The principal application of vat dyes includes coloring of cotton, silk, nylon, linen, wool, rayon, and wool. The dyes made through this way have light fastness and high washing resistance. Sulfur dyes give deep shades to materials and show high resistance against washing but low resistance to sunlight. They are generally used for dyeing of rayon, cotton, and linen. The disadvantage of sulfur dyes is the black color of these dyes which makes the fabric fragile and also weakens the structure. They are used to color cotton in alkaline media with Na2S as reducing mediator. Sulfur dyes are cheap and show virtuous resistance to acids, light, and washing.

14.3

Entry and Presence of Dyes in Environment

Dyes are released in the environment due to the dominant use of dyes in several industrial processes such as textile dyeing, paint and pigment manufacturing, pulp and paper processing, leather tanning, etc. For example, during textile dyeing processes, a significant amount of the applied dyes is lost because they do not adhere with fiber and ultimately end up as wastewater which is discharged in to the environment (Stolz 2001; Manu and Chaudhari 2003). According to some estimates, the loss of applied dyes is 10–15% for azo dyes and 50% for reactive dyes (Stolz 2001; Chen 2002; Zille et al. 2004). According to an estimate, about 200 liters of water is consumed for 1 kilogram textile product which generates large quantities of wastewater. The wastewater generated from dyeing industries comprises of several recalcitrant compounds including dyes, metals, salts, detergents, and other pollutants (Imran et al. 2015a). The configuration of wastewater originating from textile industries is influenced by the dying process and equipment used, type of fiber, type of dye applied, weight of the fabric, and the trends in fashion (Brik et al. 2006; Kehinde and Aziz 2014). Earlier scientists have studied the characteristics of actual textile wastewater generated from various industries in different countries. Although some of the wastewaters were found within the safe permissible limit, diverse range of recalcitrant compounds was found in textile wastewater. The characteristic of different wastewater samples studied by several researchers has been described in Table 14.1. The concentration of dyes in wastewater is reported to cross the permissible safe limits. For example, Laing (1991) reported 10–50 mg L
1 dyes in textile wastewater. Likewise, Shelley (1994) reported that the reactive dyes were present at concentration of 60 mg L
1 in textile wastewater. Diverse studies reported 100–200 mg L
1 of dyes in wastewater produced by textile industries (Gahr et al. 1994), ranging between 600 and 800 mg L
1 (Vandevivere et al. 1998) and

35–45 21–62

238–2430 150–10,000 487–768 268–1275 150–10,000 487–1120 1675–1850 150–10,000 1500–30,000 150–12,000

41–323  160 652.8–800 1500–12,000 1000–1500

CODa (mg L
1) 59–516

117–786 100–4000 140–320 60–450 100–4000 140–420 560–620 100–4000 80–6000 80–6000

BODa (mg L
1) 20–207 46.9–58.5 28–126  30 151.24–299.1 80–6000 300–500

1108–1907 448–485

592–1696

2210–6020

EC (μS/cm) 1700–19,820

a

100–5000 15–8000 15–8000

100–5000

49–471 100–5000 1108–1547

 30 100–336 15–8000 200–400


1

TSS (mg L ) 120–310 872.85–1282.4

a

1056–7130 1500–6000 764–3380 152–1011 1800–6000 685–1338 4271–5240 1800–6000 2900–3100 2900–3100

TDSa (mg L
1) 2758–7744 984–1148 2218–3012  250 1856–4356 2900–3100 8000–12,000 1.2–7.0

0.11–0.5 1000–6000

1.95–5.81

DOa (mg L
1) 6.7–7.0

References Maqbool et al. (2016) Ahmed et al. (2011) Khan and Guha (2012) Tafesse et al. (2015) Roy et al. (2010) Ghaly et al. (2014) Eswaramoorthi et al. (2008) Hannan et al. (2011) Hussein (2013) Sultana et al. (2013a) Sultana et al. (2013a) Kalra et al. (2011) Sultana et al. (2013b) Sultana et al. (2013b) Upadhye and Joshi (2012) Kehinde and Aziz (2014) Turhan and Turgut (2009)

Temp (temperature), COD (chemical oxygen demand), BOD (biological oxygen demand), EC (electrical conductivity), TSS (total suspended solids), TDS (total dissolved solids), DO (dissolved oxygen), mg L
1 (milligram per liter), μS Cm
1 (microsiemens per centimeter)

a

7.75–13.61 5.5–10.5 7.3–11 6.8–11 6–10 8.9–11.0 3.9–4.5 6–10 6.95–11.8 7–9

34.7–46.3 33–45

35–45

7.2–11.3 6.00–9.00 9.6–11.2 6–10 6–10

pH 8.84–10.1

27–35  37 40.5–43 35–45 35–45

Tempa ( C)

Table 14.1 Physicochemical properties of textile wastewater discharged from various textile manufacturing and dyeing industries

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7000 mg L
1 (Koprivanac et al. 1993). Likewise, Ghaly et al. (2014) found that contents of azo dyes in wastewater produced from textile dye houses ranged 10–250 mg L
1. Metals having density greater than 5 g cm
3 are also found in textile wastewater due to metal complexed dyes and dominant use of metals in dyeing process as additives. Among the different metals, chromium, zinc, iron, mercury, and lead are most widely reported to be present in textile wastewater (Adinew 2012; Hussein 2013). Hence, on the basis of this factual information, it can be concluded that dyes are the key elements present in wastewater originating from several industrial sectors especially the textile industry. Larger molecular sizes, unusual bonding, replacement with bromine or chlorine, and extraordinary stable and toxic nature are important dye properties which make them highly persistent in natural environments (Imran et al. 2015a). Moreover, multiplex aromatic ring along delocalized electrons, conjugated double bonds, and stability towards light degradation also enables them to persist in the environment (Maddhinni et al. 2006). For example, half-life of hydrolyzed Reactive Black 19 is 46 years (Hao et al. 2000). Furthermore, synthetic origin and resistance towards fading also enable dyes to exhibit high stability towards degradation in natural environment (Pandey et al. 2007; Bafana et al. 2008). The azo linkage along with sulfonate group has a strong electron withdrawing power and results in creation of electron scarce particles which decrease vulnerability of azo dyes to oxidative reactions and enhance their persistence and recalcitrant effects (Barragan et al. 2007). Because of stability, persistent nature, and carcinogenic and mutagenic characteristics, dyestuffs pose several toxic effects on environment which have been discussed in the following chapter.

14.4

Effects of Textile Dyes

Several researchers have studied phytotoxicity, zoo-toxicity, and microbicidal/ microbiostatics impacts of azo dyes for determining eco-toxicity of different azo dyes commonly found in textile wastewaters (Zablocka-Godlewska et al. 2015). Generally, eco-toxicity of textile wastewaters is determined for two major reasons. Firstly, for risk assessment of dye contaminated wastewaters discharged into surrounding water and soil resources. Addition of such wastewaters in water and soils may have negative impacts on biological components of such environments (Imran et al. 2015b). Secondly, toxicity evaluation helps determine the efficiency of different wastewater treatment technologies.

14.4.1 Effects on Environment The discharge of colored wastewater containing dyes into environment has unspecified fate and causes serious ecological concerns (Dong et al. 2003). For

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example, the existence of dyes in water resources creates nasty color and odor in receiving water as well as hinders the light dispersion, water transparency, and gas solubility which results in reduced biomass of aquatic plants (Govindwar et al. 2014; Liu et al. 2015). Moreover, existence of dyes in wastewater also reduces the concentration of dissolved oxygen which resultantly increases biological oxygen demand and chemical oxygen demand of water bodies (Saratale et al. 2009; Arar et al. 2014). It is observed that the discharge of untreated dye loaded wastewaters into the lakes and rivers alters the pH and increases the alkalinity and salinity of water (Chonkar et al. 2000; Imran et al. 2014). Few dyes and their degraded by-products are carcinogenic and mutagenic, thus, causing sudden genetic changes in water ecosystem (Khan et al. 2014; Liu et al. 2015). The untreated dye wastewaters also produce toxic aromatic amines in oxygen deficient environment in receiving media (Murugesan and Kalaichelvan 2003).

14.4.2 Effects on Plants Using untreated textile wastewater to grow crops is a common practice in developing countries due to lack of sufficient water resources. Exposure of plants to untreated textile wastewater causes a number of hazardous impacts on growth and normal physiology of plants (Ayed et al. 2011). Plant growth is severely affected by exposure to dyes which decrease the germination, root elongation, and shoot growth of the plants (Puvaneswari et al. 2006). It has been reported that benzidine causes disturbance in calcium level and DNA synthesis which lead to retardation of root cells in maize (Zaalishvili et al. 2000). Jha et al. (2015) found 20% reduction in germination rate of Phaseolus mungo and Triticum aestivum after their exposure to RB8 dye (100 milligram per liter). Similarly, Tee et al. (2015) noticed the toxicity symptoms in cattail plants exposed to Acid Orange 7 dye. Lade et al. (2015b) recorded 60 and 70% decrease in germination of Sorghum vulgare and Phaseolus mungo, respectively, in response to their exposure to 50 mg/L of Reactive Black dye. Moreover, shoot and root length was also found to decrease significantly. Similarly, direct red 81 induced enormous decline in seed germination and biomass development of Sorghum vulgare and Phaseolus mungo seedlings (Sahasrabudhe et al. 2014). In some other studies, exposure to Red HE7B and Brown 3RE resulted into a complete inhibition of seed germination of Triticum aestivum (Kalme et al. 2007; Dawkar et al. 2008). Like synthetic dyes, the dyes residues in raw textile effluents have also been found to suppress growth of different crop plants. Phugare et al. (2011b) reported that untreated effluents of a dying industry located in India significantly inhibited the seed germination of Triticum aestivum. Lade et al. (2012) also found that untreated textile wastewater showed inhibited germination of Sorghum vulgare (60%) and Phaseolus mungo (50%). Similarly, radicle length of plants was also shorter in case when seeds of both crops were soaked in untreated effluent.

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The dyes may inhibit the seed germination and plant growth due to their cytotoxic and genotoxic properties. Allium cepa have been most widely used test plant to study the gentotoxic and cytotoxic impacts of azo dyes. According to a study, when small bulbs of Allium cepa were exposed to 500 mg L
1 of Red HE3B dye, root growth of the plants was inhibited along with chromosomal aberrations with alteration in deoxyribonucleic acid of root cells (Phugare et al. 2011a). Moreover, mitotic index, total changes in chromosome, tail deoxyribonucleic acid, and tail length increased (Phugare et al. 2011a). Similarly, Waghmode et al. (2012) found decreased cell viability but enhanced mitotic index, chromosome breaks, and total number of alterations of Allium cepa in response to exposure to dyes (Waghmode et al. 2012). The cytotoxicity and genotoxicity of the dyes were further confirmed by using real textile wastewater containing dyes. Higher values of mitotic index, total number of alterations, tail deoxyribonucleic acid, and tail length were recorded when Allium cepa was exposed to the wastewater effluent of a textile industry located in India. Viability of roots cells of Allium cepa decreased upon exposure to textile wastewater (Lade et al. 2012). Thus, dyestuff existing in textile wastewater severely inhibits germination and retards crop growth. The extent of inhibition is variable depending upon the type of dye and its level in water.

14.4.3 Effects on Animals Discharge of wastewater loaded with azo dyes in freshwater creates the aesthetic problems and also affects the life of aquatic animals. Several researchers have studied the impacts of textile wastewater on aquatic animals (Meireles et al. 2018; de Oliveira et al. 2018). For example, Reactive Blue 15 dye was found toxic for embryo of fish models, bacteria, and tadpoles (de Oliveira et al. 2018). Moreover, the presence of dyes including Direct Blue 38 in water badly affects the detoxification and anti-oxidant processes, cellular energy allocation, oxidative damage, and acetylcholinesterase process during the initial life stages of zebra fish (Meireles et al. 2018). Daphnia magna is the most widely used model aquatic organism to study severe and prolonged effects of chemicals existing in aquatic environment (USEPA 1985). Data regarding toxicity evaluation of different dyes and real textile wastewaters using Daphnia magna as test organism suggested that untreated textile wastewater is highly toxic to aquatic organisms (Zablocka-Godlewska et al. 2014). Generally in such bioassays, mortality percentage, EC50 (concentration of dye effluent that causes 50% growth inhibition of tested organisms), and acute toxicity unit (TUa) of dyes are determined to assign toxicity class (I-V) to dyes (ZablockaGodlewska et al. 2014). As stated by ACE 89/BE 2/D3 final report commission of European communities, acute toxicity unit lower than 0.4 agrees to class 1 which is non-toxic, acute toxicity unit lower than 1.0 relates to class 2 which have lower toxicity, acute toxicity unit lower than 10 relates to class 3 which is toxic, acute toxicity unit lower than 100 relates to class 4 which is highly toxic, and acute toxicity unit greater than 100 relates to class 5 which is extremely toxic. Zablocka-

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Godlewska et al. (2012) reported EC50, acute toxicity unit, and toxicity class for a dye Evans blue (100 mg L
1) as 9.43  0.22, 10.6 and class IV, respectively. Przystas et al. (2012) also studied toxicity of Evans blue (50 mg L
1) to Daphnia magna and recorded 13.20 value of acute toxicity unit, and toxicity class was class IV. Franciscon et al. (2012) found complete mortality of Daphnia magna larvae upon their exposure to different dyes including Reactive Yellow 107, Reactive Black 5, Reactive Red 198, and Direct Blue 71 at concentration of 100 mg L
1 (Lade et al. 2015d). Nascimento et al. (2011) also stated elevated toxicity factor of Reactive Red 198 to Daphnia pulex. Parrott et al. (2016) also reported chronic toxicity of disperse yellow and Sudan Red G to fish. The actual wastewaters discharged without any treatment were also studied for assessment of their zoo-toxicity. Almost complete mortality of Artemia salina and Daphnia magna was noticed upon their exposure to wastewater released from textile industry located in State of Santa Catarina (Forgiarini and de Souza 2007). Bilal et al. (2016) also reported that exposure of Artemia salina and Daphnia magna to untreated textile wastewater of six different textile units located in Faisalabad Pakistan exhibited mortality ranging from 0% to 100%. Similarly, textile wastewater originating from dyeing industry located in South Korea was found toxic to larvae of Daphnia magna, i.e., 3.5 acute toxicity unit was recorded (Choi et al. 2014). Textile wastewater was also reported to change the genetic material of animal cells. For example, Gomaa et al. (2012) noticed that chromosome aberrations excluding gaps and chromosome deletion of mouse cells increased from 21 to 154 and 2 to 5, respectively, on exposure with dye effluents, whereas the mitotic index value decreased from 190 to 85. Likewise, Fernandes et al. (2015) observed testicular toxicity in mouse due to disperse red 1 dye. The improved frequency of sperm with unusual morphology and greater extent of deoxyribonucleic acid destruction was also noticed in testis cells. Based on these facts, it can be concluded that the presence of dyes in aquatic systems is a serious health risk for aquatic life. Therefore, discharge of untreated textile effluent in surrounding water bodies must be prohibited along with necessary treatment before their release in to environment.

14.4.4 Effects on Microorganisms Like animals and plant, the activities and composition of microbial communities are affected on exposure to dyes which ultimately result in significant alteration in ecological dynamics (Imran et al. 2016). Azo dyes induce oxidative stress and promote membrane damage in microorganisms. The harmful effects of dyes on microbial population result in loss of membrane integrity, inhibited electron transport chain, enhanced membrane permeability, and oxidation of polyunsaturated fatty acids. Moreover, higher content of dyes negatively affects the development and metabolic processes which consequently reduce the growth of naturally existing indigenous soil microorganisms. While studying the impact of different dyes on microbial community composition based on phospholipid fatty acids,

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Imran et al. (2015a) observed that Reactive Black 5 and other dyes significantly affected the phospholipid fatty acids of fungal and bacterial communities. A number of microbial species have been used as bioindicator to study the injurious impacts of azo dyes in environment. For example, Saratale et al. (2015) noticed the noxious impacts of Green HE4BD (400 mg L
1) on soil microorganisms belonging to different genera such as Rhizobium, Acinetobacter, Pseudomonas, and Proteus. They found that exposure of soil microbes to Green HE4BD significantly inhibited their growth. Zones of growth inhibitions were observed in the range of 5.0–7.5 mm for different microorganisms on nutrient agar medium. Dyes were also found to inhibit growth of nitrogen fixing microorganisms like Azotobacter vinelandii and Azotobacter chroococcum (Chung et al. 1998; Pozo et al. 2000). Respiration inhibition of microorganisms is another indicator of stressful environment. Orange II and Reactive Black 5 caused up to 50% and 30% respiration inhibition of Escherichia coli, respectively (Ambrosio and Campos-Takaki 2004). Likewise, textile wastewater originating from a textile dyeing and printing Egyptian company was also found to suppress the growth of Escherichia coli, and a clear zone of inhibition was detected (Gomaa et al. 2012). Similarly, textile effluents from Ksar Hellal (Tunisia) suppressed growth of Saccharomyces cerevisiae by 59% and of Bacillus cereus by 38% compared to control (Benzina et al. 2013). So, it can be concluded that untreated textile wastewater, if used for crop production, can seriously inhibit crop growth and affect beneficial microorganisms present in agricultural soils. Hence, textile effluent must be treated with suitable technology to improve the quality of such waters and then released into environment.

14.4.5 Effects on Human Health Long-term exposure of human beings to textile wastewater causes serious health hazards. For example, reactive brilliant red dye binds with human serum albumin, inhibits its function, and causes precipitation and structural changes in protein. Numerous infections and diseases in human beings like skin irritation, jaundice, and tissue necrosis are linked with the contact to dyes (Chan et al. 2012). Few reactive dyes cause frequent bladder cancer in people working in dye production industries (Rehn 1895; Watharkar et al. 2015). Not only the dyes but also the aromatic amines produced due to partial breakdown of dyes have been linked with elevated level of acute and chronic toxicity resulting into anemia, dysfunction of the kidney and liver, hematuria, and allergy (Li and Song 2010). It is also likely that carcinogenic amines of azo dyes condensed in remains of water bodies in the ecosystem cause complaints to human health, i.e., hemorrhage, nausea, and ulceration of the mucous membranes and skin, and also become a reason for acute injury to the liver, the kidney, central nervous system, reproductive system and the brain (Sarayu and Sandhya 2012).

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14.5

301

Microbial Detoxification of Dye Contaminated Wastewater

Due to high biological toxicity of textile wastewater, treatment before their introduction in environment is an emerging issue nowadays. During the last two decades, microbial detoxification of synthetic and actual textile wastewaters have showed more promising results and attracted global courtesy as a virtually feasible, economical, and ecosystem friendly technique (Maqbool et al. 2018). Therefore, various dye degrading microbes were identified for their potential to treat textile wastewater prior to discharge in environment (Imran et al. 2014; Chen et al. 2018; Das and Mishra 2019). Mostly the microbial detoxification of dye-contaminated water is an environment-friendly approach and reduces biological toxicity of dyes. However, in few cases, metabolites are produced during microbial degradation of dyes which are more toxic in nature. But these secondary metabolites can be further degraded in non-toxic by-products, or even sometime mineralization of the complete molecule takes place. Various microbial species of bacteria, fungi, yeast, and algae are found accomplished with breakdown of synthetic dyes (Waghmode et al. 2012; Jadhav et al. 2013, 2015; Jafari et al. 2013). The following sections primarily focus on the bacteria, fungi, and their consortia that have been explored principally for treatment/ detoxification of textile wastewater.

14.5.1 Bacteria Numerous bacteria having the ability to degrade azo dyes have been isolated and explored for potential treatment of wastewater coming from textile industries (Table 14.2). On the basis of bioassays based on phytotoxicity, zoo-toxicity, and microbicidal/microbiostatic analysis, several bacteria have been found to detoxify the dyes in real and synthetic wastewaters. For example, Pokharia and Ahluwalia (2016a) reported reduced toxicity and complete treatment of basic red 46 (100 mg L
1) using Staphylococcus epidermidis MTCC 10623 within only 6 h. Moreover, the treated wastewater improved germination index of Vigna radiata to 60% compared to 20% found in untreated wastewater. Similarly, Jadhav et al. (2013) recorded substantial improvement in rate of seed germination and radicle and plumule length of Phaseolus mungo, Triticum aestivum, and Sorghum vulgare when exposed to Remazol Orange (500 mg L
1) contaminated water that had been treated by Pseudomonas aeruginosa BCH before irrigation. According to Kumar et al. (2013), germination inhibition of Brassica nigra caused by disperse red F3B (100 mg L
1) can be removed by Enterococcus faecalis treatment. Microbial treatment of dye containing wastewaters also decreases their cytotoxic and genotoxic effects on plant cells, thus, decreasing dye-induced growth retardation of crop plants (Imran et al. 2016). Jadhav et al. (2011) found that treatment of

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Table 14.2 Biodegradation of different azo dyes using bacterial strains Sr. # 1

Dyes under studya AR128

2

Bacterial strain Bacillus endophyticus VITABR13 Shewanella sp. Strain IFN4

3

Bacillus sp. SG2

DR

4

Ganoderma lucidum, Coriolus versicolor

5

Bacillus subtilis

FYRD5GL, FRRDRBLS, FRRDGFL, FBRD3GR, FBRDGLN, FTSBLN CV

6

Lysinibacillus sp. RGS

RR

7

Proteus sp., Klebsiella sp.

LRD

8

Bacillus cereus

RR195

9

Enterococcus sp., Klebsiella sp.

RB19, DGD

10

Providencia rettgeri strain HSL1

RB172

11

CV

12

Pseudomonas aeruginosa, Clostridium perfringens, Proteus vulgaris Bacillus pumilus

13 14

Rhodococcus rhodochrous Stenotrophomonas sp.

TR ID

15

Acinetobacter calcoaceticus

DBMR

16

RR4E8Y5

17

Providencia spp. RMG1, Bacillus spp. RMG2 Bacillus sp. VUS

B3RE

18

Actinomyces sp.

ID

19

Enterobacter sp. strain VP-64

CV

20

Paenibacillus alvei MTCC 10625

BWNN

21

Pseudomonas sp.

ID

RB5, DR81, AR88

CR

References Prasad and Rao (2011) Imran et al. (2014, 2015c) Oak et al. (2016) Sadaf et al. (2013) Kochher and Kumar (2011) Saratale et al. (2013) Sethi et al. (2012) Modi et al. (2010) Gulati and Jha (2014) Lade et al. (2015a) Ali and Akthar (2014) Modi et al. (2015) Shah (2014) Rajendran et al. (2015) Ghodake et al. (2009) Gudmalwar and Kamble (2012) Dawkar et al. (2008) Rajendran et al. (2015) Hemapriya and Vijayanand (2014) Pokharia and Ahluwalia (2016b) Rajendran et al. (2015) (continued)

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Table 14.2 (continued) Sr. # 22

Dyes under studya AB113, BR46, DB151, DB2

25

Bacterial strain Pseudomonas aeruginosa, Pseudomonas putida Bacillus sp., Aeromonas hydrophila sp. Pseudomonas putida, Klebsiella ozaenae Bacillus sp.

26

Enterobacter aerogenes PP002

DB71

27 28

DR71 RB

29

Escherichia coli JM 109 Pseudomonas aeruginosa, Ochrobactrum sp. Pseudomonas sp. RA20

30

Enterococcus faecalis

AR27, RR2

31

Burkholderia sp.

Alizarin

32

Staphylococcus sp.

ID

33

Kocuria rosea MTCC 1532

MO

34

Lysinibacillus sp. AK2

MY

35 36

DO, DB, RG RR120, RB5, RY2, RO16

38

Pseudomonas aeruginosa Pseudomonas aeruginosa strain ZM130 Enterococcus casseliflavus Enterobacter cloacae Enterobacter spp.

CR

39

Serratia sp. RN34

RB5, RR120, RO16

40

Bacillus sp. VUS

OT4LL

41 42

Bacillus subtilis Shewanella oneidensis MR-1

Orange dye NGB

43 44 45

Shewanella decolorationis Pseudomonas sp. Pseudomonas putida

FARGR RB13 TY2G

23 24

37

PP CR, DR28, DB80 MR, NB

RB5

Orange II

References Falavarjani et al. (2012) Celia and Suruthi (2016) Shinkafi et al. (2015) Ezhilarasu (2016) Sudha et al. (2018) Jin et al. (2009) Kilic and Donmez (2012) Hussain et al. 2013 Handayani et al. (2007) Sharma and Sharma (2015) Rajendran et al. (2015) Parshetti et al. (2010) Anjaneya et al. (2011) Ahmed (2014) Maqbool et al. (2016) Chan et al. (2011) Prasad and Aikat (2014) Najme et al. 2015 Dawkar et al. (2010) Sha et al. (2014) Xiao et al. (2012) Xu et al. (2007) Lin et al. (2010) Srinivasan et al. (2011) (continued)

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Table 14.2 (continued) Sr. # 46

Bacterial strain Pseudomonas oleovorans B15

Dyes under studya CDB15

47

Escherichia coli

CR, DB38

48

Pseudomonas fluorescens Bacillus sp., Escherichia coli Halomonas elongata Escherichia coli Staphylococcus epidermidis MTCC 10623

ROM2R, RBM58, RYM4G, RB5

51

Providencia rettgeri Strain HSL1

RB172

52

Bacillus sp.

MR, OD

53

Enterococcus faecalis YZ 66

DR81

54

Pseudomonas aeruginosa BCH

RO

55

Bacillus lentus

RR120

56

Pseudomonas aeruginosa BCH

RR

57

Pseudomonas sp.

MO

58

Acinetobacter baumannii MN3

CR

59

Escherichia coli

RDR22

60

Bacillus sp. ACT2

CRM

61

Pseudomonas aeruginosa

RR2

62

Pseudomonas aeruginosa

RO

63

Clostridium perfringens

MAD

64 65

Clostridium bifermentans SL186 Citrobacter (LAJ01 and LAJ02)

RR3BA, RB5, RY3GP Amaranth dye, CR

66

Bacillus firmus (Kx898362)

DB14

67

Chromobacterium violaceum

MR, MO, amaranth dyes

49 50

References Silveira et al. (2011) Isik and Sponza (2003) Sriram et al. (2013)

MR

Eslami et al. (2016)

BR46

Pokharia and Ahluwalia (2016a) Lade et al. (2015b) Gayathri et al. (2014) Sahasrabudhe et al. (2014) Jadhav et al. (2013) Oturkar et al. (2011) Jadhav et al. (2011) Silveira et al. (2009) Kuppusamy et al. (2017) Chang and Kuo (2000) Gopinath et al. (2009) Bheemaraddi et al. (2013) Sarayu and Sandhya (2010) Morrison and John (2016) Joe et al. (2008) Schmidt et al. (2019) Neetha et al. (2019) Verma et al. (2019) (continued)

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Table 14.2 (continued) Sr. # 68 69

Bacterial strain Pseudomonas sp. Sphingomonas paucimobilis

Dyes under studya RB13 MR

70

Aeromonas hydrophila

CV, FG, BG, MG

71

Enterobacter sp.

CR, CV

72

Proteus sp. NA6

RY2

73

Shewanella putrefaciens AS96

AR88, RB5, DO3, DR81

74

Acinetobacter junii FA10

RR120, RO16, RY2, DB19, DR28

75

Aeromonas jandaei strain SCS5

MR

76

Bacillus sp. strain UN2

MR

References Lin et al. (2010) Ayed et al. (2011) Ren et al. (2006) Harry-asobara and Kamei (2019) Abbas et al. 2016 Khalid et al. (2008) Anwar et al. 2014 Sharma et al. (2016) Zhao et al. (2014)

a

AR128 (Acid Red 128), DB (Disperse Brown), AR88 (Acid Red 88), FR RDRBLS (Foron Red RDRBLS), FRRDGFL (Foron Rubine RDGFL), FBRDGLN (Foron Blue RDGLN), CV (Crystal Violet), RB19 (Reactive Blue 19), DGD (Dark Green Dye), LRD (Light Red Dye), RG (Reactive Green), TR (Toluidine Red), DB80 (Direct Blue 80), RR4E8Y5 (Reactive Red 4E8Y5), BWNN (Black WNN), AB113 (Acid Blue 113), BR46 (Basic Red 46), DB151 (Direct Blue 151), DB2 (Direct Brown 2), PP (Provisional Pink), ID (Indigo Dye), DR (Drimarene Red), MR (Methyl Red), NB (Navy Blue) RB172 (Reactive Blue 172), RB (Remazol Blue), FTSBLN (Foron Turquoise SBLN), MO (Methyl Orange), MY (Metanil Yellow), DR28 (Direct Red 28), DR71 (Direct Red 71), RR3BA (Reactive Red 3BA), FBRD3GRN (Foron Black RD3GRN), RY3GP (Reactive Yellow 3GP), OD (Orange dye), OT4LL (Orange T4LL), RR195 (Reactive Red 195), RR120 (Reactive Red 120), FG (Fuchsin Green), BG (Brilliant Green), MG (Malachite Green), DBMR (Direct Brown MR), DO3 (Disperse Orange 3), FARGR (Fast Acid Red GR), RB13 (Reactive Blue 13), RR2 (Reactive Red 2), TY2G (Tectilon Yellow 2G), CDB15 (Commercial dye B15), ROM2R (Reactive Orange M2R), RBM58 (Reactive Blue M58), RYM4G (Reactive Yellow M4G), DB14 (Direct Blue 14), DB71 (Direct Blue 71), RB172 (Reactive blue 172), RR (Remazol Red), RO (Remazol Orange), B3RE (Brown 3RE), RDR22 (Reactive Dye Red 22), CRM (Congo Red Mutant), NGB (Naphthol Green B), RO (Remazol Orange), DR81 (Direct Red 81), MAD (Mono-azo Dye), RY2 (Reactive Yellow 2), DR19 (Direct Red 19), RO16 (Reactive Orange 16), RB5 (Reactive Black 5), FYRD5GL (Foron Yellow RD5GL), CR (Congo Red), AR27 (Acid Red 27), DO (Direct Orange)

Remazol Red by Pseudomonas aeruginosa BCH increased cell viability of Allium cepa root cells to 86–95%, and a reduction in mitotic index, chromosome breaks, and total number of cells with alterations in A. cepa root cells was also observed. Franciscon et al. (2012) reported 40% mortality of Daphnia magna larvae by untreated Reactive Red 198 contaminated water, whereas no death of larvae was

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Z. Maqbool et al.

noticed after treatment with Brevibacterium sp. strain VN-15. Several other researchers have reported reduction in mortality rate and growth inhibition of Daphnia magna after bacterial treatment of dye containing wastewaters (Przystas et al. 2013; Shah et al. 2014; Lade et al. 2015b; Imran et al. 2016). Since the use of colored textile effluent for growing crops is a communal exercise in developing countries, therefore, the efficacy of bacterial treatment has also been evaluated by microbicidal/microbiostatic test. Saratale et al. (2015) reported that Reactive Orange 4 inhibited growth of soil microorganisms including Pseudomonas aeruginosa, Azotobacter sp., Cellulomonas biazotea NCIM-2550, and Lysinibacillus sp. RGS on nutrient agar plates. However, metabolites produced during dye degradation by Lysinibacillus sp. RGS did not suppress growth. Contrary to these observations, treatment of Acid Yellow 49, Acid Red 266, and Acid Blue 62 dyes containing wastewater by Bjerkandera adusta MUT 3060 showed increase in growth inhibition of green unicellular alga (Anastasi et al. 2011). These findings suggest that azo dye-induced inhibition in germination and growth of different crops, aquatic animals, and beneficial soil microorganisms could be minimized or eliminated after treatment with bacteria. In addition, treatment of dye-contaminated water reverts dye-induced alterations in genetic materials of organisms.

14.5.2 Fungi Fungi have also been studied for potential cure of industrial wastewater and reduction in biological toxicity of dyes because of their effective enzymatic system and large mycelial biomass which enhances adsorption of dyes. Several fungal strains are identified accomplished with degradation the dyes (Table 14.3). Previous studies reported efficient treatment of textile wastewater by several pure cultures of fungi and found reduction in phytotoxicity, zoo-toxicity, and microbicidal/microbiostatic effects, thus making them safe for discharge into surrounding water bodies and soils. For example, Laxmi and Nikam (2015) found that navy blue M3R (40 milligram per liter) dye containing wastewater enhanced germination (54–97%) of Triticum aestivum after it was treated with Aspergillus flavus. Moreover, plumule and radicle lengths were also statistically at par with control (distilled water treated) seedlings, indicating mitigation of negative/toxic impacts of dye on growth. Similarly, Waghmode et al. (2012) found that germination of S. vulgare was only 10% when exposed to 1000 mg L
1 of Rubine GFL solution. However, when treated with Galactomyces geotrichum MTCC 1360, metabolites produced from degradation were found non-toxic and showed 90% germination of Sorghum vulgare. Przystas et al. (2012) explored the noxious effects of Evan blue dye (80 mg L
1) and assigned class IV using Lemna minor as test plant. But, its toxicity decreased to class III after treatment with Pleurotus ostreatus (BWPH)

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Table 14.3 Biodegradation of different azo dyes using fungal strains Sr. # 1 2 3

Fungal strain Trichoderma tomentosum Aspergillus flavus Trametes versicolor

Dye under studya AR3R MG AO7, AB74, RR2, RB5

4

Trametes villosa SCS-10

AR357, AO142

5

RGFL

6 7 8

Galactomyces geotrichum MTCC 1360 Coriolopsis gallica Pleurotus ostreatus (BWPH) Galactomyces geotrichum

9

Fusarium spp.

10

P. ostreatus, P. chrysosporium, T. versicolor, and G. lucidum Pleurotus ostreatus MUT 2976, Trametes pubescens MUT 2295 Alcaligenes aquatilis 3c Permelia perlata Armillaria sp. F022 Pycnoporus sanguineus

11 12 13 14 15 16 17

RBBR, RB5, BBR, LGG EB RGFL, Azo dyes mixture

References He et al. (2018) Barapatre et al. (2017) Ramírez-Montoya et al. (2015) Ortiz-Monsalve et al. (2019) Waghmode et al. (2012) Daâssi et al. (2014) Przystas et al. (2012) Waghmode et al. (2011), (2012) Muthezhilan et al. (2014)

YMR, BM2R, BB, RSID, MMP, BMR, OM2R, O3R, BGR RB15, RRC4BL, LYC4GL

Kiran et al. (2019)

RBBR

Casieri et al. (2008)

SR6HBN Disperse dyes, SR24 AR27 OGA

Ajaz et al. (2019) Kulkarni et al. (2014) Adnan et al. (2015) Pointing and Vrijmoed (2000) Prasad et al. (2012) Kiayi et al. (2019)

RB5, RB19, AB74, Carmoisine dye

18

Pycnoporus cinnabarinus Saccharomyces cerevisiae ATCC 9763 Pleurotus ostreatus

19 20

Ganoderma sp. En3 Trametes hirsuta

RO16 AB, RB5, BR9, AB74

21

Ganoderma lucidum

RBBR, RB5

22

Dichomitus squalens, Ischnoderma resinosum, Pleurotus calyptratus Sclerotium rolfsii Aspergillus niger, A. flavus, A. fumigatus, Rhizopus sp., Penicillium sp. Irpex lacteus Marasmius cladophyllus

OG, RBBR

Abdulredha et al. (2014) Ma et al. (2014) Zapata-Castillo et al. (2012); Abadulla et al. (2000) Murugesan et al. (2007) Eichlerová et al. (2005)

RBR, ILMB, DB RR, RY

Ryan et al. (2003) Rohini et al. (2019)

RLBERA RBBR, OG, CR, MR

Kalpana et al. (2012) Ngieng et al. (2013)

23 24

25 26

BH3R, YFGR3B

(continued)

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Table 14.3 (continued)

a

Sr. # 27 28 29 30

Fungal strain Phlebia spp. P. formosus Penicillium simplicissimum Marasmius scorodonius

31 32

Aspergillus niger Candida tropicalis SYF-1

33

Peroneutypa scoparia

34 35

Coprinus cinereus Trametes pubescens strain i8

36 37

Phlebia brevispora strain TMIC33929 Pleurotus sajor-caju

38 39 40 41

Pleurotus eryngii Ceriporia lacerata Myrothecium verrucaria Podoscypha elegans

Naphthalene CR CR, MO, MR, CV OG, CR, DB15, DY27

42

Ganoderma cupreum

RV1

Dye under studya RB19, RB5 RB19, RB5 CV, MV, CB, MG MG, CV, CR, MG, RO16, RBBR RBBR, AR299 ARB, AS3R, RGKE4BD, RBRK2G, RY3RS, RBBKR AR97, AB210, AB282, AG16, AY42, AB7, AB193, AV54 OD, CV, MG RBV5R, DR5B, MG, RBB, RBBBG, AB158, IC CR, CV AR18), AB1, BB71, AR5,

References Bulla et al. (2017) Bulla et al. (2017) Chen and Ting (2015) Jeon and Lim (2017) Benghazi et al. (2014) Tan et al. (2019)

Pandi et al. (2019)

Lin et al. (2013) Rekik et al. (2019) Harry-asobara and Kamei (2019) Murugesan et al. (2006) Hadibarata et al. (2013) Wang et al. (2017) Sun et al. (2017) Pramanik and Chaudhuri (2018) Gahlout et al. (2013)

AR3R (Acid Red 3R), MG (Malachite Green), BM2R (Blue M2R), CR (Congo Red), AR27 (Acid Red 27), RO16 (Reactive Orange 16), RGFL (Rubine GFL), RLBERA (Reactive Levafix Blue ERA), EB (Evans blue), RBBR (Remazol Brilliant Blue R), SR6HBN (Synozol red 6HBN), OGA (Orange G Amaranth), RB19 (Reactive Blue 19), BB (Black-B), AR18 (Acid Red 18), BB71 (Direct Blue 71), AR5 (Alizarin Red 5), AO7 (Acid Orange 7), RR2 (Reactive Red 2), OD (Orange Dye), CV (Crystal Violet), BH3R (Blue H3R), YFGR3B (Yellow FG Red 3B), MG (Methyl Green), RSID (Red BSID), RBBR (Remazol Brilliant Blue R), AB (Acid Blue), BR9 (Basic Red 9), RB5 (Remazol Black 5), BBR (Bismarck Brown R), LGG (Lanaset Gray G), RBR (Remazol Brilliant Red), AB74 (Acid Blue 74), ILMB (Indigo Lancet Marine Blue), DB (Diamond Black), OG (Orange G dye), DB15 (Direct Blue 15), DY27 (Direct Yellow 27), RR (Reactive Red), RY (Reactive Yellow), MR (Methyl Red), RB5, YMR (Yellow MR), MMP (Magenta MP), BMR (Blue MR), OM2R (Orange M2R), O3R (Orange 3R), MO (Methyl Orange), RV1 (Reactive Violet 1), ARB (Acid Red B), AS3R (Acid Scarlet 3R), RGKE4BD (Reactive Green KE-4BD), RBRK2G (Reactive Brilliant Red K-2G), RY3RS (Reactive Yellow 3RS), RBBKR (Reactive Brilliant Blue KR), AR97 (Acid Red 97), AB210 (Acid Black 210), AB282 (Acid Brown 282), AG16 (Acid Green 16), AY42 (Acid Yellow 42), AB7 (Acid Blue 7), AB193 (Acid Blue 193), AV54 (Acid Violet 54), RRC4BL (Reactive Red C-4 BL), LYC4GL (Lemon Yellow C-4 GL), AR357 (Acid Red 357), AO142 (Acid Orange 142), RBV5R (Remazol Brilliant Violet 5R), DR5B (Direct Red5B), RBB (Remazol Brilliant Blue), RBBBG (Brilliant Blue BG), AB158 (Acid Blue 158), IC (Indigo Carmine), MV (Methyl Violet), CB (Cotton Blue), AR299 (Acid Red 299), Disperse dyes, SR24 (Solvent Red 24), RB5 (Reactive Black 5), AB1 (Acid Black 1), BGR (Brown GR)

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within 120 h. Many other fungal strains including Dichomitus squalens, Pleurotus ostreatus, Trametes pubescens, Galactomyces geotrichum MTCC 1360, Polyporus picipes (RWP17), Irpex lacteus, Galactomyces geotrichum, Ganoderma sp. En3, and Armillaria sp. F022 have been explored for their potential to mitigate hazardous impacts of dyes on development and biomass production of crops (Casieri et al. 2008; Kalpana et al. 2012; Przystas et al. 2012; Govindwar et al. 2014; Ma et al. 2014; Adnan et al. 2015). The fungal treatment has also been found to reduce cytotoxic and genotoxic impacts of azo dyes. For example, Waghmode et al. (2012) showed that the treatment of Rubine GFL (50 mg L
1) by Galactomyces geotrichum MTCC 1360 decreased mitotic index, chromosome breaks, and total number of cells with alterations in Allium cepa root cells compared to untreated water. Moreover, cell viability of Allium cepa root cells increased after treatment with Galactomyces geotrichum MTCC 1360. The zoo-toxic and microbicidal/microbiostatic effects of synthetic dyes present in textile wastewaters have also been studied using different capable fungal strain. For example, Daphnia magna larvae were exposed to untreated and treated Evans blue water to study acute toxicity unit as indicator. According to results, untreated synthetic wastewater had 13.20 acute toxicity unit, whereas it decreased to 2.40 acute toxicity unit after treatment with Pleurotus ostreatus (Przystas et al. 2012). Similarly, Zablocka-Godlewska et al. (2015) reported 13.20 acute toxicity unit and 3.33 acute toxicity unit for untreated and treated Evans blue synthetic wastewater, respectively. Moreover, EC50 increased from 9.43  0.22 to 30.03  2.07 which undoubtedly indicate significant decrease in toxicity of Evans blue in response to Daphnia magna larvae. In contrast, Congo red and Orange II (100 mg L
1) containing water treated with Irpex lacteus KUC8958 showed higher mortality of neonates of Daphnia magna (Choi et al. 2014). Similarly, Przystas et al. (2012) reported enhanced toxicity of metabolites of Evans blue to neonates of Daphnia magna than parent dye molecule. Respiration inhibition of microbial strains is used as indicator of their exposure to certain stress, whereas elimination of stress normalizes their respiration rate and activities in environment. Respiration inhibition of Escherichia coli caused by Orange 2 and reactive black 5 was lowered by treatment of these dyes with Cunninghamella elegans UCP 542 (Ambrosio and Campos-Takaki 2004). Laccases from Pleurotus ostreatus reduced growth inhibition potential of Remazol Brilliant Blue R to Bacillus cereus, strain 6E/2 by 95% in just 3 days (Palmieri et al. 2005). Fungal treatment of azo dyes also reduces their cytotoxic and genotoxic properties. Vanhulle et al. (2008) reported that treatment of Acid Blue 62 by Pycnoporus sanguineus MUCL 41582 and Perenniporia ochroleuca MUCL 41114 minimized cytotoxic and genotoxic properties of Acid Blue 62. Treated synthetic wastewater had no damage to deoxyribonucleic acid of Salmonella typhimurium strain (TA 104 recN2), whereas untreated synthetic wastewater caused high DNA damage. Thus, similar to bacteria, fungi also simultaneously decolorize and detoxify azo dyes; however fungi are slow degraders of dyes than bacteria.

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14.5.3 Consortia Previous studies also reported accelerated detoxification of dyes containing wastewaters using bacterial, fungal, and mycobacterial consortia. Although pure cultures of bacteria and fungi detoxify the dyes efficiently, sometimes their consortia exhibit better results compared to pure cultures because of synergistic metabolic actions (Tony et al. 2009; Su and Lin 2013). Some of the microbial consortia exhibiting efficient color removal of structurally variable azo dyes are summarized in Table 14.4. Su and Lin (2013) studied fungi–bacteria synergism and suggested enhanced detoxification of Reactive Red 120 (90%) by consortium comprising of Aspergillus niger and Bacillus sp. within 24 h. Additionally, complete detoxification of Reactive Red 120 was recorded by mycobacterial consortium that did not inhibit seed germination of mung bean seeds. Similarly, consortium of Proteus vulgaris NCIM-2027 and Micrococcus glutamicus NCIM-2168 exhibited efficient decolorization of Green HE4BD, and non-toxic metabolites were produced (Saratale et al. 2010a). The bacterial consortium comprising of Bacillus subtilis, Bacillus cereus, Bacillus mycoides, Bacillus sp., Micrococcus sp., and Pseudomonas sp. decreased phytotoxic effects Red dye (200 milligram per liter) to Sorghum vulgare and Zea mays in 24 h (Saratale et al. 2010b). Another consortium consisting of Providencia sp. SDS and Pseudomonas aeruginosa strain BCH was explored to decrease cytotoxicity and genotoxicity of Red HE3B. This consortium was found efficient and lowered cytotoxicity and genotoxicity to Allium cepa root cells within 1 h (Phugare et al. 2011a). Various other researchers have also demonstrated treatment of synthetic and actual textile wastewaters by microbial consortia (Saratale et al. 2009; Phugare et al. 2011b; Lade et al. 2015c; Shah et al. 2016). The effectiveness of microbial consortium in decreasing zoo-toxic and microbicidal/microbiostatic effects of azo dyes has also been widely studied (Table 14.4). Lade et al. (2015d) noticed 49% mortality of Daphnia magna larvae by untreated RB5 solution, whereas no mortality was observed after treatment with consortia consisting of Providencia rettgeri strain HSL1 and Pseudomonas sp. SUK1 together. Similarly, 60% death of larvae of Daphnia magna was recorded by trypan blue dye (50 mg L
1), but no mortality was noticed after treatment with microbial consortium consisting of 15 strains (Lade et al. 2015c). The consortium containing strain NCIM-2027 and strain NCIM-2168 was also found effective in decreasing noxious effects of azo dyes to soil microbes (Saratale et al. 2010a). The most of identified microbial strains are effective in minimizing phytotoxic, zoo-toxic, and genotoxic impacts of dyes in water. However, there are a few reports indicating that microbes may cause bio-activation of dye molecule. So, the use of microbial consortia with ability to mineralize azo dyes is preferred than pure cultures of bacteria and fungi in order to achieve mineralization.

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Table 14.4 Biodegradation of different azo dyes using microbial consortia Sr. # 1

Dyes under studya RNB

5

Microbial consortium Bacillus pumilus, Zobellella taiwanensis, Enterococcus durans Acinetobacter baumannii MN3 and Pseudomonas stutzeri MN1 Bacillus sp., Staphylococcus sp., Escherichia sp., Enterococcus sp., Pseudomonas sp. DMB5 Pseudomonas fluorescens, Enterobacter aerogenes, Shewanella sp., Arthrobacter nicotianae, Bacillus beijingensis, Pseudomonas aeruginosa Microbial Consortium (15 bacteria)

6

Aspergillus niger and Bacillus sp.

RR20

7

Providencia sp. and Pseudomonas aeruginosa strain BCH Proteus vulgaris, Micrococcus glutamicus Enterobacter sp., Phlebia brevispora

Red HE3B

Proteus spp., Pseudomonas spp., and Acinetobacter spp. Galactomyces geotrichum and Brevibacillus laterosporus Pseudomonas aeruginosa, Bacillus flexus, and Staphylococcus lentus Pseudarthrobacter sp. and Gordonia sp.

MR, CF

Enterococcus faecalis and Klebsiella variicola Bacillus pumilus HKG212, Zobellella taiwanensis AT 1–3, and Enterococcus durans GM13 Brevibacillus laterosporus and Galactomyces geotrichum Pseudomonas sp., Lysinibacillus sp., Lactococcus sp., and Dysgonomonas sp. Bacillus sp. strain AK1, Kerstersia sp. strain VKY1, and Lysinibacillus sp. strain AK2 Dichotomomyces cejpii MRCH 1–2 and Phoma tropica MRCH 1–3 Neisseria sp., Vibrio sp., Bacillus sp. and Aeromonas sp.

RR 198

2 3

4

8 9

10 11 12 13 14 15

16 17 18

19 20

CR RBV5R

References Das and Mishra (2019) Kuppusamy et al. (2017) Shah et al. (2016)

RB220

Patel and Bhatt (2015)

TB

Lade et al. (2015c) Su and Lin (2013) Phugare et al. (2011a) Saratale et al. (2010a) Harry-asobara and Kamei (2019) Joshi et al. (2015) Waghmode et al. (2019) Shanmugam et al. (2019) Eskandari et al. (2019) Eslami et al. (2019) Das and Mishra (2019)

Green HE4BD CR, CV

MR AB113 RB5

RNB

RR MY, DFBG, ABSGR, AO2, DB5B, AB P4R

CR, MR, RB NOFNR, NBBFNR, NSBG, BYS8G, BRS2B

Kurade et al. (2019) Guo et al. (2019) Masarbo et al. (2019) Krishnamoorthy et al. (2018) Karim et al. (2018) (continued)

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Table 14.4 (continued) Sr. # 21

22

23

Microbial consortium Bacillus subtilis, Bacillus cereus, Bacillus mycoides, Bacillus sp., Micrococcus sp., and Pseudomonas sp. Acinetobacter lwoffii, Acinetobacter albensis, and Enterobacter cloacae

Dyes under studya Red, green, black, and yellow dyes and mixture of all these dyes RR195, RB23, RB1

Enterococcus sp., VSS-2 as Bacillus sp., VSS-4 as Stenotrophomonas ssp., and VSS-6 as Pseudomonas sp.

ROM2R

References Mahmood et al. (2015) Velvizhi and Krishnaswamy (2018) Jagwani et al. (2018)

a

RNB (Remazol Navy Blue), CR (Congo Red), RBV5R (Remazol Brilliant Violet 5R), RB220 (Reactive Blue 220), TB (Trypan Blue), RR20 (Reactive Red 120), Red HE3B, Green HE4BD, CV (Crystal Violet), MR (Methyl Red), CF (Carbol Fuchsin dye), AB113 (Acid Blue 113), RB5 (Reactive Black 5), RR 198 (Reactive Red 198), RR (Remazol Red), MY (Metanil Yellow), DFBG (Direct Fast Black G), ABSGR (Acid Brilliant Scarlet GR), AO2 (Acid Orange 2), DB5B (Direct Blue 5B), AB (Acid Black), P4R (Ponceau 4R dye), RB (Reactive Blue), NOFNR (Novacron Orange FNR), NBBFNR (Novacron Brilliant Blue FN-R), NSBG (Novacron Super Black G), BYS8G (Bezema Yellow S8-G), BRS2B (Bezema Red S2-B), RR195 (Reactive Red 195), RB23 (Reactive Brown 23), RB1 (Reactive Black 1), ROM2R (Reactive Orange M2R)

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Chapter 15

Natural Remediation Techniques for Water Quality Protection and Restoration George Pavlidis

and Helen Karasali

Contents 15.1

Introduction and Background: Natural Remediation Systems; from Nature to Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.2 Mechanisms Underlying Natural Treatment Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.3 Applications, Advantages, and Performance of Natural Treatment Systems . . . . . . . . . . . 15.4 Potential Limitations of Natural Treatment Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.5 Plant Types Used for Each Pollutant . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Environmental deterioration caused by anthropogenic sources is steadily increasing with the years passing due to the increase in population, industrialization, and urbanization. This has driven scientists and politicians to research and develop on techniques necessary for environmental protection. Constructed wetlands are engineered systems that have been designed to employ natural processes including vegetation, soil, and microbial activity in order to treat contaminated soil and water. In addition and in parallel to constructed wetlands, phytoremediation systems can also play a significant role in the reduction of organic pollutants and metals. As such, today, up to 100% of certain contaminants may be removed from environment using natural remediation techniques. In the present chapter, the major categories of these systems, their advantages and limitations, as well as natural systems pollution reduction efficiency are presented. Keywords Natural · Treatment · Systems · Phytoremediation · Constructed wetland · Wastewater · Metals · Organic pollutants · Pesticides · Reduction · Advantages · Disadvantages

G. Pavlidis (*) · H. Karasali Laboratory of Chemical Control of Pesticides, Department of Pesticides Control and Phytopharmacy, Benaki Phytopathological Institute, Athens, Greece e-mail: [email protected]; [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 327 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_15

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Introduction and Background: Natural Remediation Systems; from Nature to Application

Nature has mechanisms to recover by itself. This simplistically describes the process of natural attenuation. Natural attenuation consists a non-invasive natural remediation process and includes a mixture of physicochemical and/or biological processes, which under certain conditions, have the potential to eliminate the mass, toxicity, mobility, volume, or concentration of an environmental pollutant without human intervention (EPA 1997). Attenuation processes may be distinguished as “nondestructive,” when the concentration of a contaminant but not the total mass in the system is reduced (e.g., in dispersion and dilution processes), or “destructive” when the procedure results in the decrease of the total pollutant mass in the system (e.g. chemical and biological degradation). The processes involved in natural attenuation and also utilized in natural remediation systems include (Adriano et al. 2004; Warężak et al. 2016; Lorion 2001; Dunbabin and Bowmer 1992; Stottmeister et al. 2003): • • • • • • • • • • •

Biodegradation Dispersion Dilution Adsorption/absorption Volatilization Evapotranspiration Precipitation Sedimentation Ion exchange Chemical or biological stabilization Biodegradation, transformation, or destruction of pollutants

Natural attenuation is a well-established remedial strategy for several inorganic compounds and organic chemicals, primarily benzene-, toluene-, ethylene-, and xylene-containing compounds as well as petroleum compounds; however, as a process, it cannot destroy metals (Adriano et al. 2004). Taking advantage of these pollution reduction capabilities, local communities have utilized natural wastewater treatment systems for several centuries, whereas scientists have targeted their research during the last decades on the functioning and optimization of natural remediation systems (Angelakis and Tchobanoglous 1995; Knight et al. 1999). In general, the natural attenuation systems may be classified to soil-based, i.e., waste applied on soil surface, and to water-based systems and include among others (Angelakis and Tchobanoglous 1995; Crites et al. 2014): • Soil based – Waste stabilization ponds/infiltration systems – Vegetative buffer strips, including agroforestry/alley cropping systems

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• Water based – Natural wetlands – Constructed wetlands: vertical flow or horizontal flow Free water surface constructed wetlands Vegetated submerged bed constructed wetlands Subsurface flow systems Vertical flow systems with upstream or downstream characteristics and continuous or intermittent loading Floating plant wetlands Hybrid systems (i.e., combinations of abovementioned systems) – Aquaculture systems Natural treatment systems have been extensively studied during the last decades, and their applications are increasingly expanding throughout the world (Vishnoi and Srivastava 2007; Lorion 2001). The major pollutants that are known to be reduced in natural treatment systems are wastewater load, i.e., biological oxygen demand, chemical oxygen demand, microorganisms, total suspended solids, nitrogen, phosphorus, etc., and can be extended even to inorganic traces, metals, and organic pollutants including pesticides and radioactives as well (Sudarsan et al. 2012, Bhatia and Goyal 2014; Lorion 2001; Dushenkov 2003). In the present chapter, the main systems and the underlying mechanisms for pollution remediation via natural treatment systems will be presented along with their applicability and benefits as well as some possible limitations.

15.2

Mechanisms Underlying Natural Treatment Systems

Accelerating the natural attenuation processes with human interference drives technically to an assisted natural remediation and as such the idea of natural treatment systems. The purpose of natural treatment systems is the re-establishment of disturbed ecosystems and their sustainability for benefits to human and nature (Mahmood et al. 2013). As mentioned by Hemond and Benoit (1988), a wetland is not a simple filter; it embodies chemical, physical, and biotic processes that can detain, transform, release, or produce a wide variety of substances. Natural treatment systems may include one or more processes as previously described, and their function is based on the ecological principals where aquatic plants, algae, and microbes absorb pollutants found in the wastewater (Mahmood et al. 2013); however, they have the advantage and flexibility of being constructed. The vast majority of systems (i.e., apart from stabilization ponds) includes plant material that contributes to the uptake or transformation of pollutants resulting in partial or substancial remediation (Vishnoi and Srivastava 2007).

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Apart from pollutants uptake, plants in a wetland provide a substrate (roots, stems, and leaves) upon which microorganisms can grow as they break down organic materials and uptake heavy metals (Zhang et al. 2010; Giripunje et al. 2015). This process is called phytoremediation and occurs through a series of complex interactions between plants, microbes, and the soil, including accumulation, hyperaccumulation, exclusion, volatilization, and degradation, while stabilization of contaminated sediments can be also performed by forming dense roots under the surface (Vishnoi and Srivastava 2007; Stottmeister et al. 2003). As explained by Cunningham et al. (1997), phytoremediation as a term derives from the combination of the Greek prefix “phyto” that means “plant” and the Latin suffix “remedium” meaning to cure or restore. The different types of phytoremediation can be distinct and are summarized below (Erakhrumen 2007; Dushenkov 2003; Khan et al. 2004; Shukla et al. 2010; Vishnoi and Srivastava 2007; Bhatia and Goyal 2014): • Phytoextraction (or phytoaccumulation): The use of pollutant accumulating plants or algae to remove pollutants from soil by concentrating them in harvestable plant parts. Mostly used for heavy metals and organics. Living plants can absorb contaminants until their harvest, and at that time a significantly lower concentration level will remain in soil; hence a new phytoextraction cycle may be necessary. • Phytotransformation: The in-plant transformation or metabolism of complex organic molecules (such as pesticides, solvents, explosives, and other chemicals) to more simple and smaller compounds. • Phytostimulation (or rhizodegradation): Plant-assisted bioremediation that enhances microbial and fungal degradation in soil by release of exudates/enzymes into the plant root zone (rhizosphere). Mostly applied for organic substances (including petroleum products and solvents). • Phytovolatilization: The use of certain plants to uptake pollutants that are water soluble and subsequently evaporate or volatize them into the atmosphere. • Rhizofiltration: Generally similar process to phytoextraction but principally applied for groundwater remediation. This involves the use of plant roots to absorb and adsorb pollutants, mainly metals, but also organic pollutants, from water and aqueous waste streams. The plants used for this process are initially acclimated to the pollutant. After root saturation the filter plants are harvested and disposed safely. • Pump and tree (dendroremediation): The use of trees to transform or extract pollutants from the soil. • Phytostabilization: The use of plants to reduce the mobility and bioavailability of pollutants, preventing their migration to water recipients and the food chain. Contaminants are absorbed and accumulated by roots, adsorbed onto the roots, or precipitated in the rhizosphere. • Hydraulic control: The control of the water table and the soil field capacity by plant canopies. • Riparian corridors and buffer strips: Phytoremediation via plant or tree root uptake along river banks or shallow groundwater. Several other natural processes

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described above are involved to restrict the pollutant transport and remediate contaminated sites. Buffers may be used at the edge of cultivated fields to control agrochemicals transport to water, whereas they have also been used along the perimeter of landfill sites. • Vegetative cover: The use of plants as a cover or cap growing over landfill sites. Apart from their aesthetical benefits, they can play a significant role in the erosion control and reduction of pollutants content. During the last four decades, specific attention was given to phytoremediation plant species selection as they can significantly contribute to the contaminants reduction efficiency, always in correlation with other system parameters (e.g., hydrology, hydraulic load, soil characteristics, vegetation, pH, wastewater retention time, temperature, air moisture, microbial diversity, and redox potential) (Eke and Scholz 2008).

15.3

Applications, Advantages, and Performance of Natural Treatment Systems

Natural treatment systems and especially the constructed wetlands that constitute the main natural systems category, have been used to treat a variety of waste, including (Sudarsan et al. 2012; Zhang et al. 2010; Haberl et al. 2003; Knight et al. 1999; Lorion 2001): • • • • • • • • • • • • • • • •

Municipal wastewater Secondary effluent Storm water Hydrocarbon effluents from petroleum industry Industrial wastewater Pulp and paper wastewater Food processing waste Slaughterhouse waste Dairy wastewater Olive mill waste Agricultural runoffs with pesticides and nutrients residues Liquid waste from farms with animal production Landfill leachates Airport and highways runoff Acid mine drainages Shallow soil and groundwater

The constructed wetlands, which in fact constitute a transitional terrestrial-aquatic system, can be classified into two categories: the free water surface and the subsurface flow systems (Crites et al. 2014; Mbuligwe et al. 2011; Lorion 2001; Stefanakis et al. 2014). In the free water surface systems, plants are rooted in sediment, and

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Fig. 15.1 Types of constructed wetland systems. (Modified after Mahmood et al. 2013). Several combinations of constructed wetland system have been found throughout the world, with varying efficiencies on pollutants reduction. The most common are subsurface flow systems as they exhibit high efficiency in line with low distortion and reduced installation area necessity

there is surface flow of wastewater, whereas, in the subsurface flow systems, plants are rooted in a porous media (e.g., sand, gravels, or aggregates) and wastewater flows through this media. Wetland plants are active components for the phytoremediation in constructed wetlands and as such must be capable of high nutrient uptake, able to grow in contaminated water, and be easily regulated within the wetland (Angassa et al. 2019). Subsurface flow systems may be further classified to horizontal flow subsurface flow systems and vertical flow subsurface flow systems (Mahmood et al. 2013), as also presented in Fig. 15.1. The systems used for domestic wastewater treatment are subsurface flow constructed wetlands only, whereas, for tertiary treatment, both surface and subsurface flow constructed wetlands may be used (Haberl et al. 2003). Constructed wetlands are considered a best available technique, where applicable, for several reasons. The major are presented hereunder (Haberl et al. 2003; Crites et al. 2014; Vishnoi and Srivastava 2007; Lorion 2001; Haarstad et al. 2012): • • • • • • • • • •

They are less expensive to build than conventional options. Their construction and operation is more simple and cost-effective. They have low operation and maintenance costs. They require only periodic operation and maintenance. They have the potential to tolerate flow fluctuations, different constituents, and concentration and exhibit a high process stability (buffering capacity). They facilitate water reuse. They provide habitat for wetland organisms and wildlife. They can be harmoniously fit into the landscape. The aquatic plants have the potential to absorb CO2, thus sequestering carbon from atmosphere. They are an environmentally sensitive “less-intrusive” approach that is viewed with favour by the general public.

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• They have the potential for self-standing use or in conjunction with other cleanup technologies. • They have the ability to reduce pollutant content in situ to remediate shallow soil and groundwater and can be also used in surface water bodies. • The waste-byproducts production (i.e. sludge) is very low. • There are advantages occurring from plants’ presence, including aesthetic amelioration and socio-economic benefits. • They are able to treat various wastewaters types, including water polluted with organic substances. • They are easily applicable in remote areas with limited energy resources. As regards phytoremediation treatment systems, some additional benefits are anticipated, including lower treatment costs by up to 1000 times (in comparison to conventional in situ and ex situ methods), easy monitoring of the plants, the possibility to recover valuable products (e.g, timber from remediation trees), and the preservation of natural state of the environment considering the natural occurrence of remediation plants (Shukla et al. 2010). Vegetation can also reduce or prevent erosion, runoff, and dust emissions. The performance of treatment wetlands and natural treatment systems in general has been previously reported in several studies during the last decades. The removal efficiency of treatment wetlands exhibits minimum percentages of 40%, 72%, 53%, 57%, and 70%, for 5-day biological oxygen demand, total suspended solids, total nitrogen, total phosphorus, and chemical oxygen demand, respectively (Knight et al. 1999; Vymazal and Kröpfelová 2009; Mbuligwe et al. 2011). In a respective full-scale constructed wetland study performed at Nea Madytos, Greece, the mean removals for the monitoring period were 90.8% for biological oxygen demand, 89.0% for chemical oxygen demand, 83.9% for total Kjeldahl nitrogen, 83.8% for ammonia, 38.8% for total phosphorus, 17.4% for ortho-phosphate, 90.4% for suspended solids, and 99.9% for coliforms (Gikas et al. 2011). Pertinent reductions were also observed by Eke and Scholz (2008), in particular 70–80% for chemical oxygen demand, 83–90% for ammonia-nitrogen, 88–94% for nitratenitrogen, and 58%–66% for ortho-phosphate-phosphorus. High average removals were observed for chemical oxygen demand, total nitrogen, and total phosphorus by Angassa et al. (2019) in a recent 3-year study at Ethiopia, with the maximum removals attained to be 94.1%, 97.3%, and 89.9%, respectively. Relevant reductions were also observed in agricultural, industrial, and livestock waste (Knight et al. 1999; Vymazal and Kröpfelová 2009), whereas removals even up to 100% have been reported for domestic wastewater as presented in an older review study (Hemond and Benoit 1988; Warężak et al. 2016). At the same rationale, total suspended solids, orthophosphate, and nitrate reductions reaching 78%, 75%, and 84%, respectively, were reported by Schulz and Peall (2001). The latter also observed up to 100% reduction of organophosphate pesticides. Finally, pesticides, pharmaceuticals, personal protection and care products and other organic substances removal has also been reported by Haarstad et al. (2012), that stated up

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to 99% disappearance of these organic contaminants in wetland systems – reductions equivalent or better to those achieved in conventional treatment systems. Constructed wetland systems have also the potential to remove sulfates at efficiencies of 55–77% (Mbuligwe et al. 2011). Lower removals were though observed for landfill leachates, a waste type that has extremely high influent chemical oxygen demand and is difficult to manage even with conventional chemical methods (Vymazal and Kröpfelová 2009). In another natural wetland system with Typha plant species, designed to treat dairy waste, the treatment resulted in significant reduction of biological oxygen demand (73%), total suspended solids (75–83%), and phosphorus (76%), whereas slightly lower was the removal in chemical oxygen demand and ammonium nitrogen (about 26% for both pollutants), mainly due to the very high pollutant content (Sudarsan et al. 2012). In general reduction of nitrogen varied from 10% to 70%, whereas the phosphorus removal varied from 11% to 94% as reported by Warężak et al. (2016). Data from petroleum industry treatment wetlands indicate that natural wetlands are equally or even more effective at removing pollutants in comparison to conventional systems, with the reductions varying between 55–98%, 38–86%, 54–94%, and 10–94%, for 5-day biological oxygen demand, chemical oxygen demand, oils, and phenols, respectively (Knight et al. 1999). The aquatic plant duckweed (Lemna minor) has also been used for the treatment of domestic wastewater in a free water surface pilot system in Israel where high removals of coliforms (approximately 95%), chemical oxygen demand, and 5-day biological oxygen demand (67.5% and 70.6%, respectively) were observed (Ran et al. 2004). Up to 99.9% coliform reduction was also reported by Mbuligwe et al. (2011). Another common plant used for phytoremediation is the water hyacinth. Fox et al. (2008) examined the pollution abatement potential of the hyacinths, and it was found to account for 60–85% of the nitrogen removed from solution. The metabolism of other organic contaminants such as polycyclic aromatic hydrocarbons (Warężak et al. 2016), perchloroethylene–trichloroethylene (Pant and Pant 2010), 2,4,6-trinitrotoluene, glyceroltrinitrate, and other chlorinated compounds has also been previously noted (Singh and Jain 2003; Cunningham et al. 1997). Natural treatment systems exhibit potential for benzenes removal, based on Braeckevelt et al. (2011) findings in a constructed wetland using Phragmites australis system observed reductions of 59–65% for monochlorobenzene, 59–69% for 1,4-dichlorobenzene, and 29–42% for 1,2-dichlorobenzene, with volatilization accounting only for 2–4% of the total amount removed. Benzene removals from 85% to 95% were also reported by Eke and Scholz (2008), in the presence of Phragmites australis. Natural treatment systems and particularly constructed wetlands have been used for the removal of metals from industrial-type wastewater as well. The removals may reach up to 100% for certain metals, i.e., zinc, cadmium, copper, and lead (Dunbabin and Bowmer 1992; Haarstad et al. 2012), whereas removal of ferrous, boron, chromium, nickel, uranium, selenium, and arsenic has also been previously reported (Marín and Oron 2007; Khan et al. 2009; Mkandawire et al. 2004; Frankenberger Jr and Arshad 2001; Giripunje et al. 2015).

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Another type of natural phytoremediation treatment systems are the soilbased systems, and principally the vegetative buffer strips (also known as vegetative filter strips), and their subcategories: i.e., alley cropping systems, conservation cover, contour buffer strips, riparian herbaceous cover, riparian forest buffers, filter strips, grassed waterways, vegetative barriers, drift buffers, and agroforestry systems (Strock et al. 2010; Pavlidis and Tsihrintzis 2018). The incorporation of buffers and filter strips into drainage systems can improve water quality as well as the effectiveness of drainage systems and can serve both for overland flow remediation and shallow groundwater. The vegetation in these buffers has a significant phytoremediation potential; however, its performance is related to several factors with the major being: field design parameters, the soil type, local hydrogeology, the weather conditions and irrigation patterns, and of course the amount/type of agrochemicals used. The reductions in agroforestry cropping systems may reach up to 100% for nitrogen, phosphorus, and certain pesticides as presented in Pavlidis and Tsihrintzis (2018) and in a recent experimental study with maize crop and poplar trees as the pollution buffer (Pavlidis et al. 2018). Τhere are a large number of studies on the advantages that they offer, which can be both environmental and socioeconomic and have been previously been reported by Pavlidis and Tsihrintzis (2018). As regards soil compartment, these include: • Reduction of erosion • Reduction of runoff • Reduction of agrochemicals (nutrients and pesticides) and their environmental effects • Raindrop interception • Soil fertility improvement • Nutrient recycling and nitrogen fixation • Increase of use of belowground resources • Weed control • Carbon sequestration • Improved soil biomass • Accelerated mineralization • Remediation of soils and shallow groundwaters Significant benefits are also observed for the water compartment, including among others (Pavlidis and Tsihrintzis 2018): • • • • • • •

Improved water infiltration Improved soil water storage, recharge, and retention Access to lower aquifer water table via deeper tree roots Water quality enhancement Reduction of groundwater leaching Removal of pollutants from unsaturated or low-depth saturated zone Flood control

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Potential Limitations of Natural Treatment Systems

Among the most important parameters during the design of a wastewater treatment system apart from its advantages are the respective limitations. Natural remediation systems exhibit an extended variety of advantages and a wide range of applications and do constitute an environment-friendly wastewater treatment technology. As in every technology, there are some shortcomings also in the case of natural treatment systems. In the case of a “design from scratch” system, it shall be clearly demonstrated whether a conventional system or a natural system better fits for the treatment purpose. The potential limitations of using natural treatment technologies are (EPA 1997; Haarstad et al. 2012): • Extensive and complicated examination is needed in order to achieve specific pollutants reduction. • Require larger treatment areas. • Preliminary studies are deemed necessary as removal rates may differ for each area, waste type, and pollutant. • The prevailing treatment process remains unclear in some points. • Potential for migration of contaminants from water to sediment and/or plants that will need specific management as hazardous waste. • Biodegradation can result in the formation of metabolites, potentially the same or even more persistent, mobile, or toxic for the environment and organisms. • Longer wastewater retention time needed for treatment in comparison with conventional systems. • System efficiency is rather dependent on location-specific conditions (e.g., temperature, hydrology, weather, and biogeochemical conditions) as well as the age of the treatment system. With reference to the phytoremediation only, some additional to the abovementioned disadvantages include (Vishnoi and Srivastava 2007): • Limited to shallow streams, soil, and groundwater only. • Involves mass transfer process (i.e., an amount of the pollutant remains in the plant and need special management). • There is always a potential for plant toxicity from certain hazardous substances. • Climatic conditions may affect the plant growth and as such the remediation rate. • It may be less effective for strongly sorbed (e.g., Polychlorinated biphenyls) and weakly sorbed contaminants.

15.5

Plant Types Used for Each Pollutant

Some plant species that are commonly used in natural treatment and remediation systems have been noted in the previous chapters. In the present section, a categorization of the plants used for phytoremediation and in constructed wetlands with

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plants will be performed. Phragmites spp. are a very commonly used plant species that has been proven efficient for the treatment of domestic wastewater, petrochemicals, mineral oil, aromatic hydrocarbons, glycols, olive mill waste, dairy waste, metals, radionuclides, and organic pollutants in general (Haberl et al. 2003; Eke and Scholz 2008; Giripunje et al. 2015; Angassa et al. 2019). Another common plant used in constructed wetlands is Typha spp., which was found to be efficient for the reduction of pollutant content of domestic wastewater, hydrocarbons, oils and grease, chlorinated compounds (e.g., trichloroethylene), and pesticides (Haberl et al. 2003). Cyperus alternifolius, Scirpus spp., Rumex hydrolapathum, Schoenoplectus spp., Heteranthera dubia, Monochoria spp., Brachiaria spp., Galega orientalis, Lolium perenne, and Villarsia exaltata are also some of the other species that if planted in constructed wetlands, can help to accumulate and absorb, among others, metals; organic matter; benzene-, toluene-, ethylene-, and xylene-containing compounds; aromatic compounds; 2,4,6trinitrotoluene; polycyclic aromatic hydrocarbons, etc. (Giripunje et al. 2015; Haberl et al. 2003; Shukla et al. 2010). Additionally, some submerged plants such as Potamogeton crispus L. and P. pectinatus L. exhibited a potential for lead uptake from sediments, Myriophyllum spicatum was found also to accumulate Hg also from sediments, whereas Cyperus alternifolius and Villarsia exaltata in constructed wetlands effectively removed cadmium, copper, manganese, zinc, and lead (Giripunje et al. 2015). Finally, as regards the soil -phytoremediation systems, among the genera/species that have received attention in agroforestry and alley cropping systems as well as tree remediation systems, these are Populus spp., Juglans nigra, Salix, Eucalyptus, Betula, Alnus, Robinia, and Nothofagus trees, whereas other minor species have also been reported (Prunus avium, Pinus pinea, Fraxinus pennsylvanica, Juniperus virginiana, and Quercus ilex) (Pavlidis and Tsihrintzis 2018). Pesticides, chlorinated phenols, and solvents were also found to be prone to phytoremediation using several grasses, herbs, legumes, as well as the crops themselves (e.g., rice, wheat, sugar beet, barley, etc.) (Shukla et al. 2010). As such, it can be realized that natural remediation systems enhanced with plants exhibit the potential for remediation and protection of soils, shallow groundwater, and surface water.

15.6

Conclusions

Natural treatment and remediation systems can play a significant role in environmental protection and restoration. The different system types can serve for the treatment of the vast majority of aquatic systems’ pollutants, including organic substances, fecal pollutants, metals, nitrogen and phosphorus, pesticides in all forms, as well as polycyclic aromatic hydrocarbons, polychlorinated biphenyls, chlorinated compounds, aromatic compounds, and benzene-, toluene-, ethylene-, and xylene-containing compounds and hydrocarbons. The removal efficiencies for the aforementioned pollutants can reach up to 100%; however, it shall be kept in mind that these systems’ performance is highly dependent on several external

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parameters, and as such, it is apparent that extensive testing is deemed necessary before field application of each system, in order to ensure about its applicability for the geoclimatic conditions of each area and the waste-specific characteristics, as both of them play a significant role in natural systems efficiency.

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Chapter 16

Phytoextraction of Heavy Metals from Complex Industrial Waste Disposal Sites Babatunde Oladipo, Aramide M. Akintunde, Sheriff O. Ajala, Samuel O. Olatunji, Olayomi A. Falowo, and Eriola Betiku

Contents 16.1 16.2 16.3 16.4

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Heavy Metal Sources in the Ecosystem . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Adverse Effects of Heavy Metals on Well-Being of Humans . . . . . . . . . . . . . . . . . . . . . . . . . Cleanup of Heavy Metals Contaminated Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.1 Surface Capping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.2 Encapsulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.3 Electrokinetic Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.4 Chemical Immobilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.5 Soil Flushing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.6 Vitrification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.4.7 Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.5 Phytoextraction: A Green Remediation Method of Heavy Metal Pollution in Complex Industrial Waste Disposal Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6 Plants as Bioaccumulators of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6.1 Metal Excluders . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6.2 Metal Indicators . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.6.3 Metal Accumulator Plant Species or Hyperaccumulators . . . . . . . . . . . . . . . . . . . . 16.7 Mechanism of Phytoextraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.8 Advantages of Phytoextraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.9 Limitations of Phytoextraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.10 Factors Affecting Phytoextraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.11 Conclusions and Future Investigation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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B. Oladipo · O. A. Falowo · E. Betiku (*) Department of Chemical Engineering, Obafemi Awolowo University, Ile-Ife, Osun State, Nigeria e-mail: [email protected] A. M. Akintunde · S. O. Olatunji Department of Environmental Engineering, Texas A&M University-Kingsville, Kingsville, TX, USA S. O. Ajala Frank H. Dotterweich College of Engineering-Sustainable Energy, Texas A&M UniversityKingsville, Kingsville, TX, USA © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 341 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_16

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Abstract Vast industrialization, expanded rate of waste generation, intense agricultural impacts, mining activities, and polluted water systems are adding to heavy metal pollution on soil and water sources. These heavy metal-polluted resources eventually get concentrated in the food system that are utilized by humans and animals and, thus, bring about a serious threat to ecological and human well-being. Removing these contaminants requires great effort and sophisticated operational techniques. Phytoextraction is a modern approach to the removal of heavy metals from contaminated sites. Thus, this chapter extensively details the background, sources, and adverse impacts of some heavy metals; techniques in phytoremediation systems, mechanism of phytoextraction as a remediation approach, and advantages and limitations of phytoextraction were also reviewed. Keywords Heavy metals · Contaminants · Phytoremediation · Phytoextraction · Hyperaccumulators · Bioremediation · Waste disposal · Industrial waste · Pollution · Cleanup

16.1

Introduction

The two pivotal foundations of natural resources on which sustainability for continuous survival lies on are land and water. Over the years, both sources have been extremely contaminated via harmful substances. The pollution problem of heavy metals is getting aggravated with expanding mechanization, industrialization, and disruption of natural, biological, geological, and chemical cycles. The organic part of wastes generated from these industrial activities is biodegradable; however, heavy metals are non-biodegradable; hence, this has led to their accumulation in soils and waters which is a threat to living habitat. Accumulation of these metals in biological organisms is referred to as bioaccumulation and as they move from lesser to greater trophic levels, their concentration increases, a process referred to as biomagnification. Due to the toxic impacts of the contaminants in soil, heavy metals are responsible for the reduction in quantities and functions of soil organisms (Khan et al. 2010). Thus, there is a high possibility of contaminants bioaccumulation and their penetration into the environmental hierarchy via food chain (Anawar et al. 2002; Reza and Singh 2010). Evacuation of the contaminants is a demanding task, which is greatly needed in order to forestall or lower their harmful impacts on humans and ecological disturbance. In a way to sustain the good mark of soils and waters and protect them from pollution, endless attempts have been made to design techniques that are simple to apply, eco-friendly, and less costly. Various processes employed for this purpose are challenged by the effects of high price, demanding labor, modification of soil attributes, and disruption of soil microflora. Interestingly, phytoremediation is a preferred approach to curb the menace. By definition, phytoremediation is the

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utilization of plants and related soil microorganisms to remove or lower the noxious impacts of contaminants in the affected areas. Phytoremediation is a modern approach and is regarded as an economical technology, effective, eco-benign, and solar-steered technique with satisfactory communal approval. In phytoextraction, novel effective metal hyperaccumulating plants are employed for their qualities in remedying and storing of various heavy metals. The present chapter, thus, reviews the different sources and harmful effect of heavy metals, different techniques in phytoremediation as it applies to wastewater pollution and complex waste disposal sites, mechanisms of phytoextraction, benefits and limitations of phytoextraction, and suggestions for future investigation in phytoextraction approach.

16.2

Heavy Metal Sources in the Ecosystem

Heavy metals are broadly dispersed in the environs due to their numerous industrial, domestic, pharmaceutical, agricultural, medical, and technological applications (González-Acevedo et al. 2018; Tchounwou et al. 2012). They enter into different media of the environment (water, soil and air) through various natural and anthropogenic (human activities) sources. Though measurable quantities of heavy metals are discovered beneath the surface of the earth, their most significant industrial sources are from anthropogenic processes such as mining and smelting activities, processing of metal in refineries, burning of coal in power plants, microelectronics, combustion of petrol, wood preservation, plastics, paper processing, and textile plants that have contributed significantly to environmental pollution (Arruti et al. 2010; Asad et al. 2018; He et al. 2005). In addition, natural incidences (e.g., volcanoes and forest fires) have been reported to adversely add to heavy metals environmental pollution (He et al. 2005). Various sources of some heavy metals are presented in Table 16.1.

16.3

Adverse Effects of Heavy Metals on Well-Being of Humans

Heavy metals are characterized as vital and non-vital based on their activities on living systems. Heavy metals essential for important physiological and biochemical activities in plants and animals are known as vital heavy metals, such as Mn, Fe, Ni, Zn, and Cu (Ali et al. 2013), while heavy metals that are not required for any biochemical and physiological activities by plants and animals are referred to as non-vital heavy metals, such as As, Cd, Cr, Pb, and Hg (Ali et al. 2013). Since plants possess the capability to amass metals from soil solution, this capacity permits them to accumulate both the vital and non-vital heavy metals which may have entered into

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Table 16.1 Various sources of some heavy metals Sources Heavy metals As

Natural Weathering and eroding of rock and soil, volcanic eruptions, and the escape of volatile methyl arsines from soil

Cd

Airborne soil particles, volcanoes and weathering of rocks, biogenic materials, sea spray, forest fires

Cr

Weathering of rocks, erosion, volcanic dust, and gases

Hg

Volcanoes, forest fires

Pb

Volcanoes, forest fires, sea-salt

Cu

Wind-blown dust, decaying vegetation, forest fires, sea spray

Zn

Erosion, wind-blown dust, sea-salt spray, forest fires

Anthropogenic (industrial and environmental) Antifungal wood preservatives, mining activities, use of arsenical pesticides, coalfired power generation, from base metal and gold production facilities Iron and steel manufacturing, phosphate fertilizer production, coal combustion, zinc production, waste incineration, disposal of sewage sludge, combustion of fossil fuels, weathering of galvanized metals Metal processing and finishing, refractory bricks manufacturing, mortars, furnace linings, tannery facilities, dyes production, wood treatment, use in corrosion control Electrical industry, medicine, nuclear reactors, caustic soda production, and through dental amalgams and fish consumption Battery manufacturing, consumption of leaded gasoline in motor vehicles, deteriorating paints made of Pb, sewage sludge, coal residues, municipal refuse incineration, wastewaters, phosphate fertilizers Textiles manufacturing, fertilizers, drinking water pipes, wood preservatives, cooking utensils, pesticides, and fungicides Electroplating, toxic waste sites, smelting and ore processing, wastewater of industrial plants, mine drainage, combustion of solid wastes, industrial and domestic sewage and fossil fuels, tire debris, road

References Chung et al. (2014), Tchounwou et al. (2012), Bhattacharya et al. (2002) and CEPA (2007)

Chen et al. (2016), Tchounwou et al. (2012) and CEPA (2007)

Pellerin and Booker (2000), Tchounwou et al. (2012) and CEPA (2007)

Ha et al. (2017), Jaishankar et al. (2014) and Dopp et al. (2004)

Pan et al. (2019), Jaishankar et al. (2014), Lanphear et al. (1998) and CEPA (2007)

Khan et al. (2007) and CEPA (2007)

Wuana and Okieimen (2011) and CEPA (2007)

(continued)

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Table 16.1 (continued) Sources Heavy metals

Ni

Natural

Anthropogenic (industrial and environmental)

Vegetation particulate exudates, forest fires, sea-salt, soil dust, volcanoes, weathering, and erosion of geological materials

surface run-off, corrosion of galvanized and zinc alloy surfaces, mining, coal steel processing, and intensive animal production Metal-plating industries, cement manufacturing, fossil fuel combustion (predominantly coal and oil), nickel mining

References

Khodadoust et al. (2004) and CEPA (2007)

Table 16.2 Permissible limits for metals Heavy metal As Cd Cr Hg Pb Cu Zn

Concentration ranges Groundwater Soils/sediment (μg/L) (μg/kg) 0.3–32,100 100–102,000 0.005–7600 100–345,000 0.42–9010 5.1–3,950,000 0.08–216,900 0.1–1,800,000 0.56–120,000 1000–6,900,000 1–3300 30–550,000 1–697,000 150–5,000,000

Contaminant limit in drinking water (μg/L) 5a 5b; 10a; 5c 100b; 50a; 100c 2b; 2a; 2c 0b; 50a; 5c 1300b; 1300c 5000d

a

Existing maximum contaminant level Proposed maximum contaminant level goals c Proposed maximum contaminant level d Non-enforceable secondary-level standard based on taste, odor, or appearance guidelines b

the soil through the natural and anthropogenic sources. Heavy metals may be toxic and non-toxic to the exposed individual depending on various factors like the amount of dose, exposure route, species of chemicals, age, genetics, and immunity (Asad et al. 2018). But heavy metal concentrations beyond threshold specific to each element have proven to have detrimental effect on plants, animals, and human wellbeing by affecting the normal functioning of biological systems. The US Environmental Protection Agency (EPA) specified the concentration ranges of some heavy metals in groundwater and soils (or sediment) and the maximum contaminant level of these heavy metals in drinking water (Table 16.2). Both acute and chronic exposures to heavy metals by inhalation, ingestion, and skin contact may result in unwanted effects and life-threatening problems even at low amounts. Therefore, heavy metal pollution of the environment from wastewater and complex industrial waste disposal sites should be given extra attention as it can

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Table 16.3 Exposure routes and harmful impacts of heavy metals on human health Heavy metal As

Exposure route Ingestion, inhalation, skin contact, and parenteral route

Cd

Inhalation and ingestion

Cr

Inhalation, skin contact, and ingestion

Hg

Ingestion

Pb

Ingestion and inhalation

Cu

Ingestion

Zn Ni

Ingestion Inhalation and ingestion

Harmful impacts Cardiovascular and peripheral vascular disease, hearing loss, diabetes, neurologic and neurobehavioral disorders, diabetes, portal fibrosis, developmental anomalies, hematologic disorders, and carcinoma Convulsions, abdominal pain, gastrointestinal tract erosion, shock, salivation, loss of consciousness, nausea, burning sensation, vomiting, vertigo, muscle cramps, pulmonary, hepatic or renal injury, and coma Respiratory cancers, nose irritation and nose ulcers, severe gastrointestinal, cardiovascular, hematological, hepatic, renal and neurological effects, and death Gastrointestinal toxicity, neurotoxicity and nephrotoxicity Brain and kidney damage, gastrointestinal diseases, blood poisoning which lowers intelligence quotient, impaired neurobehavioral development, reduced hearing sharpness, language and speech handicaps, poor attention span, and growth subnormality Cirrhosis of the liver, brain, and kidney damage, chronic anemia, intestinal and stomach irritation Dizziness and fatigue Allergic dermatitis (nickel itch), lung cancer, sinuses, reproductive toxic, stomach and throat cancer, genotoxic, hematotoxic, immunotoxic, neurotoxic, pulmonary toxic, loss of hair, nephrotoxic, and hepatotoxic

References Cheng et al. (2018), SaintJacques et al. (2018), Welch et al. (2018) and Argos et al. (2012)

Dziubanek et al. (2017), Satarug et al. (2017) and Tchounwou et al. (2012)

Tseng et al. (2019), Jobby et al. (2018), Yoshinaga et al. (2018) and Tchounwou et al. (2012)

Henriques et al. (2019) and Sharma et al. (2019) Mansel et al. (2019) and Boskabady et al. (2018)

Wuana and Okieimen (2011) and Salem et al. (2000) Hess and Schmid (2002) Mishra et al. (2010), Das et al. (2008), Duda-Chodak and Blaszczyk (2008), Khan et al. (2007) and Salem et al. (2000)

contaminate the food chain. Table 16.3 presents possible exposure routes and harmful impacts of some heavy metals on human well-being depending on their concentrations.

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16.4

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Cleanup of Heavy Metals Contaminated Soil

There are several techniques for cleaning up or treating heavy metal contaminated sites. This section reviews some of these techniques to provide information for both the industries and academic communities on the advancement of innovative heavy metals removal technologies and the adoption of such technologies, as many of these may be site specific. According to US Environmental Protection Agency (EPA), technologies for removing heavy metals from polluted soils can be categorized into source (in situ and ex situ) and containment controls. Liu et al. (2018) broadly classified remediation technologies of heavy metals contaminated soil into two, namely, in situ and ex situ methods. These were further sub-classified into physical, chemical, electrical, and biological as in situ techniques and physical, chemical, and thermal as ex situ methods. Figure 16.1 shows the various major and sub-classification techniques of remediating soil contaminated with heavy metals. Generally, these techniques employ different heavy metal removing or immobilizing mechanisms and have certain merits and limitations over one another (Liu et al. 2018). Wuana and Okieimen (2011) suggested that a combination of these remediation techniques may be employed for economic reasons. Choosing any of the available remediation technologies depends on the following factors: cost, general acceptability, resistance to high metal concentration, capability to deal with complex waste mixtures (heavy metals and organic pollutants), effectiveness in terms of permanent removal of wastes, toxicity, and volume reduction (Wuana and Okieimen 2011).

Fig. 16.1 Remediation techniques for heavy metal-polluted soil. (Modified after Liu et al. 2018)

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In situ remediation technologies do not involve contaminated soil excavation and subsequent transportation to off-site treatment facilities (Liu et al. 2018; Wuana and Okieimen 2011). The merits of these approaches include reduced treatment cost, high public acceptability due to limited exposure of workers and the surrounding populace, and reduced soil disturbance, and a wide range of contaminants can be treated (Martin and Ruby 2004). Furthermore, isolation, removal, and stabilization are the three main remediation tactics in which the in situ remediation technologies are based. Isolation technologies are those involved in the restriction of the pollutant availability by limiting the surface area of exposure and reduction of soil permeability and solubility. The removal technologies employ metal removal from contaminated soil through physical or chemical means, while in the stabilization technologies, chemical amendments and/or plants are employed for reduction of the leachability and bioavailability of the heavy metals in the soil (Martin and Ruby 2004). In situ remediation techniques used for contaminated sites cleanup are bioremediation, surface capping, encapsulation, chemical immobilization, electrokinetic extraction, soil flushing, and phytoremediation. On the other hand, ex situ technology of remediating contaminated soils requires excavation of soil from the contaminated site and its subsequent transportation to off-site treatment facilities which are then disposed after treatment under required environmental permit (Liu et al. 2018). In ex situ remediation technology, additional costs are incurred for soil excavation, transportation, site refilling, and disposal of the treated waste. Meanwhile, a better control of the treatment and time of treatment can be reduced to achieve desirable results within a short time (Liu et al. 2018; Wuana and Okieimen 2011). Various ex situ remediation practices, such as landfilling, soil washing, solidification, and vitrification, have all been employed for heavy metal removal.

16.4.1 Surface Capping This is a containment-based method of soil remediation which involves covering of heavy metal-polluted site with an impermeable material. In this case, efforts are not made to eliminate or remove the metal contaminants; rather the method minimizes the risk of human exposure either via direct skin contact or ingestion. It also prevents further contamination of surface and groundwater sources (Liu et al. 2018). Although the method is recommended for heavily contaminated soil, its major challenges include loss of natural habitat functions, which might not support plant growth. The construction of capping surface also becomes difficult when the surface area of the contaminated soil is large, thus its limitation to contaminated small piece of land (Liu et al. 2018). The cost of capping in USA ranges between $20 and $90 per m2, and this is dependent on the engineering design, material (low or high permeable material), labor, and maintenance.

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16.4.2 Encapsulation Encapsulation involves physical isolation and containment of the contaminated soil (Khan et al. 2004). This approach can also be described as the barrier wall, cutoff wall, or liner, where the contaminated site is isolated and contaminants are bounded with a view to eliminate off-site spread of the pollutants and on-site bio-exposure to the pollutants (Liu et al. 2018). Different materials, such as synthetic textile sheets and concrete, have been employed in the construction of barrier wall to reduce the contaminants infiltration to adjacent groundwater. The challenges of this method include construction of underground impermeable wall at the polluted site and decrease in encapsulation efficiency with time, and it is limited to shallow (restricted to 30 ft) contaminated soil (Khan et al. 2004; Liu et al. 2018). The cost of soil encapsulation relies on the depth of contamination, geology of the site, and the material of construction (Khalid et al. 2017; Khan et al. 2004; Liu et al. 2018).

16.4.3 Electrokinetic Extraction This method is considered a new approach and less prohibitive in situ method for heavy metal removal from contaminated soil. It is based on the principle of creating a suitable and intense electric field gradient between two electrolytic spaces created within the constructed wells at a contaminated site of interest (Gong et al. 2018; Khalid et al. 2017). The mechanisms of metal contaminants removal from soil are electrophoresis (charged particle movement), electromigration (movement of charged chemicals) or electric seepage, electrolysis (electron movement), and electro-osmosis (movement of fluid) (Yao et al. 2012). Kim et al. (2012) conducted an in situ electrokinetic remediation of Pb, As, and Cu contaminated soil and reported that about 40, 17, and 19% of the metals were, respectively, removed within 4 days of treatment. Also, Hg-polluted site was remediated by Rosestolato et al. (2015) using electrokinetic technique, and 60% removal of the Hg contaminant was observed. Li et al. (1998) employed the cation selective membrane technology to address the pH increase which limits remediation efficiency in electrokinetic approach. The authors reported more than 90% removal efficiency with the stoppage of hydroxyl ion to the anode. The merits of this method include little or no by-products, ease of installation and operation, non-disruption of natural soil structure, and short time of treatment. Meanwhile, its demerits include applicability to site with low groundwater flow, limited to the removal of leachable metals, and low effectiveness in heterogeneous soil, and energy cost may be high under the constrict of removing all or reasonable part of the contaminants (Gong et al. 2018).

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16.4.4 Chemical Immobilization Chemical immobilization also referred to as in situ solidification or stabilization is a process that involves trapping or immobilizing the metal contaminants in the soil by formation of precipitates or strongly sorbed moiety (Liu et al. 2018). This process neither removes nor extracts the contaminants but makes them immobile, thereby reducing their transportation within the soil matrix (Tajudin et al. 2016). Due to economic advantages and simplicity of application, solidification and stabilization are commonly used to fix heavy metal contaminants in solid wastes (Gong et al. 2018). In solidification, a binding agent is added to the contaminated soil, and in the process, physical stability is imparted to the material where the contaminants are contained in a solid matrix and their leachability is greatly restricted (Wuana and Okieimen 2011). Also, in stabilization, precipitating agent or stabilizing chemicals are fused into the heavy metal-contaminated soil to bring about physiochemical interaction between the stabilizing agents and the metal contaminants (Tajudin et al. 2016). In this operation, the contaminants are immobilized and the soil is not solidified (Liu et al. 2018). Organic and inorganic binders or a combination of both can be employed in solidification or stabilization process. Cements, clay, zeolites, and phosphate are widely used soil amendments (Sun et al. 2016). Meanwhile, research into the potential application of industrial wastes such as red mud, termitaria, and industrial eggshells have been recently reported for immobilizing heavy metals in contaminated soils (Khalid et al. 2017). Soares et al. (2015) studied the immobilization of Zn and Pb present in contaminated acidic soil due to mining activities using a compost derived from industrial eggshell. The authors reported the soil mobile fraction for Pb and Zn to be greater than 95%.

16.4.5 Soil Flushing Soil flushing is an in situ technique where contaminants are removed by injecting extraction solvent such as water, chelating, and acidic solutions into the polluted soil matrix (Liu et al. 2018). This technique is applicable to all forms of soil contaminants, and it may be applied in combination with other techniques such as biodegradation, pump-and-treat, and activated carbon techniques (Khan et al. 2004). The extraction liquid containing the extracted contaminants with the groundwater is recovered and adequately treated to get rid of the contaminants to meet the regulatory requirement before discharge or recycle (Khan et al. 2004; Martin and Ruby 2004). Soil permeability and groundwater flow characteristics are the major factors affecting this technology. Another challenge of this technique is selection of appropriate extraction fluid in cases where there are more than one contaminant, thus making soil flushing most useful for sites with single metal contaminant (Martin and Ruby 2004).

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16.4.6 Vitrification This technique can be employed either in situ or ex situ depending on the cleanup condition but broadly classified as an ex situ cleanup approach. Vitrification involves application of high temperature (greater than 1500  C) process to restrain the metal contaminants present in the soil which often resulted in the formation of vitreous material in form of solid oxide (Khalid et al. 2017; Wuana and Okieimen 2011). The formed vitrificated material is strong, durable, resistant to leaching, and above all, chemically inert (Liu et al. 2018). During this process, some contaminants such as Hg are volatilized which need to be captured for proper disposal due to environmental regulation of such material (Gong et al. 2018). The source of heating in this case may be through thermal heating (combustion of fossil fuel and natural gas) or electrical heating (electric arc, Joule effect, and plasma process) (Yao et al. 2012). Vitrification by Joule heating with temperature up to 1850  C was used by Dellisanti et al. (2009) to remediate heavy metals (Pb, Zr, and Zn) from ceramic waste, and in the process, these metals were immobilized within the vitrified monolith. Since temperature is paramount to metal immobilization and process cost, thus the conventional fossil fuel heating or electrical heating are sometimes costly, and search into the application of solar energy as heat source has been reported (Gong et al. 2018). Navarro et al. (2013) employed solar technology in the vitrification of mining-contaminated soil and reported immobilization of metals at two different temperatures of 1350 and 1050  C for Fe, Mn, Ni, Cu, and Zn, among others. Although this method remediates contaminated soil permanently and reduced waste volume coupled with possible reuse of produced materials, nonetheless, it has its limitations such as uneconomical for large contaminated site, and electrical conductivity of the site may reduce efficiency of the vitrification process (Gong et al. 2018).

16.4.7 Phytoremediation This is a bioremediation method which requires the utilization of specialty-growing plants to hold, remove, or immobilize contaminants that are found in the soil, thereby help in reducing the environmental impact, and revegetate such contaminated site. Phytoremediation is gaining worldwide attention not only for its efficient application for remediating heavy metals contaminated soil but also its potential to advance the managements of associated wastes due to their threat to human well-being and the surroundings (Cristaldi et al. 2017). This technology has the advantages of being eco-friendly, aesthetically appealing, low cost, public acceptability, and energy efficient to remediate low to moderately heavy metal-contaminated sites (Gong et al. 2018). Other merits include its applicability on large area of contaminated site, minimization of generation of secondary pollutants, and minimal environmental disturbances during remediation process, and in addition, the soil can be used for agricultural purposes after cleanup. Meanwhile, the challenges of this method

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include lengthy growing season for the remediation (requires more time than the conventional techniques) to take place, contaminants can enter food chain when plants used are eaten by herbivores, and drought and climatic conditions may affect plant growth. The benefits and limitations of this effective remediating approach are discussed extensively in the later part of this chapter. Based on the methods and applicability, phytoremediation technology is additionally subdivided into various classifications, viz., phytoextraction, phytovolatilization, phytofiltration, phytostabilization, phytodegradation, and rhizosphere bioremediation.

16.5

Phytoextraction: A Green Remediation Method of Heavy Metal Pollution in Complex Industrial Waste Disposal Sites

Phytoextraction which is otherwise called phytoaccumulation, phytosequestration, or phytoabsorption is the major and most effective phytoremediation approach of removing heavy metal contaminants from contaminated soil or water using green plants without damaging the soil structure and productivity, thus contributing to durable cleanup of contaminated sites. This makes phytoextraction technique to have an edge in the cleanup of complex waste disposal sites over other remediative technologies. In this approach, crops of certain species known to be accumulators of heavy metals are employed to clean up the affected areas in soil by absorbing and translocating them through the roots system and store them in the parts (leaves and stems) above the ground level (Mahar et al. 2016; Yao et al. 2012), thereby removing the contaminant from the site. Effective phytoextraction as a promising economic and environmental cleanup technique depends on (1) the chemical characteristics of the heavy metals to be removed, (2) soil speciation and properties, (3) speciation of the heavy metals and plant species involved, and (4) availability of heavy metals for uptake as well as its absorption. Figure 16.2 shows the translocation and accumulation by plants in the harvestable components. Phytoextraction is economically feasible based on the fact that the harvested biomass is mostly composted or incinerated and not often recycled for reuse. Desirable characteristics to be owned by plants ideal for phytoextraction (Adesodun et al. 2010; Jabeen et al. 2009; Sakakibara et al. 2011; Seth 2012; Shabani and Sayadi 2012) are as follows: 1. 2. 3. 4.

Capability to grow outside their area of collection Rapid growth rate and increased production of above-ground biomass Profuse root structure for exploring large soil amounts of heavy metals Tolerant to toxicity impact of high concentrations of targeted metals in plant tissues 5. High translocation factor 6. Ease of cultivation and accumulation of heavy metals in their aerial parts

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Harvesting of metal impregnated shoots

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Post Harvesting (Thermal, Chemical or Microbial Processing)

Metal accumulation in impregnated shoots Biomass reduction for metals reclamation or disposal

Metal movement to shoots

Metal uptake from roots

Metals

Root exudates thereby increasing metal availability for root uptake

Fig. 16.2 Phytoextraction processes of metals in soil. (Modified after Nascimento and Xing 2006)

7. Adaptability to climatic and environmental conditions 8. Resistance to pests and pathogens 9. Repulsion to herbivores to prevent contamination of food chain However, no known plant presently fulfills all of these criteria as these traits are challenging to couple, but a fast-growing non-accumulator plant could be altered to deliver most of the above-mentioned characteristics.

16.6

Plants as Bioaccumulators of Heavy Metals

Plants used for phytoextraction have three main approaches for growth on heavy metal-contaminated soil as described in Fig. 16.3. These plants are specially modified to flourish in heavy metal-rich soils. They are mostly of the plant family Brassicaceae and are called “metallophytes.” One attractive idea is the use of metallophytes as a stand-alone, or by coupling it with microbes for phytoremediation of heavy metal-rich soils (Bothe 2011).

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Hyperaccumulator Metal concentration in plant

Fig. 16.3 Approaches of metal levels in plant vs. increasing total metal levels in soil. (Modified after Baker 1981)

B. Oladipo et al.

at dic n i l

M

or

a et

Metal excluder Metal concentration in soil

16.6.1 Metal Excluders Metal excluders impede the transportation and entrance of heavy metals into their aerial tissues but actively concentrate heavy metals into their roots while maintaining moderate and constant metal concentration (Malik and Biswas 2012; Sheoran et al. 2010). Excluders can be utilized for maintaining the soil to prevent advance dissemination of heavy metal contaminant to erosion.

16.6.2 Metal Indicators Metal indicators amass heavy metals in their aerial members and generally reflect heavy metal concentrations in the soil (Sheoran et al. 2010). They allow existing uniform concentration of heavy metals by exuding chelators (i.e., metal binding compounds) or modify heavy metal distribution sequence by storing metals in non-sensitive areas.

16.6.3 Metal Accumulator Plant Species or Hyperaccumulators Hyperaccumulators are plants that amass heavy metals in their aerial tissues to higher levels than those available in the soil, i.e., hypertolerant to heavy metals (Memon and Schröder 2009). Hyperaccumulators are considered as an exceptional

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and intense scenario of the wider class of accumulators (Pollard et al. 2002); this gives another possible reason why phytoextraction should be employed in the remediation of heavy metals from complex industrial waste disposal sites. The Brassicaceae family contains many metal accumulating plant species with widest range of metals (Poniedziałek et al. 2010). Examples of hyperaccumulators are Alyssum bertolonii and Thlaspi caerulescens. T. caerulescens (Alpine pennycress) is known to be the best metal hyperaccumulator for Zn, Cd, and Ni and for phytomining of valuable heavy metals, such as Pd, Pt, and Au (Assunção et al. 2003). One of the main advantages of hyperaccumulating plants is their resistance against pathogens and pests (by turning leaves toxic or distasteful) (Dipu et al. 2012; Meharg 2005). The volume of heavy metals evacuated from soil by hyperaccumulators is dependent on the amount of biomass generated multiplied by the tissue metal concentration (Macek et al. 2008). Table 16.4 gives some relevant plant species that are metal hyperaccumulators and their bioconcentration potential. Some of these plant species can be used to phytoextract heavy metal from heterogeneous waste disposal sites.

16.7

Mechanism of Phytoextraction

Various mechanisms have been studied to understand the development of plants in the heavy metal contaminated soils with no detrimental impacts on their growth. Some plants behave as metal excluders by restricting uptake of metals from roots to the shoots (Küpper et al. 1999), while some other plants show tolerance to high metal concentrations in their tissues via metal compartmentalization at cellular and molecular levels and binding of metals to organic compounds (Küpper et al. 1999; Peng et al. 2006). Four processes are widely considered to be pivotal for accumulation, namely, (1) root uptake, (2) transport of metals from roots to shoots, (3) complexation with chelating molecules, and (4) compartmentalization into the vacuole (Hall 2002; McGrath and Zhao 2003). Some of the various mechanisms embraced by plants for phytoextraction of heavy metal contaminated soils are illustrated in Table 16.5. Phytoextraction of heavy metals can be carried out in two different ways, viz., natural and induced methods. In natural phytoextraction (also known as continuous phytoextraction), plants are utilized for the removal of heavy metals based on their natural ability to accumulate, translocate, and resist high concentration of metals, that is, no soil alteration is done (Garbisu and Alkorta 2001). Hyperaccumulators are the mostly used plants suited for this type of phytoextraction of metal-polluted sites. Natural phytoextraction is mainly hindered by the moderate biomass production of hyperaccumulators. This may take several years to lessen the heavy metal concentration on the site to environmentally acceptable limits (Mahmood 2010). In induced or chelate-assisted phytoextraction, various synthetic chelators and organic acids such as ammonium sulfate, elemental sulfur, ethylene-diamine-tetra-

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Table 16.4 Plants that are metal hyperaccumulators Metal type Ni

Plant species Alyssum corsicum Alyssum bertolonii Alyssum caricum Alyssum murale Alyssum pterocarpum Alyssum serpyllifolium Alyssum markgrafii Alyssum heldreichii Alyssum murale Berkheya coddii

As

Cd

Pelargonium sp. Psychotria douarrei Sebertia acuminata Thlaspi caerulescens Pteris vittata L. Pteris vittata Pteris biaurita Pteris cretica Pteris ryukyuensis Pteris quadriaurita Thlaspi caerulescens Rorippa globosa Pelargonium sp. Atriplex halimus L. Alyssum murale

Metal accumulation (mg/kg) 18,100

References Li et al. (2003)

10,900

Li et al. (2003)

12,500

Li et al. (2003)

10,000–22,800 13,500

Li et al. (2003) and Chaney et al. (2008); Baker et al. (1988) Li et al. (2003)

10,000

Prasad (2005)

19,100

Bani et al. (2010)

11,800

Bani et al. (2010)

4730–20,100 >17,000 >1190 47,500

Bani et al. (2010) and Li et al. (2003) Mesjasz-Przybyłowicz et al. (2004), Robinson et al. (1997) and Moradi et al. (2010) Dan et al. (2002) Baker and Walker (1990)

250,000

Jaffré et al. (1976)

>10,000

Baker and Brooks (1989)

13,800–23,000 1000–8331 2000 1800–3030 3647

Ma et al. (2001) and Tu et al. (2002) Kalve et al. (2011) and Baldwin and Butcher (2007) Srivastava et al. (2006) Srivastava et al. (2006) and Zhao et al. (2002) Srivastava et al. (2006)

2900

Srivastava et al. (2006)

263–3000

Baker and Walker (1990) and Lombi et al. (2001)

>100 >750 830

Wei et al. (2008) Dan et al. (2002) Lutts et al. (2004)

1300

Bernal et al. (1994) (continued)

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Table 16.4 (continued) Metal type

Cr

Mn

Zn

Pb

Plant species Arabidopsis halleri Colocasia antiquorum Helianthus annuus L. Sedum alfredii Solanum photeinocarpum Azolla pinnata Eleocharis acicularis Leptospermum scoparium Pimelea suteri Pteris vittata Macadamia neurophylla Schima superba Eleocharis acicularis T. caerulescens Thlaspi caerulescens Sedum alfredii Polygonum aviculare Jatropha dioica Arabidopsis halleri Ambrosia artemisiifolia Armeria maritima Pisum sativum Sedum alfredii Fagopyrum esculentum Zea mays

Metal accumulation (mg/kg) 5722

References Küpper et al. (2000)

>400

Kashem et al. (2008)

>1500

Elkhatib et al. (2001)

570–1800 158

Yang et al. (2004) and Baker et al. (1994) Zhang et al. (2011)

740 239

Rai (2008) Sakakibara et al. (2011)

>30,000

Negri and Hinchman (1996)

30,000 20,675 51,800

Negri and Hinchman (1996) Kalve et al. (2011) Baker and Walker (1990)

62,412 11,200

Yang et al. (2008) Sakakibara et al. (2011)

>100,000 5100

Negri and Hinchman (1996) Baker et al. (1994)

5000–34,000 >9000

Yang et al. (2002), Baker et al. (1994) and Brown et al. (1995) González and González-Chávez (2006)

>6000 7429–32,000

González and González-Chávez (2006) Küpper et al. (2000) and Zhao et al. (2000)

>2000

Huang et al. (1997)

>10,000

Baker and Brooks (1989)

>6500 1182 4200

Huang et al. (1997) Yang et al. (2002) Tamura et al. (2005)

>2000

Huang et al. (1997) (continued)

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Table 16.4 (continued) Metal type Cu

Au

Plant species Aeollanthus biformifolius Haumaniastrum katangense Ipomoea alpine Brassica juncea

Metal accumulation (mg/kg) >1000

References Brooks et al. (1992)

>1000

Brooks et al. (1992)

12,300 100

Baker and Walker (1990) Anderson et al. (1999)

Table 16.5 Phytoextraction mechanisms of heavy metal-contaminated soil Plant Sedum alfredii

Metal Pb, Cd

Zn Alyssum species, Brassica juncea

Ni

Arabidopsis halleri Thlaspi goesingense

Cd, Zn

Pteris vittata

As

Thlaspi caerulescens, Thlaspi ochroleucum Brassica juncea

Zn, Cd, Cr, Cu, Ni, Pb

Ceratophyllum demersum

Ni

Cd

Cd

Mechanism Introduction and accumulation of phytochelatin that binds metals in above-ground parts Metals loaded into leaf sections and protoplast Attachment of the metals with histidine for detoxification Accumulation in trichomes and mesophyll cells Reducing the soil pH Release of ligands into rhizosphere Exploring more soil, increased colonization Reducing the pH of rhizosphere; hence, improving metal solubilization

Synthesis of phytochelatins (PCs), non-protein thiols, glutathione for metal binding in shoots glutathione reductase Production of phytochelatin for metal binding in shoots Activation of glutathione-S-transferase, glutathione, cysteine synthase

References Zhang et al. (2008)

Yang et al. (2006) Kerkeb and Krämer (2003) and Krämer et al. (1996) Küpper et al. (2000) Puschenreiter et al. (2003) and Wenzel et al. (2003) Leung et al. (2007) McGrath et al. (1997)

Seth et al. (2008)

Mishra et al. (2009)

acetic acid (EDTA), and citric acid are added to the soil to boost the metal phytoavailability and translocation from root to shoot (Lone et al. 2008; Salt et al. 1995; Sun et al. 2011). Induced phytoextraction employs high biomass production plants in which metal hyperaccumulation is prompted by soil amendments. The chelates form water-soluble complexes with the heavy metals in soil and aid in desorption from soil particles. Nevertheless, these chemical treatments can lead to environmental pollution of great concern. Proper measures should be put in place

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Table 16.6 Main features of the phytoextraction techniques Natural phytoextraction Plants are naturally metal accumulators Steady growth, moderate biomass production Effective uptake of metals from roots to shoots High resistance to high concentrations of metals in tissues Uptake of high concentration of metals from soils No environmental risk with respect to metals leaching

Induced phytoextraction Plants are typically metal excluders Rapid growing, significant biomass production Chemical alterations raise the metal transfer from roots to shoots Low resistance to metal, increase in absorption rate results to plant death Metal uptake is enhanced using synthetic chelators and organic acids Possibility of leaching of metal chelates to groundwater

when performing induced phytoextraction because synthetic chelating agents can also be harmful to plants in higher amounts (Song et al. 2012; Zhao et al. 2011). Main features of the two phytoextraction systems (Nascimento and Xing 2006) are presented in Table 16.6.

16.8

Advantages of Phytoextraction

Phytoextraction techniques offer numerous benefits, such as being a low-cost technique, an eco-friendly technology, and its applicability and effectiveness in reducing various range of contaminants, especially those peculiar to wastewater pollution and complex industrial waste disposal sites. One major benefit of using biomass in adsorption technology for heavy metals is its high efficiency in reducing the heavy metal ion concentration to lower levels coupled with the utilization of cheap biosorbent materials (Rakhshaee et al. 2009). Phytoextraction is inexpensive in approach and the cleanest remediation technology, majorly applicable to environmental media of large areas, such as complex industrial waste disposal sites which contain heterogeneous contaminants that have comparatively low to shallow levels of pollution and for large quantities of water with low amounts of pollutants (Mahar et al. 2016). It is less costly (60–80% or can be lesser) than conventional physical and chemical techniques, because it does not demand highly skilled personnel with no need for expensive equipment. Phytoextraction approach can also help to improve poor soils like those having high level of salts or aluminum. In comparison to the common techniques, in situ applications lower the level of disturbance in soil. Moreover, considering the agricultural practices involved, this technique which can be carried out with very low environmental effect offers dual benefits to the contaminated site as it initiates both detoxification of metals injurious to the soil and it subsequently enhances site restoration through preservation of the soil surface by plant cover (Bhargava et al.

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2012). The organic contaminants may be reduced to carbon dioxide and water, thereby getting rid of the toxic attributes from the environment (Mwegoha 2008). Another benefit of phytoextraction is the generation of metal-rich plant residue that are recyclable (Liu et al. 2000). To decontaminate heavy metals pollution in complex industrial waste disposal sites, phytoextraction strategy could be a better option especially when the biomass generated from the remediation process could be monetarily valorized as bioenergy. The utilization of metal-accumulating biofuel crops might be well-fitted for this function. By using oil crops to phytoremediate soils polluted by heavy metals, biodiesel synthesis from the subsequent plant oil could be a feasible choice to yield biofuel (Van Ginneken et al. 2007). In great-scale applications, thermal energy can then be generated from the stored potential energy (Mwegoha 2008). Furthermore, other than soil cleaning process, a technological-modified phytoextraction could demonstrate to be important in extracting valuable metals, for example, Ni, Au, and Pt, which might be reclaimed by phytomining for business purpose (Anderson et al. 2005; Kidd et al. 2009). Besides, economic feasibility of phytomining in combination with phytoextraction could offer a dual benefit to the contaminated site. The technical implication of this phytoextraction approach as an added advantage is that it requires little labor compared with other cleanup techniques used in contaminated site. Practices such as digging of the soil, landfilling, washing of soil using water, vitrification of polluted soil at high temperatures, chemical and electrical separation, and use of solubilizing and stabilizing agents for solidification are practically avoided (Cristaldi et al. 2017).

16.9

Limitations of Phytoextraction

Despite the fact that phytoextraction offers numerous appealing alternatives over conventional soil remediation systems, there are peculiar drawbacks to phytoextraction approach. Among them are being time consuming, applicability to sites with low to slight ranges of metal contaminant concentrations, possible adulteration of food chain, and ecological exposure. Phytoextraction may be time-intensive, and it can take numerous developing seasons to remediate a site to a satisfactory level. Extensive period of time may be needed to phytoextract a site in comparison with other cleanup strategies. In this manner, for contaminated areas that present intense danger to human well-being and different environmental receptors, phytoextraction may not be the cleaning system of choice (Mwegoha 2008). Phytoextraction may then be said to be maximally suitable for distant geographical location where human touch is restricted or where soil pollution does no longer require a prompt action (Salido et al. 2003). Besides, contingent upon the proposed future utilization of the land, the time necessities may be inhibitory. Specifically, property developers may not wish to delay for the several

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developing seasons required for phytoextraction to successfully tidy up the site (Schnoor et al. 1995). Contaminants with high level of concentration may restrain plant development; consequently, this may hinder the application of phytoextraction on the site or certain areas of the site. This phytotoxicity could prompt a curative method where wastes of high concentration are managed with costly ex situ procedures that rapidly lessen intense hazard. The optimum level that can be amassed by plants has in a way proved to be the main restriction in the use of phytoextraction of toxic metals. For example, plants with the highest level of toxic element content, regarded as “hyperaccumulators,” by and large show, on a dry weight basis, from about 0.2% (2000 ppm) and to greater than 2% for high toxic metals (e.g., Pb and Cd) and low toxic metals (e.g., Zn, Ni, and Cu), respectively (Tangahu et al. 2011). Hyperaccumulator plants most often amass a particular metal only, demonstrating a restricted applicability to areas with various blended contaminants. Phytoremediation is generally limited to treating soils with shallow contaminated sites employing species that are familiar with the climatic and environmental conditions of the locality. Phytoextraction is efficient if soil defilement and groundwater contamination are restricted to not more than 3 ft and 10 ft of the surface, respectively (Ghosh and Singh 2005). This limitation brings about its applicability majorly to large site areas with moderate soil contamination and to areas with huge quantities of groundwater containing low levels of contamination which must be remediated to allowable limits. This cleanup technique may likewise necessitate that the site be sufficiently enormous to allow for conventional cropping systems of agricultural practices. The surface topology of the site, instability of surface materials, and unstableness of slopes will affect effectiveness and cost. Phytoextraction may also require extensive input costs such as in pre-treatment of the waste material or the areas on which the wastes are discarded. Insecticides, artificial chelating agents, soil amendments, and irrigation may be needed in large amounts to properly clean the site. Furthermore, apart from the natural exposure which may happen at whatever point the plants are utilized to relate with the pollutants in the soil, another limitation of phytoextraction is the potential contamination of food cycle via animals or insects that feed on plant materials that may have been contaminated.

16.10

Factors Affecting Phytoextraction

Generally, factors limiting successful application of phytoextraction are concentration of soil contaminant, biomass production, and bioavailability of metals (Sheoran et al. 2016; Lasat 1999). However, the key factor affecting this technique is bioavailability of heavy metals in soil solution (Felix 1997). This can be described as the absorbable form of heavy element present in the soil which can be taken up by plant. Bioavailability of metals to plants is affected by both soil-associated (external) factors and plant-associated (internal) factors (Sheoran et al. 2016). Soil-associated

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factors include cation exchange capability, soil type, redox potential, soil moisture, soil pH, and biochemical operations, while plant-associated factors are transpiration rate, root absorption factor, acidification of rhizosphere, rhizosphere microorganisms, plant root depth, and plant density. Phytoextraction is limited by the plant root depth which is a major plantassociated factor. Root contact is a common impediment on the usage of phytoextraction as most of the heavy metal remediative plants employed have shallow depth of root penetration. For effective remediation with plants to take place, it is necessary that contaminants must interact within the root region of the plants. There might be need to alter the ground surface of the site to avoid erosion or flooding since the technique is restricted to sites with little contamination within the rooting region of accumulating plants (Mwegoha 2008). In addition, age influences the normal functioning of a plant, particularly its roots. Mostly, roots of a youthful plant show higher capacity to absorb metal ions in comparison to an older plant of similar size. It is essential to utilize rich youthful plants for progressively effective outcome; notwithstanding, this does not eliminate the utilization of bigger aged plants whose bigger size may make up for their less impactful physiological behavior when contrasted with little more youthful plants (Tu et al. 2004). For any cleanup strategy, the challenges related with waste segregation and the impacts of mixed site conditions likewise go for phytoextraction (Glass 1999). The soil-associated factors must all be within the specific boundaries of plants tolerance. Contaminants that have high solubility may leach outside the root zone making less effective plant uptake. Microclimates on the site may also be unfavorable, thereby leading to small amounts of germination. Biological interactions of species may have an inducing or constraining impact on cleanup. For instance, plants used for cleanup may be specifically susceptible to pests and herbivores. Sufficient time and efforts are in this manner required to build up a total informative knowledge of biological mechanisms and to carry out comprehensive design investigation for biological remediation processes. Other factors such as biomass production affect phytoextraction efficiency. Several metal hyperaccumulator plants have slacken development rates and generate little quantity of biomass, while other plants that are fast growing have less heavy metal extracting capacity (Zhuang et al. 2007). Water and weather-related conditions may hinder the pace of development of plants that can be used. In addition, metal detainment within the root and absorption rate of metal in the root (Sharma and Pandey 2014) are other parameters affecting metal phytoextraction. Additionally, numerous hyperaccumulator plants are generally uncommon, with little populaces that usually grow up in remote areas. This makes them to be very restricted in distribution (Brooks 1998) and, thus, limit the usage of phytoextraction in affected areas. The use of untamed or non-indigenous plants for phytoextraction likewise brings up specific concerns of hazard evaluation and may influence biodiversity (Mwegoha 2008).

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363

Conclusions and Future Investigation

Soil and water contaminated by heavy metals from complex industrial waste disposal sites is a major threat in every part of the world; consequently, efficient cleanup techniques are essential. Phytoremediation is eco-benign with great acceptance in the environment. The prospect of phytoextraction of heavy metals as a major phytoremediation technique is expected to be a market plausible technology in years to come due to its numerous advantages over its few limitations. For future prospects, optimization of agricultural practices is essential to increase the uptake potential of new hyperaccumulative plants because metal sorption in roots is most often hindered by low solubility in soil solution. To shorten the extensive time in developing season of phytoremediative plants and maximize the use of the growing time, optimization of the harvest time is important. Also, it is necessary to identify cheap and ecofriendly chemical substances with standard metal chelating qualities to promote the bioavailability of metals. In addition, identification of plant species with the ability of being rotated is needed to sustain the rate of metal removal, and detailed environmental and risk assessment investigations should be carried out before the use of foreign plant species for phytoextraction.

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Yang X, Li T, Yang J, He Z, Lu L, Meng F (2006) Zinc compartmentation in root, transport into xylem, and absorption into leaf cells in the hyperaccumulating species of Sedum alfredii Hance. Planta 224(1):185–195. https://doi.org/10.1007/s00425-005-0194-8 Yang S, Deng H, Li M (2008) Manganese uptake and accumulation in a woody hyperaccumulator, Schima superba. Plant Soil Environ 54(10):441–446. https://doi.org/10.17221/401-PSE Yao Z, Li J, Xie H, Yu C (2012) Review on remediation technologies of soil contaminated by heavy metals. Procedia Environ Sci 16:722–729. https://doi.org/10.1016/j.proenv.2012.10.099 Yoshinaga M, Ninomiya H, Al Hossain MA, Sudo M, Akhand AA, Ahsan N, Alim MA, Khalequzzaman M, Iida M, Yajima I (2018) A comprehensive study including monitoring, assessment of health effects and development of a remediation method for chromium pollution. Chemosphere 201:667–675. https://doi.org/10.1016/j.chemosphere.2018.03.026 Zhang Z, Gao X, Qiu B (2008) Detection of phytochelatins in the hyperaccumulator Sedum alfredii exposed to cadmium and lead. Phytochemistry 69(4):911–918. https://doi.org/10.1016/j. phytochem.2007.10.012 Zhang X, Xia H, Za L, Zhuang P, Gao B (2011) Identification of a new potential Cd-hyperaccumulator Solanum photeinocarpum by soil seed bank-metal concentration gradient method. J Hazard Mater 189(1–2):414–419. https://doi.org/10.1016/j.jhazmat.2011.02.053 Zhao F, Lombi E, Breedon T (2000) Zinc hyperaccumulation and cellular distribution in Arabidopsis halleri. Plant Cell Environ 23(5):507–514. https://doi.org/10.1046/j.1365-3040. 2000.00569.x Zhao F, Dunham S, McGrath S (2002) Arsenic hyperaccumulation by different fern species. New Phytol 156(1):27–31. https://doi.org/10.1046/j.1469-8137.2002.00493.x Zhao H-Y, Lin L-J, Yan Q-L, Yang Y-X, Zhu X-M, Shao J-R (2011) Effects of EDTA and DTPA on lead and zinc accumulation of ryegrass. J Environ Prot 2(07):932. https://doi.org/10.4236/ jep.2011.27106 Zhuang P, Yang Q, Wang H, Shu W (2007) Phytoextraction of heavy metals by eight plant species in the field. Water Air Soil Pollut 184(1–4):235–242. https://doi.org/10.1007/s11270-007-9412-2

Chapter 17

Biosorption of Nickel (II) and Cadmium (II) Rajeswari M. Kulkarni

, K. Vidya Shetty

, and G. Srinikethan

Contents 17.1 17.2 17.3

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Heavy Metals: Environmental Threat . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicological Aspects of Nickel and Cadmium and the Need for Removal from Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.3.1 Nickel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.3.2 Cadmium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4 Standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.5 Physicochemical Remediation Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.6 Biosorption as a Remediation Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.7 Biosorption Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.8 Microorganisms Used in Biosorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.9 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

374 374 375 375 376 378 378 379 379 380 384 384

Abstract Water is a finite limited resource in this planet. This is being threatened both quantitatively and qualitatively by the stress brought about by increasing population and industrial activities. Heavy metals are natural elements extracted from earth and exploited in industrial applications which, generally, have high densities. Among the many heavy metals present in wastewater, cadmium [Cd (II)] and nickel [Ni (II)] are present in industrial wastewater generated from a wide variety of processing applications (smelting, mining, chemical processing, R. M. Kulkarni (*) Department of Chemical Engineering, M. S. Ramaiah Institute of Technology, Bangalore, Karnataka, India Department of Chemical Engineering, National Institute of Technology, Surathkal, Karnataka, India e-mail: [email protected] K. Vidya Shetty · G. Srinikethan Department of Chemical Engineering, National Institute of Technology, Surathkal, Karnataka, India e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 373 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_17

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electroplating etc.). Heavy metals Ni (II), Cd (II), and their compounds are hazardous. This necessitates the removal of Ni (II) and Cd (II) ions from wastewater. Biosorption has emerged as a promising technology for exploring algae, bacteria, fungi, and yeasts as a potential metal biosorbents. The review begins with introduction on nickel (II) and cadmium (II) heavy metals, their applications, toxicity, and discharge standards. Different conventional techniques applied for heavy metal remediation have been discussed. Reports about biosorption process, mechanism, and biosorption capacity of microorganisms employed for remediation of Ni (II) and Cd (II) are presented. Keywords Heavy metals · Nickel · Cadmium · Applications · Toxicity · Health effects · Environmental effects · Biosorption · Biosorbents · Mechanisms

17.1

Introduction

Water, one of the most important resource for the mankind, is available but in limited quantity. The release of heavy metals into the water will pose a serious problem to the ecosystem. Metal ions present in wastewater are categorized by their movement in the ecosystem and their impact on the higher forms of life, even at lower concentrations (Volesky and Holan 1995). Moreover, these ions are non-biodegradable, chronic, and lead to ecological and health issues (Volesky 2001). The progressive awareness of the ill-effects of untreated discharges and the anti-pollution laws from governments has helped in controlling the problem to some extent (Gadd and White 1993; McHale and McHale 1994). The penalizations for the release of untreated water pose huge threat financially for the industrialists and the industries are looking for financially sustainable solutions. Development of efficient and economical treatment technologies may lead to reuse of treated effluents and hence offer solutions to wastewater generation problem through efficient use of water and minimization of wastewater generation (Jooste 2000). Nickel and cadmium are widely used together in many industries and are potential hazardous metals if discharged in large quantity.

17.2

Heavy Metals: Environmental Threat

“Heavy metals,” are metallic chemical elements that are relatively higher in density (Passow et al. 1961). Heavy metals even at dilute concentrations are toxic (Gadd and White 1993; Febrianto et al. 2009). Among various heavy metals, cadmium, mercury, lead, and chromium are considered highly toxic. Metals like copper, cobalt, nickel, and zinc are not regarded toxic as the former, but their widespread usage and

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increasing levels in the environment are causing a serious concern (Brown and Ahsanullah 1971). Due to the exponentially growing population and the growth of industrial production to meet their requirements, there is increasing risk of heavy metal contamination in our ecosystem. Increased industrial activity has led to increased discharge of heavy metals into soil and water, thus posing a serious risk to biota (Veglio and Beolchini 1997). Hence, there is an increasing necessity to control the discharge of metals at the source of emissions. The toxic metals penetrate in the food chain, and humans eventually are at highest risk (Volesky 2001).

17.3

Toxicological Aspects of Nickel and Cadmium and the Need for Removal from Wastewater

17.3.1 Nickel Nickel is a silvery-white lustrous metal with atomic number 28 and atomic weight 58.69. It has a density of 8.9 g/cm3 at 20  C. It is a fair conductor of electricity and heat. Nickel forms a part of group of transition metals; it is hard and ductile in its nature. The oxidation state that nickel is most commonly found is +2. The consumption of nickel in industrial applications may be distributed in making nickel steels, special alloys, plating, tinted glass, and batteries (Davis 2000; Alyüz and Veli 2009). Nickel is also used as a catalyst in many important chemical reactions (Davis 2000; Padmavathy et al. 2003).

Health Effects Plants accumulate nickel, and hence nickel uptake by humans from vegetables and fruits is obvious (Somers 1974). Humans may get exposed to nickel by consuming contaminated water or food, smoking, breathing contaminated air, etc. (Mastromatteo 1967; Lisk 1972; Bencko 1983). Nickel is essential to humans when consumed in small quantities; however, a larger amount of uptake is dangerous to health. If Ni (II) is ingested through water, in higher concentration, it may cause (i) severe damage to kidneys and lungs, (ii) birth defects, (iii) heart disorders, (iv) gastrointestinal distress, (v) respiratory failure, (vi) skin dermatitis (also known as nickel itch) in sensitized individuals, (vii) pulmonary fibrosis, etc. (Sunderman Jr 1976; Subbaiah and Yun 2013).

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Environmental Effects Nickel is an essential metal for plants and is metabolically important as long as it is available for plants in required levels. However, it can become toxic to most plant species when concentration exceeds in soil and water. Higher nickel concentration can severely retard seed germination of many crops and also retards shoot and root growth at vegetative stages. This can affect branching development; deform the plant parts including producing abnormal flower shape and disturbing mitotic root tips; decrease biomass production; induce leaf spotting; and produce iron deficiency that leads to chlorosis and foliar necrosis. Besides this, excess nickel can affect nutrient absorption by roots, impairs plant metabolism, inhibits photosynthesis and transpiration, and causes ultrastructural modifications (Mishra and Kar 1974; Chen et al. 2009). Also, excessive nickel exposure can also alter processes and result in reduced yields of agricultural crops (Ahmad et al. 2011). Nickel disturbs plant growth, reducing seed germination and affecting the production of proteins, enzymes, and chlorophylls (Seregin and Kozhevnikova 2006). The effect of nickel on plants is evident however the information on its effect upon organisms other than humans is not much available. Algae growth is affected due to high nickel concentration in soil and surface waters. Microorganisms also suffer from excessive presence of nickel, but they develop resistance, eventually. In case of animals, nickel is favorable as an essential element, when exposed in small quantities (Nielson and Ollerich 1974; Nielsen et al. 1974). However, when the amount of nickel exposure exceeds maximum tolerable amounts in animals, nickel can be very harmful and dangerous; it can cause cancer (Sunderman Jr 1976).

17.3.2 Cadmium Cadmium is a bluish-white metallic element with atomic number 48 and atomic weight 112.41. Its density is 8.65 g/cm3 at 20  C. +2 is the most common oxidation state of cadmium. Cadmium has several physical and chemical properties that are desirable and industry-friendly, viz., resistance to corrosion and chemicals, withstand high temperatures, good electrical conductivity, and a low melting point that make it one of the widely used elements in industrial and consumer applications (Morrow 2000). Due to these properties, cadmium is found suitable for use in Ni-Cd batteries, stabilizers, coatings, pigments, and alloys (Page and Bingham 1973; Morrow 2000; Wang et al. 2010).

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Health Effects Cadmium is a hazardous metal. Human exposure to cadmium is primarily from industries like steel, cement production and related activities, nonferrous metals, electroplating, and Ni-Cd batteries (Friberg et al. 1974). Cadmium uptake in humans happens mainly through food and water. Cadmium-rich food stuff like shell fish, mussels, mushrooms, cocoa powder, dried seaweed, etc. can significantly increase cadmium concentration in human bodies (Satarug et al. 2009). Blood transports cadmium to all parts of the body including the liver. There it forms complexes, then transported to the kidneys, and damages filtering mechanism resulting in the excretion of essential proteins and sugars from the body. Higher level of cadmium in the body leads to brittle bones. The damage to the kidney and brittle bones in combination, due to cadmium exposure, is called as itai-itai disease. Even exposure to low cadmium levels, chronically over a period, may also result in progressive lung diseases, heart diseases, anemia, depressed immune system response, skeletal weakening, kidney and liver diseases, etc. Some other health effects of cadmium on humans are diarrhea, severe vomiting and stomach pain, infertility, and damage to the central nervous system (Flick et al. 1971; Ryan et al. 1982; Järup et al. 1998).

Environmental Effects Cadmium has serious effects on our environment. Industrial effluent streams with cadmium waste mainly disposed in soils. Cadmium impacts seed germination, plant growth, lipid content, and phytochelatin production (Das et al. 1997; Seregin and Ivanov 2001). Cadmium absorbs the organic matter from soil. Soil that is acidic enhances the cadmium absorption by the plants. This causes serious endangerment to the animals that depend on plants for their prey. When animals consume these plants, cadmium accumulates in their bodies causing high blood pressures, liver disease, and nerve and brain damage (Dudka and Miller 1999; de Vries et al. 2007). Soil organisms like earthworms are extremely susceptible to cadmium poisoning. Earthworms are very sensitive to this contamination, and this has major consequence on the whole soil ecosystem (Spurgeon et al. 1994; Das et al. 1997). Excess cadmium in aquatic ecosystems can bio-accumulate, and the susceptibility to excess cadmium varies from organism to organism. Saltwater life is more resistant to cadmium poisoning than freshwater organisms (Eisler 1971; Wright and Welbourn 1994).

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Standards

The drinking water standards and industrial discharge standard for effluents containing nickel and cadmium prescribed by various regulatory bodies are given in Tables 17.1 and 17.2, respectively.

17.5

Physicochemical Remediation Processes

The common physicochemical treatment processes that are used to treat the effluents containing Ni (II) and Cd (II) are chemical precipitation (Charerntanyarak 1999), ion exchange (Maliou et al. 1992), adsorption (Huang and Ostovic 1978; Ho et al. 1995), membrane processes (Qdais and Moussa 2004), etc. Most of the aforementioned processes have serious drawbacks that limit their practical application. Precipitation is a treatment process that is generally applied because of its simple equipment, being capable of treating large volume at a low cost (Ahalya et al. 2003; Kurniawan et al. 2006). Precipitation leaves a large amount of sludge, which is hard to handle and is not capable to treat the water to the required standards. Expensive membrane separation and ion exchange processes face membrane fouling problems. Another well-recognized method for heavy metal removal from wastewater is adsorption on activated carbon; however, its high cost restricts its application (Pollard et al. 1992; Lalvani et al. 1998; Volesky 2001; Fu and Wang 2011). This has led many researchers to search for low-cost alternative sources from industrial and agricultural wastes such as red mud (Apak et al. 1998), sawdust Table 17.1 Drinking water standards for nickel (II) and cadmium (II)

Metal Nickel (II) (mg/L) Cadmium (II) (mg/L)

IS:10500 BIS (Bureau of Indian Standards) 0.02

EU (European Union) 0.02

USEPA (US Environment Protection Agency) 0.1

WHO (World Health Organization) 0.07

0.003

0.005

0.005

0.003

Table 17.2 Industrial discharge standards for nickel (II) and cadmium (II)

Metal Nickel (II) (mg/L) Cadmium (II) (mg/L)

CPCB (Central Pollution Control Board) 3

EU (European Union) 0.2

USEPA (US Environment Protection Agency) 0.1

WHO (World Health Organization) 0.5

2

0.05

0.01

0.1

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(Taty-Costodes et al. 2003), fly ash (Srivastava et al. 2006), lignin (Guo et al. 2008), rice straw (Ding et al. 2012), sugarcane bagasse (Alomá et al. 2012), baker’s yeast (Wang 2012), zeolite (Kulkarni et al. 2014b), Saccharomyces cerevisiae (Galedar and Younesi 2013), brewery sludge (Kulkarni et al. 2019), etc. However, due to low metal uptake, large amounts of adsorbents are needed. Biosorption has been tested for metal removal and found to be successful in treating the metal-contaminated water (Beveridge 1989; Lovley et al. 1991; Wang and Chen 2006).

17.6

Biosorption as a Remediation Process

Biosorption offers several advantages including low operating cost, high efficiency of treating dilute metal ion solutions, biosorbents regeneration and metal recovery opportunity, rapid kinetics of adsorption and desorption, no sludge generation, etc. Biosorption process deals with the interaction of functional groups of biological material and the contaminating element in the wastewater (Gadd 1990; Kratochvil and Volesky 1998). It has been found that both living and dead microbes adsorb metal ions. Microorganisms bind heavy metals either actively, called as bioaccumulation, or passively, called as biosorption, or with a combination of both processes (Naja and Volesky 2011). Biosorption is a rapid process and is not dependent on the presence of specific nutrients. They can be regenerated and reused for many cycles. However, bioaccumulation is relatively a slow process and has dependence on the presence of nutrients. Dead cells have proved to be more effective in treating large-scale metal contaminated water than the growing or resting cells (Gabr et al. 2008; Ahmad et al. 2019).

17.7

Biosorption Mechanisms

The metal biosorption is a complicated process influenced by the types of biomass, chemistry of metal biomass, and environmental conditions. Figure 17.1 shows various ways by which metal can be captured by the cell (Veglio and Beolchini 1997; Taty-Costodes et al. 2005). Metal sequestering happens in two modes – (i) energy-independent passive mode by dead or inactive cells and (ii) energy-dependent active mode by living cells. Physical adsorption, ion exchange, complexation, and precipitation may be responsible for metabolism-independent metal sequestering. Active mode of biosorption involves passive uptake of metal ions onto the cell walls or on to extracellular polymeric substances (EPS) secreted by cells and also intracellular metal transport (Gadd 1990; Veglio and Beolchini 1997).

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BIOSORPTION MECHANISMS

METABOLISM DEPENDENT

Transport across cell membrane (Intracellular accumulation) Extra cellular accumulation (through EPS) Precipitation(Intracellular / extracellular)

NON METABOLISM DEPENDENT

Physical adsorption Ion exchange Complexation Precipitation

Fig. 17.1 Biosorption mechanisms based on dependency on cell metabolism and on location where the metal removed is found. (Modified after Veglio and Beolchini 1997)

Biosorption of metal ions onto cell surface is of two types: physisorption and chemisorption. In physisorption, van der Waals’ force (electrostatic interaction) is involved in binding metal ions on to the cell surface (Kuyucak and Volesky 1988; Tsezos and Volesky 1982; Pradhan et al. 2019). Chemisorption takes place through ion exchange, complexation, and precipitation mechanism (Veglio and Beolchini 1997; Escudero et al. 2019).

17.8

Microorganisms Used in Biosorption

Over the last few decades, adsorption using microbial biomass (bacteria, fungi, yeast, and algae) as biosorbent has emerged as a potential technique alternative to the existing methods for the removal of Ni (II) and Cd (II). Many bacteria have shown excellent Ni (II) and Cd (II) metal treatment ability which include genre of Bacillus (Hu et al. 2007; Oves et al. 2013; Kulkarni et al. 2014a, b; Kim et al. 2015; Ramya and Thatheyus 2018), Pseudomonas (Cabral 1992; Pardo et al. 2003; Gabr et al. 2008), and Streptomyces (Mattuschka and Straube 1993; Puranik et al. 1995; Veneu et al. 2012), etc. Fungi and yeast groups including genera Rhizopus (Tobin et al. 1984), Penicillium (Siegel et al. 1983; Fan et al. 2008), Aspergillus (Panday and Banerjee 2012), Streptomyces (Puranik et al. 1995; Selatnia et al. 2004), and Saccharomyces (Galedar and Younesi 2013) also proved to be capable of detoxifying the metals.

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There are numerous research reports that are available on using algal biomass for metal treatment (Davis et al. 2003; Romera et al. 2006). The algal groups include green algae (Hashim and Chu 2004; Aksu and Donmez 2006), brown algae (Holan et al. 1993; Romera et al. 2007), and red algae (Prasher et al. 2004; Sarı and Tuzen 2008). Brown alga stands out as a very good biosorbent for heavy metal removal. The walls of the cell contain fucoidan and alginic acid (Romera et al. 2006). Wastes from industrial operations containing microbial biomass such as food and fermentation wastes (El-Sayed and El-Sayed 2014; Kulkarni et al. 2019), activated sludge process wastes (Al-Qodah 2006; Hammaini et al. 2007), pharmaceutical wastes (El-Sayed and El-Sayed 2014), agricultural wastes (Sud et al. 2008; Mohammed et al. 2014; Vijayaraghavan et al. 2016), etc. have been widely investigated as potential biosorbents for heavy metals. Several researchers have adopted different microbial biomass as biosorbent for the removal of Ni (II) and Cd (II) from contaminated water. Tables 17.3 and 17.4 list these biosorbents and the uptake efficiency. Table 17.3 Cadmium (II) uptake by various microbial biosorbents Biomass type Penicillium notatum Rhizopus arrhizus Bacillus subtilis Penicillium digitatum Ascophyllum nodosum Sargassum natans Fucus vesiculosus Streptomyces noursei Saccharomyces cerevisiae Penicillium chrysogenum Rhizopus nigricans Streptomyces pimprina Pseudomonas aeruginosa PU21 Pseudomonas putida Live cells of Mucor rouxii Dead cells of Mucor rouxii Sargassum baccularia Chaetomorpha linum Gracilaria changii Palmaria palmata Streptomyces rimosus

Biomass class Fungus Fungus Bacteria Fungus Brown marine alga Brown marine alga Brown marine alga Bacteria

Metal uptake (mg/g) 5 30 101 3.5 215

References Siegel et al. (1983) Tobin et al. (1984) Brierley (1985) Galun et al. (1987) Holan et al. (1993)

135

Holan et al. (1993)

73

Holan et al. (1993)

3.4

Mattuschka and Straube (1993) Volesky et al. (1993) Volesky and Holan (1995) Volesky and Holan (1995) Puranik et al. (1995) Chang et al. (1997)

Yeast Fungus Fungus Bacteria Bacteria

20–40 56 19 30.4 42.4

Bacteria Fungus Fungus Brown algae Green algae Red algae Red algae Bacteria

8.0 8.46 8.36 35.52 23.04 11.04 4.75 64.9

Pardo et al. (2003) Yan and Virarghavan (2003) Hashim and Chu (2004) Hashim and Chu (2004) Hashim and Chu (2004) Prasher et al. (2004) Selatnia et al. (2004) (continued)

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Table 17.3 (continued) Biomass type Bacillus laterosporus Bacillus licheniformis Bacillus circulans Chlorella vulgaris Enterobacter sp. J1 Penicillium simplicissimum Ceramium virgatum Geobacillus thermodenitrificans Aspergillus aculeatus DBF9 Streptomyces lunalinharesii Exiguobacterium Saccharomyces cerevisiae

Biomass class Bacteria Bacteria Bacteria Green alga Bacteria Fungus Red algae Bacteria

Metal uptake (mg/g) 159.5 142.7 26.5 86.6 46.2 61.35 39.7 37.86

Fungi

14.24

Bacteria

24.8

Bacteria Yeast

43.5 1.179

Pseudomonas florescence Bacillus thuringiensis Geobacillus thermantarcticus Ochrobactrum sp. Halomonas Bacillus laterosporus Lessonia nigrescens Durvillaea antarctica Trichoderma sp. Ulva lactuca Exiguobacterium Trichoderma sp. Scenedesmus obliquus Hydrodictyon reticulatum Bacillus cereus AVP12 Bacillus cereus NC7401 Microbacterium sp. MC3B10 Parachlorella sp. Nannochloropsis oculata

Bacteria Bacteria Bacteria

52.6 59.17 6.26

Bacteria Bacteria Bacteria Algae Algae Fungi Algae Bacteria Fungi Algae Algae Bacteria Bacteria Bacteria

83.33 12.023 85.47 115.3 106.4 133.33 84.6 15.6 0.45 144.93 12.74 434.0 212.7 97

Algae Algae

96.20 232.55

Bacteria Yeast Fungi

9.86 13.96 71.43

Bacteria

131.58

Streptomyces rimosus Brewery sludge Phanerochaete chrysosporium Bacillus badius

References Zouboulis et al. (2004) Yilmaz and Ensari (2005) Aksu and Donmez (2006) Lu et al. (2006) Fan et al. (2008) Sarı and Tuzen (2008) Chatterjee et al. (2010) Panday and Banerjee (2012) Veneu et al. (2012) Alam and Ahmad (2013) Galedar and Younesi (2013) Safari et al. (2013) Oves et al. (2013) Özdemir et al. (2013) Khadivinia et al. (2014) Rajesh et al. (2014) Kulkarni et al. (2014a) Gutiérrez et al. (2015) Bazrafshan et al. (2016) Ibrahim et al. (2016) Park and Chon (2016) Rahman et al. (2016) Zhang et al. (2016) Ammari et al. (2017) Akhter et al. (2018) Camacho-Chab et al. (2018) Dirbaz and Roosta (2018) Kaparapu and Prasad (2018) Yous et al. (2018) Kulkarni et al. (2019) Noormohamadi et al. (2019) Vishan et al. (2019)

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Table 17.4 Nickel (II) uptake by various microbial biosorbents Biomass type Bacillus subtilis

Metal uptake (mg/g) 6

References Beveridge (1986)

29

Beveridge (1986)

6 18

Cabral (1992) Fourest and Roux (1992) Fourest and Roux (1992) Mattuschka and Straube (1993) Holan and Volesky (1994) Holan and Volesky (1994) Holan and Volesky (1994) Holan and Volesky (1994) Holan and Volesky (1994) Volesky and Holan (1995) Volesky and Holan (1995) Veglio et al. (1997) Tan and Cheng (2003)

Pseudomonas syringae Rhizopus arrhizus

Biomass class Bacterial cell walls preparation Bacterial cell walls preparation Bacteria Fungus

Rhizopus arrhizus

Fungus

18.7

Streptomyces noursei

Bacteria

0.8

Ascophyllum nodosum

Brown marine alga

30

Fucus vesiculosus

Brown marine alga

40

Sargassum natans

Brown marine alga

24–44

Ascophyllum nodosum

Algae

70

Fucus vesiculosus

Algae

17

Absidia orchidis

Fungus

5

Rhizopus nigricans

Fungus

5

Arthrobacter sp. Penicillium chrysogenum Live cells of Mucor rouxii Dead cells of Mucor rouxii Palmaria palmata Streptomyces rimosus Chlorella vulgaris Mucor rouxii Bacillus thuringiensis Saccharomyces cerevisiae Geobacillus thermoleovorans Bacillus thuringiensis Bacillus laterosporus Sinorhizobium

Bacteria Fungus

13 19.2

Fungus

11.09

Fungus

6.34

Red algae Bacteria Green alga Fungus Bacteria Yeast

3.03 32.6 58.4 6.34 45.9 1.683

Bacteria

2.19

Öztürk (2007) Galedar and Younesi (2013) Özdemir et al. (2013)

Bacteria Bacteria Bacteria

43.13 44.44 25.13

Oves et al. (2013) Kulkarni et al. (2014a) Jobby et al. (2015)

Bacillus licheniformis

Yan and Virarghavan, (2003)

Prasher et al. (2004) Selatnia et al. (2004) Aksu and Donmez (2006)

(continued)

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Table 17.4 (continued) Biomass type Bacillus cereus Mucor indicus chitosan Pleurotus mutilus Streptomyces roseorubens Streptomyces rimosus Sargassum sp. Brewery sludge Sargassum filipendula Phanerochaete chrysosporium

17.9

Biomass class Bacteria Fungi Fungi Bacteria

Metal uptake (mg/g) 344.80 10.4 47.169 208.39

Bacteria Algae Yeast Algae Fungi

22.8 67.31 7.87 37.03 46.50

References Naskar et al. (2016) Ruholahi et al. (2016) Daoud et al. (2018) Long et al. (2018) Yous et al. (2018) Barquilha et al. (2019) Kulkarni et al. (2019) Moino et al. (2019) Noormohamadi et al. (2019)

Conclusions

This chapter presents a brief review on the studies carried out on the processes involving heavy metal removal from aqueous solutions. The major focus was updating information on nickel (II) and cadmium (II) which are in larger usages together, and they are non-degradable in nature. The technology of metal treatment in wastewater of process industries has grown from conventional technologies to biosorption. Biosorption of heavy metals from aqueous solutions, a relatively new route, confirms as a very promising cost-effective method in the removal of heavy metal contaminants. Recently there has been lot of attention on use of dead and live biomass as biosorbents for detoxification of metals. This review also consolidated information on biosorption capacity of numerous microbial biosorbents which have been explored for Ni (II) and Cd (II) removal by various researchers.

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Veglio F, Beolchini F (1997) Removal of metals by biosorption: a review. Hydrometallurgy 44 (3):301–316. https://doi.org/10.1016/S0304-386X(96)00059-X Veglió F, Beolchini F, Gasbarro A (1997) Biosorption of toxic metals: an equilibrium study using free cells of Arthrobacter sp. Process Biochem 32(2):99–105. https://doi.org/10.1016/S00329592(96)00047-7 Veneu DM, Pino GA, Torem ML, Saint’Pierre TD (2012) Biosorptive removal of cadmium from aqueous solutions using a Streptomyces lunalinharesii strain. Miner Eng 29:112–120. https:// doi.org/10.1016/j.mineng.2011.08.005 Vijayaraghavan K, Rangabhashiyam S, Ashokkumar T, Arockiaraj J (2016) Mono-and multicomponent biosorption of lead (II), cadmium (II), copper (II) and nickel (II) ions onto cocopeat biomass. Sep Sci Technol 51(17):2725–2733. https://doi.org/10.1080/01496395.2016. 1212889 Vishan I, Saha B, Sivaprakasam S, Kalamdhad A (2019) Evaluation of Cd (II) biosorption in aqueous solution by using lyophilized biomass of novel bacterial strain Bacillus badius AK: biosorption kinetics, thermodynamics and mechanism. Environ Technol Innov 14:100323. https://doi.org/10.1016/j.eti.2019.100323 Volesky B (2001) Detoxification of metal-bearing effluents: biosorption for the next century. Hydrometallurgy 59(2–3):203–216. https://doi.org/10.1016/S0304-386X(00)00160-2 Volesky B, Holan ZR (1995) Biosorption of heavy metals. Biotechnol Prog 11(3):235–250. https:// doi.org/10.1021/bp00033a001 Volesky B, May H, Holan ZR (1993) Cadmium biosorption by Saccharomyces cerevisiae. Biotechnol Bioeng 41(8):826–829. https://doi.org/10.1002/bit.260410809 Wang Y (2012) Optimization of cadmium, zinc and copper biosorption in an aqueous solution by Saccharomyces cerevisiae. Int J Chem 1:1–3 Wang J, Chen C (2006) Biosorption of heavy metals by Saccharomyces cerevisiae: a review. Biotechnol Adv 24(5):427–451. https://doi.org/10.1016/j.biotechadv.2006.03.001 Wang FY, Wang H, Ma JW (2010) Adsorption of cadmium (II) ions from aqueous solution by a new low-cost adsorbent—bamboo charcoal. J Hazard Mater 177(1–3):300–306. https://doi.org/ 10.1016/j.jhazmat.2009.12.032 Wright DA, Welbourn PM (1994) Cadmium in the aquatic environment: a review of ecological, physiological, and toxicological effects on biota. Environ Rev 2(2):187–214. https://doi.org/10. 1139/a94-012 Yan G, Viraraghavan T (2003) Heavy-metal removal from aqueous solution by fungus Mucorrouxii. Water Res 37(18):4486–4496. https://doi.org/10.1016/S0043-1354(03)00409-3 Yilmaz EI, Ensari NY (2005) Cadmium biosorption by Bacillus circulans strain EB1. World J Microbiol Biotechnol 21(5):777–779. https://doi.org/10.1007/s11274-004-7258-y Yous R, Mohellebi F, Cherifi H, Amrane A (2018) Competitive biosorption of heavy metals from aqueous solutions onto Streptomyces rimosus. Korean J Chem Eng 35(4):890–899. https://doi. org/10.1007/s11814-018-0004-1 Zhang X, Zhao X, Wan C, Chen B, Bai F (2016) Efficient biosorption of cadmium by the selfflocculating microalga Scenedesmus obliquus AS-6-1. Algal Res 16:427–433. https://doi.org/ 10.1016/j.algal.2016.04.002 Zouboulis AI, Loukidou MX, Matis KA (2004) Biosorption of toxic metals from aqueous solutions by bacteria strains isolated from metal-polluted soils. Process Biochem 39(8):909–916. https:// doi.org/10.1016/S0032-9592(03)00200-0

Chapter 18

Biological Strategies for Heavy Metal Remediation Memory Tekere

Contents 18.1 18.2

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Microbial Strategies in Metal Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.1 Biosorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.2 Bioleaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.3 Bioaccumulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.4 Biotransformation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.2.5 Biomineralization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3 Microbes in Heavy Metal Pollution Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.1 Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.2 Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.3.3 Algae . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4 Phytoremediation of Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4.1 Plants for Metal Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.5 Engineered Organisms for Enhanced Metal Remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.6 Conclusion and Further Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Heavy metal contamination of the environment is a worldwide concern, and due to the toxic effects of metals, efforts are concerted on the remediation of the contamination. Sources of heavy metal pollution in the environment include mining, metal smelting, agriculture, waste disposal and atmospheric deposition. Physical, chemical, and biological methods are all explored in cleaning up the environment from metal pollution. Biological remediation of heavy metals involves the use of organisms such as various microorganisms and plants, and their derivatives or products are applied as a mild environmentally friendly method for decontaminating the environment. Biological strategies, unlike other methods of remediation, are unique in that biological strategies are environmentally friendly and acceptable, the diversity of organisms involved is wide and of diverse capabilities that have not yet M. Tekere (*) Environmental Science Department, University of South Africa, Johannesburg, South Africa e-mail: [email protected] © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature 393 Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4_18

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been exhaustively exploited and also amenable to genetic modification for accelerated bioremediation. Biological strategies for heavy metal remediation are applied to all the main environmental components: soil, air, and water. Bioremediation processes can be active (metabolic and energy dependent) or passive and can use dead or live organisms/biomass. Microorganisms can be applied directly to contaminated sites or used in designed bioreactors optimized to achieve the remediation goals. In phytoremediation, plants are grown on contaminated sites or in wetlands or their biomass used. This chapter gives an overview of biological remediation of heavy metals using plants and microorganisms. The strategies that are used by microorganisms in bioremediation include biosorption, biotransformation, bioaccumulation, bioleaching, and biomineralization. Further elaboration of bacteria, fungi, algae, and plants’ contribution to biological remediation of metals is given. Keywords Bioremediation · Phytoremediation · Microorganisms · Biosorption · Heavy metals · Pollution · Bacteria · Fungi · Plants · Algae

18.1

Introduction

Heavy metals in the environment arise from natural sources as well as anthropogenic activities and are acknowledged as major causes of concern for environmental pollution. Soil and aquatic environments receive a wide range of metal pollutants and of different concentrations resulting from human actions such as intentional uses, poor disposal of waste, accidental metal waste discharges, and inappropriate use (Tekere and Kamika 2016; Banerjee et al. 2018). Human activities (anthropogenic) through mining, agriculture, and other industrial activities result in most of the environmental pollution by heavy metals (Kavamura and Esposito 2010; Tekere and Kamika 2016). Compared to organic pollutants, heavy metals are non-biodegradable; they persist in the environment and can accumulate to levels that are not safe for humans, plants, animals, and other soil and aquatic living organisms. Non-biodegradability of heavy metals as inorganic compounds makes it difficult to clean up the environment of the pollution. Cadmium, arsenic, lead, cobalt, copper, mercury, nickel, selenium, and zinc are among toxic metals of concern frequently found in the environment (Kav amura and Esposito 2010). The reported order of toxicity, in order of the increasing toxicity, for these metals is as follows: lead < zinc < copper < cadmium and is influenced by prevailing abiotic and biotic factors (Ren et al. 2009; Kavamura and Esposito 2010). Living organisms inhabiting in and near metal-contaminated sites are exposed to the metals, and depending on the organism (excluding plants), the metals can reach target organisms through drinking, eating, breathing, and dermal contact (Forstner and Wittmann 2010; Gupta et al. 2016). Accumulated heavy metals in the human

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body, for example, result in multiple health effects such as lung, kidney, and DNA damage. Concerted effort has to be instituted in dealing with metal pollution. Strategies applicable in heavy metal remediation can be physical, chemical, and biological. Chemical and physical methods of heavy metal remediation include filtration, reverse osmosis, ion exchange, oxidation, reduction, precipitation, evaporation, and electrochemical treatment (Liu et al. 2018; Jacob et al. 2018). The drawback on physical and chemical methods however is that they are expensive and often not very efficient with metals at low concentrations or dissolved in wastewaters and result in generation of large amounts of secondary wastes (Ahluwalia and Goyal 2007; Dixit et al. 2015; Liu et al. 2018; Jacob et al. 2018). On the other hand, biological remediation of metals that relies on biological organisms (plants, algae, fungi, and bacteria) and their derivatives as the main driving agents can be used (Kulshreshtha et al. 2014; Eccles 1995; Gupta et al. 2016). The process where biological organisms are used in mediating the removal of pollution from the environment is called bioremediation. Microorganisms and plants that find applications in metal remediation are those which have developed mechanisms of coping with metal contaminants upon exposure. Microbial adaptation in heavy metal-polluted habitats includes detoxifying mechanisms such as biosorption, bioaccumulation, biotransformation, and biomineralization as shown in Fig. 18.1 (Dixit et al. 2015; Bansuri et al. 2018; Jacob et al. 2018). Bioremediation can be applied in situ, that is, on site of contamination, or ex situ, where the treatment of the contaminated media is done off the site of contamination. Ex situ bioremediation often requires movement of contaminated media and treatment offsite and has the advantage of treatment being done under optimized controlled environmental conditions (Liu et al. 2018).

Fig. 18.1 Simplified diagram of different ways of bacterial interaction with heavy metals. (Modified after Bansuri et al. 2018)

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Microbial Strategies in Metal Remediation

With increasing environmental pollution by metals from different sources, microorganisms have to thrive in contaminated habitats. Microorganisms have and continue to develop ingenious mechanisms of dealing with heavy metals in their environmental habitats. The mechanisms include enzyme secretion, cell morphological changes, metal efflux, metabolite secretion, resistance to high metal concentration, metal accumulation within intracellular systems, chelation, biomethylation, and transformation (Kavamura and Espsito 2010; Yao et al. 2012; Dixit et al. 2015; Jacob et al. 2018). In metal remediation, these mechanisms are used in developing the necessary remediation strategies (Kavamura and Espsito 2010; Gupta et al. 2016; Mashangwa et al. 2017; Ojuederie and Babalola 2017). Microbial remediation can involve natural attenuation where activities of naturally occurring organisms are harnessed without human intervention, bio-stimulation where activities of naturally occurring organisms are enhanced to promote remediation, e.g., through the addition of metabolic substrates, aeration, or bioaugumentation where microorganisms are brought onto the contaminated site to aid in the remediation. Live or non-viable microorganisms are exploited in bioremediation, and the microorganism can be natural or genetically engineered to enhance their remediation competences (Gupta et al. 2016). The proceeding sections describe the microbial processes of biosorption, biotransformation, bioaccumulation, bioleaching, and biomineralization that are utilized in remediation of heavy metals. Figure 18.1 shows the different ways by which microorganisms interact with metals.

18.2.1 Biosorption Biosorption capitalizes on the natural affinity of biological compounds to heavy metals (Aly et al. 2018). The biosorption mechanisms for metal remediation can involve intracellular precipitation, adsorption onto cell surfaces (physical adsorption), complexation, ion exchange, and precipitation (Ojuederie and Babalola 2017). Biosorption is a non-enzymatic process, and the sorption of metals onto cell surfaces relies on structure of the cell walls and various physical and chemical factors such as metal combinations, ionic strength, and temperature (Bansuri et al. 2018; Aly et al. 2018). Different microbial species (fungi, bacteria, and algae) have different abilities in metal absorption. A variety of microbial biomass of varying capacities is used in biosorption especially for dilute metallic ions in contaminated water (Gupta et al. 2016; Jacob et al. 2018). Additionally, inactive non-living microbial or plant-based biomass when applied as biosorbents in the removal of metals from aqueous media help circumvent the need to maintain conditions that are often required for optimal growth of the organisms (Ahluwalia and Goyal 2007; Mashangwa et al. 2017). Biosorption is considered as an innovative technology for heavy metals that is

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cost-effective and eco-friendly in nature and more preferable to bioaccumulation (Banerjee et al. 2018). Low operational cost, simple, very effective, and environmentally friendly are also attributes of the biosorption strategy. The absorption of metals by different microbial biosorbents has been found to be selective depending on metal affinities exhibited, ionic strength, metal mixtures, pH, and temperature (Aly et al. 2018). Constituents such as proteins, carbohydrates, and functional carboxyl, amine and hydroxyl groups, which are part of microbial cells, function in metal absorptions (Ren et al. 2009; Mashangwa et al. 2017). Filamentous fungi are composed of cell wall components such as chitin, chitosan, melanins, glycoproteins, lipids, and D-galactosamine polymers which offer binding sites for metals (Vendruscolo et al. 2017). Different fungi, algae, and bacteria can act as biological-based heavy metal ion exchangers in absorption (Ahluwalia and Goyal), and metal desorption can be easily done to recover the metals. Among the disadvantages to the use of biosorption is the saturation of binding sites that limits the use of the biomass in further absorption unless desorption is done, and if non-living biomass is used, the heavy metals often remain in their valence without any potential for alteration to less toxic or soluble forms (Ahluwalia and Goyal 2007). Where dead biomass is used, fouling occurs as the biomass degrades over time and available binding sites in turn become reduced.

18.2.2 Bioleaching Bioleaching utilizes living organisms for metal extraction from their ores in environmental samples (Yin et al. 2018; Okoh et al. 2018). Not only is bioleaching applicable to bioremediation but also to biomining of low-grade mineral or waste mineral resources (Mishra et al. 2005; Pathak et al. 2017). Several minerals can be recovered from low-grade metallic ores or waste streams including sewage, metal industry sludge, mine dumps, and electronic wastes, with examples including copper, iron, lead, nickel, gold, cobalt, silver, and zinc (Okoh et al. 2018). Bioleaching in bioremediation is applicable to metal recovery from sewage, coal fly ash, and low-grade ore mine dumps, thus detoxifying the environment (Yin et al. 2018; Gu et al. 2018). Heterotrophic microorganisms are able to produce organic acids and other metabolites that are involved in the indirect processes where metals are dissolved from minerals by displacement from the source or by forming soluble metal complexes and chelates (gluconic, citric, oxalic acid) (Ren et al. 2009; Okoh et al. 2018). Bacteria including Acidothiobacillus, Sulfobacillus, Ferrimicrobium, and Leptospirillum and fungi Fusarium oxysporum, Penicillium glabrum, Aspergillus niger, and Penicillium simplicissimum find applications in bioleaching (Xu and Ting 2009; Yin et al. 2018; Gu et al. 2018). Aspergillus niger due to its high organic acid productivity is extensively explored in bioleaching. Bioleaching as a remediation strategy offers advantages of being low cost, environmentally friendly, low energy, and easy to operate (Ren et al. 2009). High removal efficiency is achievable even in very dilute effluents as the microorganisms

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can efficiently target the specific metals (Yin et al. 2018). In bioleaching, there is reduction of emissions that would normally arise if physical and chemical processes are applied thus the strategy’s environment-friendliness, there is potential for recovery of valuable metals in low grade ores, and there is simplicity in process design and operation making it cheap (Yin et al. 2018; Gu et al. 2018). The process is thus used in biomining as well as in decontamination of water and soil environments (Okoh et al. 2018). Disadvantages associated with bioleaching are that as a biological process, it is slow when compared to other applicable physical/chemical processes, and there is potential for generation of toxic acid mine drainage into the environment (Yang et al. 2016).

18.2.3 Bioaccumulation Bioaccumulation takes the form of active (energy dependent) absorption of metals into the cells using importer complexes for translocating metals into the cells through the lipid bilayer (Bansuri et al. 2018; Diep et al. 2018). The process is alternatively described as metabolism-dependent biosorption (Igiri et al. 2018). Concentration of heavy metal pollutant becomes high in the cell cytoplasm as compared to the external environment as the metals are sequestered and localized into the cells. Metal binding proteins (metallothionens) are involved in binding and complexing the metals in cells. Further cationic groups on cell surfaces such as teichoic acids, peptidoglycan phosphate, and carboxyl groups offer binding sites for heavy metals (Aly et al. 2018). Biosynthesis of metal binding proteins and peptides under heavy metal stress through cell signaling helps in heavy metal accumulation. Metalloregulatory proteins function to regulate the accumulation of metals in the cell (Banerjee et al. 2018). Several bacteria have been used for intracellular metal bioaccumulation including Pseudomonas putida, Rhizobium leguminosarum, and Pseudomonas syringae (Igiri et al. 2018). Figure 18.2 gives an illustration of how bioaccumulation of heavy metals takes place in Gram-negative bacteria.

18.2.4 Biotransformation Heavy metal biotransformation involves changing metal species from one form to another including inorganic and organic forms through processes like physical adsorption, membrane transport, ion exchange complexation, and biosorption (Banerjee et al. 2018). Heavy metal transformation has great potential to reduce toxic effects of environmental pollution by the metals. High valence metals such as Cr(VI), U(VI), Hg(II), and Se(VI) are reduced to lower oxidation states through action microorganisms. Examples of biotransformative metal reductive detoxification include reduction of mercury from +2 to 0 state, chromium +6 to +3 state, and selenium from +6 to 0 state (Banerjee et al. 2018). Cr(VI) is said to be hundred times

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Bioaccumulation

Biosorption

Periplasm

Outer Lipid Membrane

Porins

Secondary Carriers

Primary Active Transporters

H+

Inner Lipid Membrane

Channels

ADP + Pi

H+

Cytoplasm

ATP + H2O

P P PP P P Phytochelatins

Metallothioneins

Polyphosphates

Storage system

Fig. 18.2 Bioaccumulation of metals in Gram-negative bacteria. Biosorption, red arrows, metal binding entity; metal, black circles; purple arrows, transportation of heavy metals through lipid membranes; blue, brown, red, yellow, import-storage protein machinery. (Modified after Deip et al. 2018)

more toxic when compared to Cr(III) (Vendruscolo et al. 2017). On the downside, however, conversion of metals to methylated organic forms, for example, methylmercury, makes them more volatile and toxic (Gupta and Nirwan 2015). A lot of microorganisms have been described for their ability for metal biotransformation including the fungi Aspergillus, Verticillium, Hymenoscyphus, Cephalosporium, Pleurotus, Phlebia, and Botryosphaeria (Banerjee et al. 2018; Sutjaritvorakul et al. 2015). An analysis of mercury by X-ray diffraction analysis in the biomass of a heavy metal-tolerant Alcaligenes strain showed that sequestered mercuric ions were biotransformed into monovalent mercury (Hg2Cl2), a form of mercury which is not bioavailable (Gupta and Nirwan 2015). The mercurous chloride (calomel) can be easily collected out, being insoluble in water, and utilized in electrochemical industry (Gupta and Nirwan 2015).

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18.2.5 Biomineralization Microorganisms can directly or indirectly mediate the formation of mineral substrates through biomineralization during their metabolic activity. In biomineralization, specific forms of inorganic substrates are synthesized using the substrate molecule relevant to performing various functions in the biological system (Banerjee et al. 2018). Microorganisms immobilize heavy metals in mineral crystal lattices or convert heavy metals into insoluble minerals through metabolism, thereby reducing their mobility and bioavailability. This allows for efficient sequestration of the heavy metals within relatively stable solid phases (Li et al. 2013). Oxides, phosphates, sulfides, and oxalates are some of the most common biominerals precipitated by microbes and have different unique chemical properties such as high capacities of metal capacities and redox catalysis (Gadd 2010). Bacteria producing urease, for example, promote the formation of carbonate by producing urease to hydrolyze urea into NH4+ and CO32 and NH4+ releases NH3, which increases the average pH until the NH4+/NH3 and HCO3 /CO3 ratios achieve equilibrium (Zhao et al. 2019). Under alkaline conditions, the metal ions combine with CO32 to form a precipitation that could convert the soluble heavy metals into insoluble carbonate mineral crystals. In a study by Zhao et al. (2019), a urease (urea amide hydrolase, EC3.5.1.5)-producing bacteria could mineralize 2 mM Cd by +72%. Urease-producing bacteria therefore promote mineral precipitation, thus removing and inactivating metals in soil and waste (Li et al. 2013).

18.3

Microbes in Heavy Metal Pollution Remediation

Microbial remediation utilizes living or non-living algae, fungi, and bacteria. Several bacterial, algal, and fungal species existing in heavy metal-contaminated soils have been described (Kavamura and Esposito 2010; Gupta et al. 2016; Igiri et al. 2018 ; Timková et al. 2018). Microbial bioremediation as described in previous sections takes the form of bioaccumulation, biomineralization, biosorption, bioaccumulation, and biotransformation. The remediation can be carried out in situ or ex situ in natural settings or engineered bioreactors. Engineered bioreactors (e.g., stirred tanks, biofilters) providing for optimum conditions for microbial growth and biodegradation are often applied in ex situ bioremediation to achieve the remediation goals (Chikere et al. 2012). Several microorganisms from the different microbial groups as shown in Tables 18.1, 18.2, and 18.3 have been evaluated and demonstrated heavy metal remediation capabilities.

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Table 18.1 Examples of bacteria that have been evaluated for metal remediation Bacteria Bacillus licheniformis Bacillus coagulans, Bacillus megaterium Enterobacter sp. JI Pseudomonas aeruginosa

Metal Cr(VI) Cd Cr(VI) Cd Cr

Bacillus firmus

Pb, Cu, Zn

Bacillus cereus Desulfovibrio desulfuricans

Hg Cu, Ni, Cr

Enterobacter cloacae Streptomyces noursei Cupriavidus metallidurans Alcaligenes sp. Micrococcus luteus

Pb Cu, Zn, Co, Ag Cu, Cr Pb Cu, Pb

Bacillus cereus Bacillus subtilis Pseudomonas veronii Micrococcus sp. Geobacillus thermantaracticus, Anoxybacillus amylolyticus

Cr(VI), Cu Cr(VI) Cd, Zn, Cu Ni Cd, Cu, Co, Mn

Removal capacity 62 mg/g 142.7 mg/g 39.9 mg/g 46.2 mg/g 99.6% 467, 381, 418 mg/g 104.1 mg/g 98.2, 90.1, 99.8 mg/g 2.3 mg/g 9, 1.6, 1,2, 38 mg/g – 56.8 mg/g 408, 1965 mg/ g 82–100% 99.6% 49.8% 55% –

References Zouboulis et al. (2004) Srinath et al. (2002) Lu et al. (2006) Benazir et al. (2010) Ahluwalia and Goyal (2007) Sinha et al. (2012) Kim et al. (2015) Kang et al. (2015) Ahluwalia and Goyal (2007) Fan et al. (2014) Jin et al. (2017) Jacob et al. (2018) Nayak et al. (2018) Igiri et al. (2018) Vullo et al. (2008) Igiri et al. (2018) Ozdemir et al. (2013)

18.3.1 Bacteria Heavy metal bioremediation capacity of bacteria arises from basic inherent coping mechanisms that prompt the cells to mitigate metal toxicity (Jacob et al. 2018). Bacteria take up high amounts of metal ions accumulating them from the environment into their cells (Ahluwalia and Goyal 2007). The anionic nature of microbial cell surfaces through electrostatic force is mainly responsible for binding metals. The functional groups that include hydroxyl, carboxyl, sulfonate, amide, and phosphate are mainly involved in the uptake of metal ions by different bacterial species. The Gram-positive bacteria are more efficient in trapping metal ions than Gram-negative due to their thick walls and component teichoic, acids, and peptidoglycan that are not present in Gram-negative bacteria (Gupta et al. 2016). As indicated in Table 18.1, a numerous number of bacteria have been evaluated for metal remediation. Living and non-living bacteria with high metal binding affinities find application in bioremediation as biosorbents. Biomass from spent fermentation processes such as Pseudomonas and Bacillus can be used (Ghosh

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Table 18.2 Examples of fungi and yeasts studied for bioremediation Fungi and yeast Ganoderma lucidum Phanerochaete chrysosporium, Aspergillus awamori Mucor rouxii Pleurotus platypus Aspergillus niger, Lepiota hystrix Rhodotorula mucilaginosa Rhizopus oryzae Aspergillus brasiliensis, Penicillium citrinum Aspergillus niger Aspergillus niger Aspergillus terreus, Trichoderma longibrachiatum Aspergillus niger NCIM548 Lentinula edodes Pleurotus sapidus Aureobasidium pullulans CCY 27-1-111 Saccharomyces cerevisiae

Metals Cr, Cu, Pb Pb, Cd, Cr, Ni Ni Ag Cr Pb, Cu Ag Cu Cu, Mn, Zn Cr (IV) Pb Pb, Cd, Cr Ni, Co

Metal removal –

References Jacob et al. (2018)

–

Joshi et al. (2011)

0.36 mg/g 46.7 mg/g 5.1–6.6 mg/g 3.89, 8.50 mg/g 26 mg/g 34 mg/g –

Jacob et al. (2018) Das et al. (2010) Jacob et al. (2018) Kariuki et al. (2017) Gomes et al. (2002) Fu et al. (2012) Pereira et al. (2014)

11.792 mg/g 5.3–34.4 mg/g 59.67, 16.25, 0.55, 0.55 mg/g 70.4%, 66.9%

Mondal et al. (2017) Jacob et al. (2018) Joshi et al. (2011)

Cd Cd, Hg Cd

96.3% 127, 287 mg/g –

Mn

22.5 mg/g

Biswas and Bhattacharjee (2014) Wang et al. (2017) Jacob et al. (2018) Breierová et al. (2002) Fadel et al. (2017) and Gupta et al. (2016)

and Paul 2016). Several bacteria were identified using DGGE for their role in metal remediation (Kavamura and Esposito 2010). Igiri et al. (2018) provide a comprehensive list of bacteria and other microorganisms evaluated for heavy metal remediation. Table 18.1 provides some examples of bacteria, the different metals, and efficiency of removal under the studied conditions.

18.3.2 Fungi Several fungi can tolerate and detoxify heavy metal contamination (Igiri et al. 2018; Jacob et al. 2018; Srivastava and Thakur 2006). Fungi have a greater cell to surface ratio and are good sorption material in bio-absorption (Jacob et al. 2018). In soils, fungal hyphae penetrate deep into soil aggregates where they chelate and absorb the heavy metals (Aly et al. 2018). Cell walls of fungi are composed of chitin among other polysaccharides, proteins, lipids, polyphosphates, and inorganic ions. Species such as Pleurotus ostreatus, Irpex lacteus, Agaricus bisporus, Trametes versicolor, and Bjakendera adusta are some of the superior fungal biosorbents for metals

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Table 18.3 Metal removal biopotential of different algae biomass. (Adapted for Jacob et al. 2018) Algae strain Spirogyra sp. Aphanothece halophytica Spirogyra sp. Chlorella sorokiniana Dunaliella sp. Sargassum wightii Sargassum sp. Palmaria palmata Spirulina maxima Spirogyra hyalina Micrasterias denticulata Cystoseira barbata Spirogyra neglecta Cladophora sp. Cladophora hutchinsiae Chlorella vulgaris Scenedesmus obliquus Chara aculeolata Nitella opaca Eucheuma denticulatum Ulva lactuca

Metal Cr Zn Zn Cd Cr Ni Ni Pb Pb Cd, Hg, Pb, As Cd Cd, Ni, Pb Cu Pb Se U Zn Pb, Cd, Zn Pb, Cd, Zn Pd, Cu, Fe, Zn Cu2+, Pb2+, Cd2+, Cr2+

Removal capacity (mg/g) 133.30 133 – 13.33 58.3 37.2 181 15.17 – – – 37.6, 78.7, 196.7 30.17 45.4 74.9 14.3 429.6 105.3, 23.0, 15.2 104.2, 20.5, 13.4 81.97, 66.23, 51.02, 43.48 64.5, 62.5, 60.9, 68.9

(Jacob et al. 2018). Table 18.2 provides a list of some of the fungi studied for metal remediation. Production of high amounts of fungal biomass is achievable using simple fermentation techniques and also can be obtained from spent fermentation production processed, e.g., Saccharomyces cerevisiae from breweries. The fungus Coprinopsis atramentaria was shown to remove 94.7% of Pb2+ at a concentration of 800 mgL 1 (Igiri et al. 2018). Dead fungal biomass was also shown to useful in converting the toxic Cr(VI) to Cr(III) which is less toxic (Park et al. 2005; Igiri et al. 2018). Yeast strains that include S. cerevisiae, Yarrowia lipolytica, Hansenula polymorpha, Pichia guilliermondii, Rhodotorula pilimanae, and Rhodotorula mucilage bio-convert Cr(VI) to Cr(III) (Igiri et al. 2018). In 2015, Fazli et al. described several fungi isolates that had high Cd tolerance and remediation potential, with minimum inhibitory concentrations of 1000–4000 mgL 1.

18.3.3 Algae Advantage of the autotrophic algae as biosorbents is that they have high biomass productivity at low nutrient requirements (Jacob et al. 2018). The cell walls of algae provide for metal binding sites through their amino and carboxyl groups as

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well as the nitrogen and oxygen moieties (Jacob et al. 2018). Red algae can have 7–50% protein as constituents of their cell wall. The algae Chlorella vulgaris, Synechococcus sp., Cladophora crispata, Anabaena sp., Eklonia sp., Ulothrix zonata, Caulerpa lentillifera, Chlorella vulgaris, and Microcystis aeruginosa have been reported capable of biosorption of metals (Ahluwalia and Goyal 2007; Imani et al. 2011). Removal of Cd and Hg from industrial wastewater was at 90–95% with the brown algae tested (Lodeiro et al. 2005). pH affected the uptake of chromium by Anabaena sp. and Synechococcus sp. with uptake increasing on pH decrease (Ahluwalia and Goyal 2005). Chemical moieties instrumental on the metal binding sites of algae include hydroxyl, carboxyl, phosphate, and amide groups (Igiri et al. 2018). Metal removal biopotential of different biomass of some of the studied algae is shown in Table 18.3.

18.4

Phytoremediation of Heavy Metals

Phytoremediation is defined as the use of plants to mitigate environmental pollution. In contrast to microbial bioremediation, in phytoremediation plants are responsible for the transformation and/or immobilization of the metal contaminants through physical and chemical processes. Heavy metals cause necrosis and chlorosis in plants, and plants that survive metal contamination have to be tolerant and adapted to grow in heavy metal-pollutant environments. The ability of some plant species to tolerate extreme metal contamination has been evaluated, resulting in increasing interest for using phytoremediation as a cost-effective alternative method to conventional ones (Prasad and de Oliveira Freitas 2003; Banerjee et al. 2018; Jacob et al. 2018). Plants which can tolerate heavy metals, with high bioaccumulation factor, and possessing short life cycle are found suitable for the remediation of contaminated soil (Prasad and de Oliveira Freitas 2003). Plants used for phytoremediation on contaminated land in wetlands, for example, are beneficial in that they provide aesthetic quality of the environment, encourage microbial population establishment, help to stabilize the effluent/substrate and encourage a more even flow of effluent in the wetland system (Tekere and Kamika 2016; Herniwanti et al. 2013). Though phytoremediation seems cheaper than other conventional methods, the method is complicated by the need to plant and grow hyperaccumulating plants in the metalpolluted areas requiring highly technical strategies by expert project designers of relevant field experience choosing the proper plants for particular metals and area. Phytoremediation is classified based on processes occurring in plants as follows: phytosequestration, rhizofiltration, phytoextraction, phytodegradation, and phytovolatilization. All or a combination of these processes can occur at the same time depending on the type of metals and media affected (soil or water), as well as clean-up goals that include stabilization or sequestration, reduction, assimilation, detoxification, and degradation. The different phytoremediation techniques are defined in Table 18.4.

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Table 18.4 Type of phytoremediation mechanisms Mechanism Phytosequestration

Rhizofiltration Rhizodegradation Phytoextraction Phytodegradation Phytovolatilization Phytostabilization Phytostimulation

Description The ability of plants to take in certain metals in the rhizosphere through exudation of phytochemicals and on the root through transport proteins and cellular processes. This reduces movement of the contaminant and migration to soil, water, and air The plants through the rhizomes absorb, concentrate, and/or precipitate contaminants in the aqueous system Contaminants in the rhizosphere are degraded via microbial activities and exuded phytochemicals The ability of plants to take up contaminants from both soil and water into the plant and is stored or translocated to the shoot Uptake and breakdown of pollutants via internal enzymatic activity and photosynthetic oxidation/reduction Some ions of elements like Hg, Se, and Ar are absorbed by the root and converted into less toxic forms and released The inorganic compound is incorporated to the lignin or to soil humus The growth activities of the roots promote development of rhizosphere microorganisms that can degrade the contaminant, using exudates as carbon source

18.4.1 Plants for Metal Remediation During phytoremediation, plants serve as a medium either alone or in association with microorganisms in the removal of heavy metal pollution. The plant species used in the remediation of heavy metals differ in the uptake and accumulation of different trace elements in their tissues. Plants that have the ability to take up very high concentrations of metals into their system (hyperaccumulators) are preferentially used. Such capabilities may occur due to some genetic natural selection of metals which can also be variable among populations and within organisms of the same population (Herniwanti et al. 2013). A number of studies have been done that examine the abilities of some plant species to accumulate metals in their biomass (Herniwanti et al. 2013; Mkandawire et al. 2004). Some examples of plants that are used in metal phytoremediation are given in Table 18.5. In a study that evaluated some local water plant characteristics for phytoremediation of acid mine drainage, Eleocharis dulcis was found to reduce iron best, while Pistia stratiotes was effective at reducing manganese levels (Herniwanti et al. 2013). Species of the plants Brassica juncea, Arabidopsis thaliana, and Chara canescens were found capable of taking up forms of AS, Hg, and Se and transfer them to gaseous forms inside the plants before release into the atmosphere (Khalid et al. 2017). Volatilization can be enhanced by genetic engineering of the plants. It is reported that to enhance Hg volatilization, Hg reductase genes have been inserted into plants from bacteria and these transgenic plants can volatilize Hg at 100–1000 times than the wild-type plants (Khalid et al. 2017; Rugh et al. 1996). Plant families that dominate among metal hyperaccumulators include

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Table 18.5 Some examples of plants applicable for metal remediation Plants material Carica papaya wood Berkheya coddii Ni Sawdust(Acacia arabica) Euphorbia cheiradenia Ricinus communis L. Water hyacinth Oryza sativa husk Helianthus annuus Alyssum bertolonii Alyssum murale Ni Arabidopsis halleri Minuartia verna Sedum alfredii Erica australis Pteris vittata Viola boashanensis Thlaspi caerulescens Zea mays L. Lemna gibba L.

Metal Hg (II)

References Gupta et al. (2016)

Ni Pb(II), Hg (II), Cr (VI), Cu(II) Cu, Fe, Pb, Zn

Ojuederie and Babalola (2017) Gupta et al. (2016) Nematian and Kazemeini (2013)

Hg (II)

Gupta et al. (2016)

Pb (II), Cu (II), Co (II), Zn (II) Pb(II) Pb, Cd, Zn, Cu Ni Ni Zn, Cd

Gupta et al. 2016) and Kumar et al. (2017)

Zn, Cd, Pb Pb Fe and Al Hg, As Pb, Zn, Cd Zn, Cd

Gupta et al. (2016) and Zulkali et al. (2006) Angelova et al. (2016) Ojuederie and Babalola (2017) Broadhurst and Chaney (2016) Zhang et al. (2017) and Claire-Lise and Nathalie (2012) Ojuederie and Babalola (2017) Chen et al. (2013) Abreu et al. (2008) Xie et al. (2009) and Su et al. (2008) Zhuang et al. (2005) Assunção et al. (2008)

Pb and Zn U, As

Li et al. (2013) Mkandawire et al. (2004)

Brassicaceae, Asteraceae, Violaceae, Fabaceae, Caryophyllaceae, Euphorbiaceae, Lamiaceae, Flacourtiaceae, and Poaceae. Heavy metals are able to induce a range of morphological, biochemical, and physiological disorders in plants and often plants respond to these effects (Shahid et al. 2016). Mechanisms such as stress enzyme production help plants resist adverse damages. Hypertolerance is therefore key in making hyperaccumulation possible. Plants that include thatching grass (Hyparrhenia cuspidate), hippo grass (Vossia cuspidata), and wiregrass (Gentiana pennellianna) are reported as excellent candidates for phytostabilization of Zn-, Cu-, Cr-, and Pb-contaminated soils (Liu et al. 2018; Radziemska et al. 2017). Though not best suited to extremely contaminated sites where plants can barely grow and survive, phytostabilization helps reduce ecological risks from such contaminated sites (Liu et al. 2018).

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407

Engineered Organisms for Enhanced Metal Remediation

As one of the advantages for using bioremediation, biological organisms are amenable to genetic engineering to enhance remediation competences. Heavy metal exposure in plants and microorganisms induces gene expression changes with genes involved in uptake, transportation, and sequestration being upregulated or downregulated (Diep et al. 2018). The genetic capacity of microorganisms can be engineered to improve on bioremediation competences such as increasing the rate of bioleaching, metal tolerances, and over- or co-expression of metal transporter with metal-chelating/binding proteins (Okoh et al. 2018; Kavamura and Esposito 2010; Khalid et al. 2017). Also the metal binding specificity of the proteins can be altered so as to allow a wide range of metals to be bound and transported for bioaccumulation and transformation in the cells. Recombinant metal-binding proteins and peptides that enhance specificity and capacity by microbial biosorbents have been expressed in the microorganisms (Diep et al. 2018). Studies on Bacillus subtilis identifying genes inducible by heavy metal exposure show that many of metal stress affected genes were controlled by metalloregulatory proteins Fur, MntR, PerR, ArsR, and Cu (Moore et al. 2005; Kavamura and Esposito 2010; Diep et al. 2018). Genetically engineered strains of P. putida and E. coli were reported for effective mercury removal from sediments (Chen and Wilson 1997; Ojuederie and Babalola 2017). The use of genomics, proteomics, and metabolomics goes a long way to help with understanding the players in decontamination, the metabolic pathways, and mechanisms in metal resistance, among others. However despite the advantages that genetic engineering may offer to heavy metal remediation, scientists are weary of possible horizontal gene transfer between engineered and indigenous species. Mechanisms of monitoring and controlling the genetically engineered microorganisms beyond their application as well as regulation are reported (Azad et al. 2014; Jan et al. 2014).

18.6

Conclusion and Further Considerations

Heavy metals as inorganic pollutants are not easily amenable to remediation. While physical and chemical methods for heavy metal remediation are available, there is however continued search for more cost-effective and environmentally friendly methods for clean-up of metal pollution from the environment. Biological methods utilizing microorganisms and plants fill in this gap and can be fully optimized for field scale bioremediation applications and with higher public acceptability. Not only are microorganisms applicable in bioremediation but also in recovery of heavy metals from different industrial wastes and low-grade ores. Several plants and microorganisms have inherent metal-resistant mechanisms that give them selective advantages of thriving in increasingly metal burdened environments. These metal-

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resistant microorganisms and plants when applied in heavy metal remediation result in efficient environmentally friendly detoxification. While many microbial and plant species are applicable in different technologies for bioremediation, the competitiveness of biological processes is an ongoing area of research and development. The variability and complexity of heavy metal pollution in different environment make the requirement for site-specific methods imperative. Also in as much as some in roads have been made in development and application of bioremediation techniques, most of the microorganisms and plants remain untapped for the bioremediation potential and also potential of those described not yet fully optimized. Optimized processes that make use of immobilized biomass and engineered biomass can be further developed to facilitate economic competitiveness of this green technology, and merging genetic engineering and ecological understanding is prerequisite. For biosorption, obtaining reliable and inexpensive biomass applicable to mixed metal ions and of prolonged capacity is identified as one of the inhibitory factors. Overall the ultimate goal in further biotechnology research on microbial remediation of heavy metals is to increase metal tolerance and the biosorption, biotransformation, bioleaching, and bioaccumulation as is necessary for economical bioremediation.

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Index

A Abadi, B., 22–57 Abadulla, E., 307 Abbas, N., 305 Abdulredha, S., 307 Abisamra, N.S., 51 Abreu, M.M., 406 Adishesh, S.C., 254–262 Adnan, L.A., 307 Adsorption, 8, 88, 138, 182, 218, 226, 259, 270, 278, 290, 328, 359, 378, 396 Advantages, 76, 94, 121, 147, 164, 171, 190, 193, 214, 215, 226, 233, 242, 268, 282, 290, 328, 329, 331, 333, 335, 336, 350, 351, 355, 359, 360, 363, 379, 395, 397, 403, 407 Agenda, 29–56 Ahluwalia, S.S., 301, 302, 304, 401 Ahmad, S., 382 Ahmadi, E., 260 Ahmed, M.K., 295 Ahmed, S.G.K.A., 303 Aikat, K., 302, 303 Ajala, S.O., 342–363 Ajaz, M., 307 Akhter, K., 382 Akinbile, C., 190 Akintunde, A.M., 342–363 Aksu, Z., 382, 383 Akthar, N., 302 Alam, M.Z., 382 Algae, vi, 3, 88, 94, 139, 162, 163, 166, 170, 171, 183, 184, 186, 204–207, 221, 230,

232–234, 246, 301, 329, 330, 376, 380–384, 395–397, 400, 403, 404 Ali, S.A.M., 302 Altaf, M., 182–206 Ammari, T.G., 382 Amrane, A., 204 Anaerobic digestion, 68–70, 72 Anderson, C., 358 Aneesh, E.M., 268–274 Angassa, K., 333 Angelova, V.R., 406 Anjaneya, O., 303 Anjum, M.S., 204 Anoopkumar, A.N., 268–274 Anwar, F., 305 Applications, vi, 4, 6, 70, 87, 88, 119, 123, 127, 128, 140–142, 157, 162, 170, 186, 226–228, 230–232, 234, 235, 240, 242, 246, 255, 273, 274, 284, 294, 328, 329, 331, 336, 338, 343, 350, 351, 359–361, 375, 376, 378, 395, 397, 401, 407, 408 Aquatic environments, 117–129, 166, 167, 172, 183, 298, 394 Argos, M., 346 Arica, M.Y., 149, 151 Aromatic compounds, 124, 125, 142, 200, 337 Ashraf, R.S., 182–206 Ashraf, S.S., 138–157 Assunção, A.G., 406 Aydin, S., 77 Ayed, L., 305 Azeem, F., 290–312 Aziz, H.A., 295

415 © The Editor(s) (if applicable) and The Author(s), under exclusive license to Springer Nature Switzerland AG 2020 Inamuddin et al. (eds.), Methods for Bioremediation of Water and Wastewater Pollution, Environmental Chemistry for a Sustainable World 51, https://doi.org/10.1007/978-3-030-48985-4

416 Azo dyes, 90, 91, 152, 157, 189, 198, 202, 204, 290–292, 294, 296, 298–302, 305, 307, 309–331

B Babalola, O., 406 Bacteria, 2, 66, 87, 118, 139, 163, 183, 219, 230, 255, 271, 278, 291, 380, 395 Bacterial metabolites, 85–108, 120 Bacterial swarm, 119 Bagewadi, K.Z., 150 Baker, A., 356, 357 Baker, A.J., 356–358 Baldwin, P.R., 356 Banerjee, D., 240, 382 Bani, A., 356 Barapatre, A., 307 Barquilha, C.E., 384 Bayer, A.S., 21 Bayramoglu, G., 149, 151 Bazrafshan, E., 382 Benazir, J.F., 401 Benghazi, L., 308 Benoit, J., 329 Bernal, M., 356 Bestawy, E.E., 200 Betiku, E., 342–363 Beveridge, T.J., 383 Bharagava, R.N., 200 Bhatt, N., 311 Bhattacharjee, K., 402 Bhattacharya, P., 344 Bheemaraddi, M.C., 304 Bilal, M., 138–143, 157, 299 Binod, P., 268–274 Bio-sorption, 2, 128, 186, 214, 226, 283, 374, 395 Biochemical methods, 181–206 Biodegradable, 3, 98, 106, 279, 342 Biofilm, vi, 4–8, 76, 96–98, 105, 107, 118–129, 234, 235, 238, 241, 260 Biological material, 94, 282 Bioremediation, v, vi, vii, 2–4, 19, 22, 88, 95, 97, 98, 104, 106–108, 118–129, 138–157, 171, 183, 185, 189, 192, 193, 195, 196, 204–206, 226, 232, 233, 235, 241, 254–262, 278–286, 330, 348, 351, 352, 400–404, 407, 408, 395297 Biosorbents, vi, 2, 4, 8, 10, 11, 15, 16, 22, 95, 97, 101, 106, 218, 220, 222, 229–233, 242, 245, 282, 374, 379, 381, 383, 384, 396, 397, 401–403, 407

Index Biosurfactant, 6, 96, 123, 125, 128 Bisphenols, vi, 253–262 Biswas, S., 402 Blaszczyk, U., 346 Bølling, A.K., 256 Booker, S.M., 344 Boskabady, M., 346 Braeckevelt, M., 334 Breierová, E., 402 Brierley, J., 381 Broadhurst, C.L., 406 Brooks, R., 356–358 Brown, S., 357 Bryant, B., 42 Buitron, G., 63 Bulla, L.M.C., 308 Butcher, D.J., 356 Buthelezi, C.-S., 200 Buthelezi, S., 200

C Cabral, J.P., 383 Cadmium, 2, 28, 89, 124, 148, 172, 191, 228, 254, 268, 279, 334, 374, 394 Camacho-Chab, J., 382 Carrier-free immobilization, 154–157 Casieri, L., 307 Celia, M.P., 303 Chan, F.C., 303 Chandra, R., 200 Chaney, R.L., 356, 406 Chang, J.S., 304 Chatterjee, S., 200 Chatterjee, S.K., 382 Chaudhuri, S., 308 Cheah, C., 226–247 Chen, B., 406 Chen, C.-H., 201 Chen, F.P., 256 Chen, P., 344 Chen, S.H., 308 Chen, Z., 63, 73 Cheng, P., 381, 383 Cheng, Y.-Y., 346 Chojnacka, K., 238 Chon, H.T., 382 Chromium, vi, 2, 4, 19, 21, 28, 88, 90, 101, 106, 125, 126, 191, 194, 195, 200, 228, 229, 254, 269, 279, 282–284, 286, 296, 298, 334, 374, 404 Chu, K.H., 230, 381 Chung, J.-Y., 344

Index Claire-Lise, M., 406 Cleanup, 106, 119, 123–125, 270, 274, 333, 347–352, 360–363 Codd, G.A., 167 Comoglu, B., 76 Constructed wetland, vii, 187, 329, 331–334, 336–338 Contaminants, 2, 30, 107, 123, 139, 162, 182, 214, 229, 254, 269, 278, 328, 342, 384, 395 Contamination, v, vi, vii, 28, 30, 31, 50, 56, 62, 127, 138, 170, 194, 222, 227, 254, 255, 269–271, 273, 274, 278, 282, 290, 348, 349, 353, 361, 362, 395, 402, 404 Costa, A.C.A.D., 230, 232 Cunningham, S.D., 330 Cyanobacteria, 88, 105, 127, 161–174, 184, 230, 232

D Dan, T.V., 356 Daoud, N., 384 Das, A., 311 Das, D., 402 Das, K., 346 Das, N., 204 Datta, S., 86–108 Dauta, A., 166 Dawkar, V.V., 302, 303 Day, J.A., 172 De Philippis, R., 230 Deivasigamani, C., 204 Dellisanti, F., 351 Deng, L., 200 Desorption, 102, 154, 222, 246, 271, 358, 379, 397 Devassy, E., 268–274 Dhir, B., 172 Ding, Z., 190 Dionisio Pires, L.M., 166 Dirbaz, M., 382 Disadvantages, 214–217, 221, 280, 294, 336, 398 Djelal, H., 204 Do, M.T., 71 Donmez, G., 303, 382, 383 Dopp, E., 344 Drobniewska, A., 172 Duda-Chodak, A., 346 Duta, F.P., 230, 232 Dutta, K., 72

417 Dye degradation, 139, 142, 148–150, 152, 156, 157, 198, 306 Dyes, 75, 86, 138, 214, 232, 278, 290, 344 Dziubanek, G., 346

E Ebrahimpour, H., 51 Eichlerová, I., 307 Eke, P.M., 333, 334 El Mhammedi, M.A., 162–174 Elkhatib, E., 357 Endocrine disruptors, 254, 255, 260 Ensari, N.Y., 382 Environmental, v, vi, vii, 2, 22, 29, 30, 32, 34, 42–47, 49, 53, 55, 56, 62, 88, 98, 106, 118–120, 123–125, 128, 129, 138–140, 142, 164, 166, 167, 183, 191, 226, 227, 229, 232, 233, 242, 254, 255, 258–260, 268–272, 274, 280, 290–312, 328, 332, 335, 342–345, 347, 348, 351–353, 355, 358–361, 363, 374, 376, 379, 394–398, 404, 407 Environmental effects, 289–312, 376, 377 Environmental pollution, 29, 56, 62, 98, 123, 255, 268, 272, 343, 358, 394, 396, 398, 404 Eskandari, F., 311 Eslami, H., 311 Eslami, M., 304 Eswaramoorthi, S., 295 Exopolymeric substances, 226–247 Exopolysaccharide, 5–7, 95–97, 99, 104, 105, 107, 118, 123, 124, 126–128 Ezhilarasu, A., 303

F Fadel, M., 402 Fahiminia, M., 32 Falavarjani, E.R., 303 Falowo, O.A., 342–363 Fan, J., 401 Fan, T., 382 Farooq, U., 228 Feitz, O., 168 Feuillade, J., 166 Flaws, J.A., 256 Fourest, E., 383 Fox, L.J., 334 Franciscon, E., 299, 305 Freshwater, 124, 162–174, 206, 260, 298, 377

418 Freundlich model, 188 Fu, Y.Q., 402 Fungi, vi, vii, 2, 3, 15, 88, 94, 95, 139, 141, 181, 183, 196, 198, 202–204, 206, 219, 220, 228–230, 232, 234, 246, 254, 255, 260, 291, 301, 306–310, 374, 380, 382, 384, 395–397, 399, 400, 402, 403

Index Heavy metals, 2, 28, 86, 119, 171, 182, 214, 254, 268, 278, 330, 343, 374, 394 Hemapriya, J., 302 Hemond, H.F., 329 Henriques, M.C., 346 Hepatotoxins, 165, 166 Hinchman, R.R., 357 Holan, Z.R., 381, 383 Hospital, 61–78 Hou, J., 73 Huang, J., 149, 152, 154, 155 Huang, J.W., 357 Hussain, M.A., 182–206 Hussain, S., 303 Hussein, F.H., 295 Hussein, H., 230, 232 Hussein, S., 290–312 Hyperaccumulators, 354–356, 361, 362, 405

G Gahlout, M., 308 Galedar, M., 382 Galun, M., 381 Ganta, M., 254–262 Gao, X., 256 Gayathri, R., 304 Genetic factors, 104–105 George, K.S., 278–286 Ghaly, A.E., 190, 295, 296 Ghodake, G.S., 302 Gomaa, O.M., 299 Gomes, N., 402 González, R.C., 357 González-Chávez, M., 357 Gopinath, K.P., 304 Gou, M., 204 Goyal, D., 401 Grigoras, A.G., 2–22 Guba, E.G., 51 Gudmalwar, R.M., 302 Guha, A.K., 295 Gulati, D., 302 Guo, G., 311 Gupta, S., 402, 406 Gupta, V.K., 190 Gurbanov, R., 118–129 Gutiérrez, C., 382

I Ibrahim, W.M., 382 Igiri, B.E., 401, 402 Ikawa, M., 166 Iman, M.T., 39, 51 Immobilization, v, vi, 93, 107, 128, 140, 145, 147–150, 152, 154, 243, 258, 271, 348, 350, 351, 404 Immobilized enzymes, 140 Imran, M., 300, 302 Industrial waste, 4, 21, 62, 73, 89, 106, 182, 187, 198, 203, 230, 341–363 Industrialization, v, 62, 183, 227, 254, 278, 282, 342 Iqbal, H.M.N., 138–157 Isik, M., 304 Isotherm models of absorption, 10

H Ha, E., 344 Haarstad, K., 333 Hadibarata, T., 204, 308 Handayani, W., 303 Hannan, M.A., 295 Hannon, P.R., 256 Hao, J., 204 Hardware, 27–56 Harry-asobara, J.L., 305, 308, 311 Hashim, M.A., 230, 381 He, X., 307 Health effects, 72, 98, 255, 375, 377, 395 Heavy metal uptake, 274

J Jacob, J.M., 401, 402 Jadhav, S.B., 301, 304 Jaffré, T., 356 Jagwani, J., 312 Jain, P.P., 278–286 Jaishankar, M., 344 Janga, N., 240 Järvenpää, S., 168 Jasiñska, A., 143 Javed, M.T., 172 Jeon, S.J., 308 Jha, I., 302 Jha, P., 297

Index Jin, D., 260 Jin, R., 303 Jin, Y., 401 Jiranuntipon, S., 200 Jobby, R., 346, 383 Joe, M., 304 John, G.H., 304 Joshi, N.C., 402 Joshi, P.A., 311 Joshi, S.S., 295

K Kalpana, D., 307 Kalra, S.S., 295 Kalve, S., 356, 357 Kamble, L.H., 302 Kamei, I., 308, 311 Kang, C.H., 401 Kaparapu, J., 382 Karasali, H., 328–338 Karim, M.E., 311 Kariuki, Z., 230, 232, 402 Kashefi, S., 148, 149 Kashem, M.A., 357 Kavya Raj, K., 268–274 Kaya, Y., 72 Kazemeini, F., 406 Kehinde, F.O., 295 Kerkeb, L., 358 Khadir, A., 62–78 Khadivinia, E., 382 Khalid, A., 305 Khan, M.A., 344, 346 Khan, M.A.A., 295 Khodadoust, A.P., 345 Khouni, I., 201 Kiayi, Z., 307 Kilic, N.K., 303 Kim, I.H., 401 Kim, W.-S., 349 Kinetics, 2, 8–11, 15, 22, 65, 72, 98, 157, 186– 189, 193, 199, 202, 218, 273, 348, 349, 379 Kiran, S., 307 Kline, R.R., 34 Knauert, S., 172 Kochher, S., 302 Krämer, U., 358 Krishnamoorthy, R., 311 Krishnaswamy, G.V., 312 Kulkarni, A.N., 307 Kulkarni, R.M., 375–384

419 Kumar, J., 302 Kumar, S.S., 301 Kumar, V., 406 Kumar, V.V., 156 Kumari, S., 15, 17 Kunnel, S.G., 254–262 Kuo, T.S., 304 Küpper, H., 357, 358 Kuppusamy, S., 304, 311 Kurade, M.B., 204, 311 Kurella, B.R., 254–262 Kurki-Helasmo, K., 167

L Laccase, vi, 137–157, 191, 192, 194, 198, 199, 201–203, 258–260, 309 Lade, H., 297, 302, 310, 311 Lade, H.S., 297 Laing, I.G., 294 Lakshmipathy, R., 230, 231 Langmuir model, 187–189 Lanphear, B.P., 344 Laxmi, S., 306 Lead, 2, 28, 66, 87, 124, 157, 162, 203, 227, 268, 278, 296, 334, 344, 374, 394 Lefevre, M., 166 Leo, V.V., 143 Leung, H., 358 Li, D., 17, 20 Li, H., 190 Li, M., 259, 406 Li, W.C., 75 Li, Y.-M., 356 Li, Z., 227, 349 Lim, L.B., 190 Lim, S.J., 308 Lin, C.H., 310, 311 Lin, J., 303, 305 Lincoln, Y.S., 51 Ling, Z.R., 144 Lins, P., 69 Liu, L., 347 Liu, W.J., 240 Lombi, E., 356 Lone, M.I., 357 Long, J., 384 Low-cost, 2, 62, 73, 94, 107, 129, 154, 171, 199, 214, 218, 221, 282, 351, 359, 378, 397 Lu, W.B., 382, 401 Lutts, S., 356

420 M M-Hamvas, M., 167 Ma, L., 307 Ma, L.Q., 356 Madhu, G., 190 Mahapatra, S., 240 Mahdavinia, G.R., 190 Mahmood, F., 290–312 Mahmood, R., 312 Malik, O.A., 32 Management, 29–34, 36, 37, 39, 43, 50–56, 95, 138, 206, 254, 262, 278, 286, 291, 336, 351 Manonmani, S., 190 Mansel, C., 346 Maqbool, Z., 290–312 Mardy, A., 62–78 Martínez-Juárez, V.M., 204 Mary Lissy, P.N., 190 Masarbo, R.S., 311 Mattuschka, B., 383 Mawgoud, Y.A., 19 Mbuligwe, S.E., 334 McGrath, S., 358 Mechanisms, 2, 37, 66, 92, 118, 142, 162, 184, 215, 226, 271, 285, 328, 343, 377, 395 Mehmood, A., 182–206 Membrane, 5, 69, 89, 122, 147, 169, 193, 214, 229, 268, 279, 299, 349, 378, 398 Mendoza, L., 152 Meo, M., 63 Meriluoto, J., 167 Mesjasz-Przybyłowicz, J., 356 Metal removal, 108, 125–128, 200, 201, 218, 225–247, 348, 349, 363, 378, 379, 381, 384, 402–404 Metals, 2, 28, 86, 118, 144, 171, 182, 214, 226, 254, 268, 293, 328, 342, 374, 394 Microbial detoxification, 289–312 Microorganisms, v, vi, vii, 2, 4, 5, 8, 15, 22, 66, 76, 86–89, 91, 94, 95, 98, 101, 106, 119, 124, 125, 140, 144, 163, 171, 183, 187, 206, 219–221, 226, 229–232, 255, 258, 261, 272, 278, 291, 299, 300, 306, 329, 330, 343, 374, 376, 380, 395–400, 402, 405, 407, 408 Miller, S.M., 190 Mishra, S., 311, 346, 358 Mkandawire, M., 406 Modi, H.A., 302 Modi, S., 302 Moino, B.P., 384 Mollahosseini, A., 62–78

Index Mondal, N.K., 402 Monsalvo, V., 70, 71 Morrison, J.M., 304 Muhammad, G., 182–206 Murugesan, K., 307, 308 Muthezhilan, R., 307 Muzammil, S., 290–312 Mycoremediation, vi

N Nadeem, H., 290–312 Nahar Ali, Z., 278–286 Najari, R., 51 Najme, R., 303 Nandy, T., 63 Nascimento, C., 299 Naskar, A., 384 Nathalie, V., 406 Natural, 3, 39, 87, 118, 166, 193, 217, 243, 255, 268, 278, 290, 328, 342, 394 Navada, K.K., 142, 143, 145 Navarro, A., 351 Nayak, A.K., 401 Neetha, J.N., 304 Negarestani, M., 62–78 Negri, C.M., 357 Nematian, M.A., 406 Nesterenko-Malkovskaya, A., 190 Neurotoxins, 165, 166 Nezakati, R., 53 Ng, K.K., 63 Ngieng, N.S, 307 Nguyen, T.A.H., 229, 230 Nickel, vii, 28, 88, 90, 97, 101, 127, 128, 192, 200, 228, 229, 268–270, 272, 273, 279, 280, 283, 284, 294, 297, 334, 345, 346, 374–384 Nikam, T.D., 306 Ning, Y.J., 144 Noormohamadi, H.R., 382, 384

O Oak, U., 302 Ojuederie, O., 406 Okieimen, F.E., 344, 346, 347 Oladipo, B., 342–363 Olatunji, S.O., 342–363 O’Neill, C., 290 Organic pollutants, v, 171, 190, 205, 227, 330, 337, 347, 394, 407 Ortiz-Monsalve, S., 307

Index Oturkar, C.C., 304 Oudra, B., 162–174 Oves, M., 382, 383 Ozdemir, G., 243 Özdemir, S., 382, 383, 401 Öztürk, A., 383

P Pan, H., 344 Pandey, A., 268–274, 382 Pandi, A., 143, 146, 308 Paradigm, 37–56, 142 Park, J.H., 382 Parmar, M., 229, 230 Parrott, J.L., 299 Parshetti, G.K., 303 Patel, V.R., 311 Pavlidis, G., 328–338 Payus, C., 190 Peall, S.K., 333 Pellerin, C., 344 Pereira, A.R., 402 Pesticides, 2, 30, 53, 142, 228, 270, 278, 279, 329–331, 333, 335, 337, 344 Peuthert, A., 169 Pflugmacher, S., 173 Pharmaceutical, 6, 62–78, 125, 235, 333, 343, 381 Philips, E.M., 256 Phthalates, vi, 253–262 Phuengprasop, T., 230, 231 Phugare, S.S., 311 Phytoextraction, vii, 268–274, 330, 342–363, 404, 405 Phytoremediation, vi, vii, 4, 20, 162, 163, 171–174, 182, 183, 185, 189, 190, 255, 261, 270, 273, 274, 284, 330–337, 342, 343, 348, 351–353, 361, 363, 404–406 Pierce, J., 290 Plants, 3, 33, 64, 87, 124, 140, 162, 183, 214, 227, 257, 268, 278, 291, 329, 343, 375, 394 Pointing, S.B., 307 Pokharia, A., 301, 302, 304 Polepalli, S., 190 Pollution, vi, 20, 28–56, 62, 86, 98, 123, 126, 139, 157, 162, 189, 205, 226, 227, 254, 255, 268–270, 272, 278, 282, 283, 286, 291, 328, 329, 334, 335, 342, 345, 352, 358–360, 366, 374, 378, 394–396, 398, 400 Pradhan, S., 238

421 Pramanik, S., 308 Prasad, A., 302 Prasad, M.K., 382 Prasad, M.N.V., 356 Prasad, N.K., 307 Prasad, S.S., 303 Prasher, S.O., 381, 383 Priester, J.H., 244 Priya, E.S., 172 Przystas, W., 299, 306, 307, 309 Pseudomonas, 2, 89, 123, 189, 230, 257, 300, 380, 398 Puschenreiter, M., 358 Puthur, S., 268–274

Q Qu, Y., 200 Quorum sensing, vi, 118, 121–129

R Rahman, N.N., 382 Rai, P.K., 357 Rajendran, R., 302, 303 Rajesh, V., 382 Ramírez-Montoya, L.A., 307 Rao, C.P., 190 Raza, M.A., 182–206 Rebello, S., 268–274 Reda, F.M., 149 Reduction, 53, 66, 67, 74, 88, 103, 105, 125–127, 142, 167, 184, 189, 195, 199, 260, 272, 279, 281–286, 297, 305, 306, 328, 331–337, 342, 347, 348, 353, 395, 398, 404, 405 Rehman, A., 201, 204 Rekik, H., 308 Remediation, 2, 28, 88, 118, 147, 169, 186, 221, 273, 328, 347, 378, 395 Remediation (674) Removal, 4, 31, 64, 88, 125, 139, 162, 183, 226, 258, 280, 310, 333, 347, 375, 395 Ren, S., 305 Resources, 4, 27–56, 106, 119, 162, 163, 185, 231, 268–271, 273, 278, 279, 290, 296, 297, 333, 335, 342, 397 Ressom, R., 168 Robinson, B., 356 Rodriguez, E., 169 Rodríguez, M.J., 63 Rohini, A., 307 Romero-Oliva, C.S., 173

422 Roosta, A., 382 Rosestolato, D., 349 Roux, J.C., 383 Roy, R., 295 Ruholahi, F., 384 Ryan, S., 307

S Sadaf, S., 302 Safari, M., 382 Saha, I., 86–108 Sahasrabudhe, M.M., 304 Saif, M.M.S., 190 Saint-Jacques, N., 346 Sakakibara, M., 357 Salem, H.M., 346 San, N.O., 200 Sandhya, S., 304 Santhi, T., 190 Saqrane, S., 162–174 Sar, A., 382 Sarada, N.C., 230, 231 Saratale, R., 200 Saratale, R.G., 300, 302, 306, 311 Saravanan, A., 214–223 Sarayu, K., 304 Sasmaz, M., 172 Satarug, S., 346 Saunders, F.M., 172 Schmidt, C., 304 Scholz, M., 333, 334 Schulz, R., 333 Selatnia, A., 381, 383 Selvan, P.S., 172 Sengupta, D., 86–108 Senthil Kumar, P., 214–223 Seow, N., 156 Seth, C.S., 358 Sethi, S., 302 Severcan, F., 118–129 Sha, M.P., 303 Shabbir, S., 157 Shah, B., 311 Shah, M., 302 Shahid, M., 182–206, 290–312 Shahvali, M., 22–57 Shahzad, T., 290–312 Shaker, M.A., 11, 12, 16, 18, 19 Shanmuga, B.K., 311 Sharma, B.M., 346 Sharma, R., 302 Sharma, S., 303

Index Sharma, S.C.D., 305 Shelley, T.R., 294 Sheng, G.P., 242, 244 Shetty, K.V., 375–384 Sheveleva, I.V., 230, 231 Shilli, A., 254–262 Shinkafi, M.S., 303 Siddique, M.H., 290–312 Siegel, S.M., 381 Silveira, E., 304 Sim, C.S.F., 238 Sindhu, R., 268–274 Singh, R.K., 204 Singh, R.P., 200 Sinha, 401 Sinirlioglu, Z.A., 155, 156 Sirianuntapiboon, S., 201 Sisodiya, S.J., 278–286 Sludge blanket, 73–76, 78 Smith, A.L., 70 Snow, A., 190 Soares, M.A., 350 Software, 14, 28–56 Sponza, D.T., 304 Sreekanth, D., 63 Srinath, T., 401 Srinikethan, G., 375–384 Srinivasan, R., 303 Sriram, N., 304 Srivastava, M., 356 Straube, G., 383 Stuckey, D.C, 71 Su, W.T., 310, 311 Sud, D., 229 Sudha, M., 303 Sultana, Z., 295 Sun, J., 308 Suruthi, S., 303 Sustainable approach, 144 Suyamud, B., 257 Synergism, 15, 19, 310 Systems, 7, 29, 65, 87, 121, 139, 164, 184, 218, 227, 255, 269, 279, 299, 328, 377, 396

T Tafesse, T.B., 295 Tamura, H., 357 Tan, T., 381, 383 Tchounwou, P.B., 344, 346 Technology, 5, 29, 32–34, 36, 37, 53, 55, 56, 70, 104, 108, 123, 147, 171, 186, 192,

Index 205, 300, 336, 343, 348–352, 359, 363, 384, 396, 408 Tee, H.C., 297 Tekere, M., 394–408 Textile dyes, 100, 101, 290–312 Textile wastewater, 196, 204, 290, 294–301, 306, 309, 310 Thakur, L.S., 229, 230 Thakura, R., 63 Ting, A.S.Y., 226–247, 308 Tobin, J.M., 381 Tondee, T., 201, 204 Toxic dyes, 86–108, 140, 188, 203 Toxic heavy metals, 2–22, 87, 89, 95, 97, 104, 183, 184, 268, 270, 278, 279, 282 Toxic pollutants, 214–223 Toxicity, 20, 87, 125, 145, 162, 192, 227, 255, 270, 279, 296, 328, 346, 394 Toxicology, v, 167, 290 Treatment, 5, 29, 62, 87, 142, 162, 182, 215, 231, 278, 291, 328, 344, 374, 395 Tree bark, 278–286 Trovaslet, M., 144 Tseng, C.-H., 346 Tseng, I.L., 256 Tsihrintzis, V.A., 335 Tu, C., 356 Tunggolou, J., 190 Turgut, Z., 295 Turhan, K., 295 Tursi, A., 259 Tuttolomondo, M.V., 190, 200 Tuzen, N., 382

423 W Waghmode, T.R., 298, 306, 307, 309, 311, 313 Walker, P.L., 356–358 Wang, H.S., 256 Wang, N., 11, 12, 16, 17, 308 Wang, Y., 402 Warężak, T., 334 Waste disposal, v, 98, 268, 270, 291, 342–363 Wastewater, 2, 28, 62, 95, 124, 139, 171, 182, 214, 226, 258, 270, 278, 290, 328, 343, 374, 395 Wastewater treatment, 29, 30, 32–34, 62, 64, 65, 69, 72, 73, 77, 78, 155, 171, 183, 185, 187, 193, 205, 206, 221, 244, 296, 301, 328, 332, 336 Watanapokasin, R.Y., 200 Water, 3, 28, 62, 87, 118, 138, 162, 182, 214, 227, 254, 268, 278, 290, 328, 342, 374, 396 Water purification, 126, 127, 162, 182–206, 279, 285 Water treatment, 29, 30, 32–36, 50, 53, 54, 56, 64, 69, 72, 73, 77, 78, 155, 171, 183, 184, 187, 193, 205, 206, 215, 221, 223, 244, 296, 328, 332, 336 Wei, S., 356 Wei, X., 238 Welch, B., 346 Wen, X., 149 Wenzel, W., 358 Wijekoon, K.C., 70, 71 Williams, D.R., 168 Wu, G., 149, 152, 153 Wuana, R.A., 344

U Upadhye, V.B., 295

V Vale, M., 230, 232 Van Donk, E., 166 Vanhulle, S., 309 Veglio, F., 381, 383 Velvizhi, M., 312 Veneu, D.M., 382 Verma, K., 304 Vijayanand, S., 302 Vijayaraghavan, K., 230, 231 Virarghavan, T., 381, 383 Vishan, I., 382 Volesky, B., 383 Volesky, V., 381 Vrijmoed, L.L.P., 307

X Xia, J., 143 Xiao, X., 303 Xiao, Y., 72 Xing, Z.-P., 63 Xu, H.M., 149 Xu, M., 303 Xu, X., 11, 12, 17, 19

Y Yan, G., 381, 383 Yang, J., 156 Yang, K.L., 156 Yang, Q., 204 Yang, T., 260, 261 Yang, X., 357, 358

424 Yetis, U., 245 Yilmaz, E.I., 382 Yin, F., 68 Yin, L., 168, 173 Yoshinaga, M., 346 Younesi, H., 382 Yous, R., 382, 384 Yusoff, M.S., 190

Z Zablocka-Godlewska, E., 200, 299, 309 Zahoor, A., 201 Zapata-Castillo, P., 307

Index Zhang, S., 259 Zhang, X., 172, 242, 260, 357, 382 Zhang, Y., 170 Zhang, Z., 358, 406 Zhao, F., 356, 357 Zhao, M., 305 Zhao, X., 400 Zhao, Y., 230 Zhu, X., 200 Zhuang, P., 406 Zhuo, R., 144 Zouboulis, A.I., 382, 401 Zulkali, M.M., 406