Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, 2-Volume Set [2 ed.] 3527347224, 9783527347223

Citation preview

Sustainable Resource Management

Sustainable Resource Management Technologies for Recovery and Reuse of Energy and Waste Materials

Edited by Wenshan Guo Huu Hao Ngo Rao Y. Surampalli Tian C. Zhang

Volume 1

Sustainable Resource Management Technologies for Recovery and Reuse of Energy and Waste Materials

Edited by Wenshan Guo Huu Hao Ngo Rao Y. Surampalli Tian C. Zhang

Volume 2

Editors Prof. Wenshan Guo

Civil & Environmental Engineering University of Technology Sydney 15 Broadway 2007 Ultimo Australia

All books published by Wiley-VCH are carefully produced. Nevertheless, authors, editors, and publisher do not warrant the information contained in these books, including this book, to be free of errors. Readers are advised to keep in mind that statements, data, illustrations, procedural details or other items may inadvertently be inaccurate. Library of Congress Card No.:

Prof. Huu Hao Ngo

Civil & Environmental Engineering University of Technology Sydney 15 Broadway 2007 Ultimo Australia Prof. Rao Y. Surampalli

Global Institute for Energy Environment and Sustainability (GIESS) P.O. Box 14354 KS United States Prof. Tian C. Zhang

College of Engineering University of Nebraska-Lincoln 200E 110 South 67th Street Scott Campus NE United States Cover

© Courtesy of Wenshan Guo

applied for British Library Cataloguing-in-Publication Data

A catalogue record for this book is available from the British Library. Bibliographic information published by the Deutsche Nationalbibliothek

The Deutsche Nationalbibliothek lists this publication in the Deutsche Nationalbibliografie; detailed bibliographic data are available on the Internet at . © 2021 WILEY-VCH GmbH, Boschstr. 12, 69469 Weinheim, Germany All rights reserved (including those of translation into other languages). No part of this book may be reproduced in any form – by photoprinting, microfilm, or any other means – nor transmitted or translated into a machine language without written permission from the publishers. Registered names, trademarks, etc. used in this book, even when not specifically marked as such, are not to be considered unprotected by law. Print ISBN: 978-3-527-34722-3 ePDF ISBN: 978-3-527-82537-0 ePub ISBN: 978-3-527-82538-7 oBook ISBN: 978-3-527-82539-4 Typesetting

SPi Global, Chennai, India

Printed on acid-free paper 10 9 8 7 6 5 4 3 2 1

v

Contents

Volume 1 Preface xix 1

1.1 1.2 1.2.1 1.2.2 1.2.2.1 1.2.2.2 1.2.2.3 1.2.2.4 1.3 1.3.1 1.3.1.1 1.3.1.2 1.3.1.3 1.3.2 1.4 1.4.1 1.4.2 1.4.3 1.5

Resource Recovery and Reuse for Sustainable Future Introduction and Overview 1 Wenshan Guo, Huu Hao Ngo, Lijuan Deng, Rao Y. Surampalli, and Tian C. Zhang Introduction 1 Background 2 Hierarchy of Resource Use 2 Analyzing the Needs for Resource and Energy Recovery and Reuse 2 Population Growth 2 Resource Scarcity 4 Environmental Impacts 4 Economical Aspect 4 Current Status of Resource Recovery and Reuse 5 Wastewater 5 Nutrient Recovery 6 Organic Carbon Recovery 6 Heat Recovery 7 Waste 7 Research Needs 9 Development of Novel Technologies 9 Social and Economic Feasibility of Resource Recovery and Reuse 9 Development of Internationally Coordinated Framework and Strategy 10 Book Overview 10 References 17

vi

Contents

2

2.1 2.1.1 2.1.2 2.1.2.1 2.1.2.2 2.1.2.3 2.1.2.4 2.1.3 2.1.3.1 2.1.3.2 2.1.3.3 2.1.3.4 2.2 2.2.1 2.2.2 2.2.3 2.2.4 2.3 2.3.1 2.3.2 2.3.3 2.3.4 2.3.5 2.4 2.5

3

3.1 3.2 3.3 3.4 3.4.1 3.4.2 3.4.3 3.4.4

Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy 21 Ranaprathap Katakojwala, Hari Shankar Kopperi, Althuri Avanthi, and S. Venkata Mohan Introduction 21 Global Food Waste Production 22 Conventional Food Waste Management Practices 23 Land Filling 23 Fertilizer/Animal Feed 23 Incineration 23 Composting 24 Advanced Food Waste Management Methods 24 Acidogenesis 24 Solventogenesis 24 Biodiesel 25 Bioplastics 26 Significance of Hydrothermal Liquefaction of Food Waste 26 HTL Reactor Operation 27 Isothermal HTL and Fast HTL 30 HTL Products 30 Greenhouse Gas Emissions 31 Factors Influencing HTL During FW Treatment 32 Temperature 34 Reaction Time 35 Solid-to-Solvent Ratio 35 Composition of Food Waste 36 Catalyst Concentration 36 HTL of Food Waste: Case Studies 37 Conclusions and Future Scope 39 Acknowledgement 40 References 40 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England 47 Pauline Deutz and Anne Kildunne Introduction 47 Sustainability Transitions Literature 48 Waste Management in England 51 Research Design and Methods 52 Research Design 53 Methods 53 Selection of Interviewees 54 Secondary Data 58

Contents

3.5 3.5.1 3.5.2 3.5.3 3.5.3.1 3.5.3.2 3.5.3.3 3.5.3.4 3.5.4 3.5.5 3.6

Results and Discussion 58 English Waste in the Context of the EU 58 Influences in the UK Context for LAs 64 Implementation of the 2000 Waste Strategy 66 LA Implementation of Waste Policy 67 Targets 70 Financial Instruments 70 Regional Governance 72 Local Authorities and the Public 72 Legacy of the Strategy 74 Conclusions 75 Acknowledgements 77 References 77

4

Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production 83 Huihui Chen, Gang Luo, and Shicheng Zhang Introduction 83 Composition of Lignocellulosic Biomass and their Degradation in HTL Processes 85 Composition of Lignocellulosic Biomass 85 Brief Review on the Development of HTL Technology 85 Main Components Degradation of the Lignocellulosic Biomass During HTL 87 Cellulose and its Degradation in HTL Processes 87 Hemicellulose and its Degradation in HTC Process 88 Lignin and its Degradation in HTC Processes 88 Research Status in HTL of Lignocellulosic Biomass 90 Products Description 90 Bio-oil 90 Solid Residue 90 Other By-products 91 Operating Parameters for Bio-oil Production by HTL 91 Bio-oil 92 Temperature 93 Heating Rate 93 Residence Time 94 Pressure 94 Catalysts 95 Liquid-to-Solid Ratio 96 Limitations and Prospects for Bioenergy Production from Lignocellulosic Biomass by HTL 97 Poor Quality of Crude Bio-oil 97

4.1 4.2 4.2.1 4.2.2 4.2.3 4.2.3.1 4.2.3.2 4.2.3.3 4.3 4.3.1 4.3.1.1 4.3.1.2 4.3.1.3 4.3.2 4.3.2.1 4.3.2.2 4.3.2.3 4.3.2.4 4.3.2.5 4.3.2.6 4.3.2.7 4.4 4.4.1

vii

viii

Contents

4.4.2 4.4.3 4.5

Aqueous By-products Utilization 97 Prospects 98 Conclusion and Future Work 98 References 99

5

Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management 109 Sudheer Salana, Tuhin Banerji, Aman Kumar, Ekta Singh, and Sunil Kumar Introduction 109 Present Scenario 111 Ecological Sanitation 112 Rottebehaelter and Centrifugal Separation Sanitation 113 Biofilters, Vermicomposting Units, Bag Toilets 114 Failure, Success, and Lessons 115 Resource Recovery Options in Rural Areas 116 Nutrient Recovery from Urine 117 Anaerobic Digestion or Composting? 119 Community-Scale or Household Models? 121 Resource Recovery Sanitation in Urban Context 121 Energy Matters 121 Johkasou Systems 123 Possibilities of Industrial-Scale Units 124 Life Cycle Assessment of Sanitation Systems 125 Human Excreta and Sustainable Future 127 Economics of Resource Recovery Sanitation 127 Sanitation Access and Resource Recovery 128 Conclusion and Recommendations 130 References 131

5.1 5.2 5.2.1 5.2.1.1 5.2.1.2 5.2.2 5.3 5.3.1 5.3.2 5.3.3 5.4 5.4.1 5.4.2 5.4.3 5.5 5.6 5.6.1 5.6.2 5.7

6

6.1 6.2 6.2.1 6.2.1.1 6.2.2

Resource Recovery and Recycling from Livestock Manure: Current Statue, Challenges, and Future Prospects for Sustainable Management 137 Tao Liu, Hongyu Chen, Junchao Zhao, Parimala Gnana Soundari, Xiuna Ren, Sanjeev Kumar Awasthi, Yumin Duan, Mukesh Kumar Awasthi, and Zengqiang Zhang Introduction 137 Present Scenario and Global Perspective of Manure Generation and Recycling 139 Sanitization and Hygiene in Manure Management 139 Aerobic Composting 139 Importance and Significance of Resource Recovery 141

Contents

6.2.2.1 6.2.2.2 6.3 6.3.1 6.3.2 6.3.3 6.3.4 6.4 6.4.1 6.4.2 6.4.2.1 6.4.2.2 6.4.2.3 6.5 6.5.1 6.5.1.1 6.5.1.2 6.5.1.3 6.5.1.4 6.5.1.5 6.5.2 6.6 6.7 6.7.1 6.7.2 6.7.3 6.8 6.9

7

7.1

Nitrogen and Phosphorus Recovery from Livestock Manure 141 Heavy Metal Recovery from Livestock Manure 142 Resource Recovery Technologies and Logistics for Handling, Transport, and Distribution of Manures 142 Nutrient Recovery from Manure 142 Bioenergy Production by Anaerobic Digestion/Co-digestion 147 Composting/Co-composting 147 Centralized and De-centralized Models? 148 Energy Matters and Economic Feasibility 149 Energy Production 149 Mineral Reutilization 150 Ammonia Stripping 150 Struvite Crystallization 150 Mineral Concentrates 150 Resource Recovery Sanitation in Developed and Developing Countries 151 Operational Guidelines for Septage Treatment and Disposal 153 Storage 154 Pasteurization 154 Chemical Treatments 154 Anaerobic Treatments 154 Composting 155 Testing the Possibilities of Commercial-Scale Resource Recovery 155 Life Cycle Assessment of Sustainable Manure Management Systems 156 Innovation in Sustainable Manure Management Systems and Recycling 157 Economics of Resource Recovery from Manure and Sanitation 157 Business Models for a Circular Economy 158 Enabling Environment Sanitation and Financing for Resource Recovery 159 Challenges and Limitation 160 Conclusion and Future Prospects 160 Acknowledgements 161 References 161 Utilization of Microalgae and Thraustochytrids for the Production of Biofuel and Nutraceutical Products 167 Ying Liu and Jay J. Cheng Introduction 167

ix

x

Contents

7.1.1 7.1.2 7.1.3 7.1.4 7.2 7.2.1 7.2.2 7.2.2.1 7.2.2.2 7.2.2.3 7.2.2.4 7.2.2.5 7.3 7.4 7.5 7.6 7.6.1 7.6.2 7.6.3

8

8.1 8.2 8.3 8.3.1 8.3.2 8.3.3 8.4 8.5 8.6 8.7 8.8 8.9

Microalgae 167 Thraustochytrids 167 Biodiesel and Biobased Jet Fuel 168 Docosahexaenoic Acid (DHA) and Eicosapentaenoic Acid (EPA) 168 Microalgae for Biodiesel and Jet Fuel Production 169 Selection of Microalgae 169 Processes of Microalgae to Biofuel 170 Microalgae Cultivation 170 Microalgae Harvesting 172 Extraction of Oil from Microalgae 174 Biodiesel Production from Microalgal Oil 175 Jet Fuel Production from Microalgal Oil 176 Thraustochytrids for Biodiesel Production 177 Challenges of Microalgae and Thraustochytrids to Biofuel 178 Microalgae and Thraustochytrids for DHA and EPA Productions 179 Future Perspectives 183 Integrated Microalgae/Thraustochytrids Cultivation and Harvesting System 183 Genetically Modified Microalgae/Thraustochytrids for High Oil and Easy Extraction of Lipids 184 Integrated Microalgae/Thraustochytrids System for Biofuel and DHA/EPA Production 186 References 186 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective 199 Goldy De Bhowmick and Ajit K. Sarmah Introduction 199 Current Status of Algal Energy and Bio-product Formation Analysis of Conversion Methods 202 Dynamics of Algal Biomass Composition 202 Conversion Routes 203 Product Yield and Market Value 204 Competent Applications of Algae 205 Biorefinery and Integrated Approaches 207 Technological Issues: Pros and Cons 208 Life Cycle Assessment 210 Techno-Economic Analysis (TEA) 211 Futuristic Options 212 References 213

200

Contents

9

9.1 9.2 9.2.1 9.2.2 9.2.2.1 9.2.2.2 9.3 9.3.1 9.3.1.1 9.3.1.2 9.3.2 9.3.2.1 9.3.2.2 9.3.3 9.3.3.1 9.3.3.2 9.4 9.4.1 9.4.1.1 9.4.1.2 9.4.2 9.4.2.1 9.4.2.2 9.4.2.3 9.4.2.4 9.4.3 9.5 9.5.1 9.5.2 9.6

Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater 217 Dong Wei, Bin Du, and Qin Wei Introduction 217 Types of Sludge in Wastewater Treatment Process 218 Activated Sludge 218 Granular Sludge 219 Anaerobic Granular Sludge 219 Aerobic Granular Sludge 220 Sludge-Based Activated Carbon for Wastewater Treatment 222 Production Method 222 ZnCl2 223 H3 PO4 223 Treatment of Dye Wastewater 224 MG Sorption onto Sludge-Based ACs 224 Mineral Acid Modification of AGS-Derived AC for MG Sorption 225 Treatment of Heavy Metal-Contained Wastewater 226 Heavy Metal Sorption onto Sludge-Based AC 226 Cu(II) Sorption onto AGS-AC in the Presence of HA and FA 227 Granular Sludge Biosorbent Applied for Wastewater Treatment 229 Treatment of Dye Wastewater 229 Role of EPS in Aerobic Granular Sludge for MB Sorption 229 Biosorption of Dye Wastewater and Photocatalytic Regeneration of AGS 230 Treatment of Heavy Metal-Contained Wastewater 232 Zn(II) Sorption onto AGS 232 Cu(II) Sorption onto AGS 232 Ni(II) Sorption onto AGS/AnGS 233 Magnetic Modification of AnGS for Pb(II) and Cu(II) Removal 234 Treatment of Multicomponent Contaminants 235 Applications of EPS Extracted from Sludge for Wastewater Treatment 236 Bioflocculant 236 Biosorbent for the Removal of Various Pollutants 237 Conclusion 238 References 238

xi

xii

Contents

10

10.1 10.2 10.3 10.4 10.4.1 10.4.1.1 10.4.1.2 10.4.2 10.4.2.1 10.4.2.2 10.4.3 10.4.3.1 10.4.3.2 10.4.4 10.4.4.1 10.4.4.2 10.5 10.5.1 10.5.2 10.5.3 10.5.4 10.5.5 10.5.6 10.6 10.6.1 10.6.2 10.6.3 10.6.4 10.7

Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery 247 Mian Hu, Dabin Guo, Yingqun Ma, and Yu Liu Introduction 247 Sewage Sludge and Its Impact on Environmental Sustainability 248 Characterization of Sewage Sludge 250 Thermal-Chemical Treatment of Sewage Sludge 250 Incineration 250 Typical Incineration Processes 250 Performance–Cost–Benefit Analysis of Incineration Technology 253 Pyrolysis 253 Typical Pyrolysis Processes 253 Performance–Cost–Benefit Analysis of Pyrolysis Technology 255 Gasification 255 Typical Gasification Processes 255 Performance–Cost–Benefit Analysis of Gasification Technology 257 Liquefaction 257 Typical Liquefaction Processes 257 Performance–Cost–Benefit Analysis of Liquefaction Technology 258 Recovery of Energy and Resource from Sewage Sludge 258 Combustible Gas 258 Bio-oils 259 Biochar 260 Ashes to Value-Added Materials 261 Nutrient Recovery 261 Heavy Metals Removal and Recovery 263 Technology Limitations and Challenges 264 Deactivation of Catalyst 264 Tar Formation 264 NOx and SOx Emission 265 High Moisture Content 265 Conclusions and Perspectives 266 References 267

Contents

11

11.1 11.2 11.2.1 11.2.2 11.3 11.3.1 11.3.2 11.4 11.4.1 11.4.1.1 11.4.1.2 11.4.1.3 11.4.1.4 11.4.1.5 11.4.2 11.4.2.1 11.4.2.2 11.4.2.3 11.4.2.4 11.4.3 11.5 11.5.1 11.5.1.1 11.5.1.2 11.5.1.3 11.5.1.4 11.5.1.5 11.5.1.6 11.5.2 11.5.2.1 11.5.2.2 11.5.2.3 11.5.2.4 11.5.2.5 11.5.2.6

Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge 275 Qilin Wang, Jing Wei, Huan Liu, Dongbo Wang, Long D. Nghiem, and Zhiyao Wang Introduction 275 Characteristics of Sewage Sludge 276 Primary Sludge 276 Waste Activated Sludge 276 Anaerobic Digestion for Bioenergy Recovery 279 Theory of Anaerobic Digestion 279 Bioenergy Recovery by Anaerobic Digestion 279 Technologies for Enhancing Methane Production from Sludge 280 Physical Pretreatment 280 Thermal Hydrolysis Pretreatment 280 Mechanical Pretreatment 281 Ultrasonic Pretreatment 282 Microwave Pretreatment 282 Focused Pulsed Pretreatment 282 Chemical Pretreatment or Dosage 283 Ozonation Pretreatment 283 Alkaline Pretreatment 283 Free Nitrous Acid Pretreatment 283 Free Ammonia Pretreatment 283 Biological Pretreatment 284 Technologies for Enhancing Hydrogen Production from Sludge 284 Physical Pretreatment 284 Thermal Pretreatment 284 Freezing/Thawing Pretreatment 288 Sterilization Pretreatment 288 Microwave Pretreatment 288 Ultrasonic Pretreatment 288 Gamma Irradiation Pretreatment 288 Chemical Pretreatment 289 Acid Pretreatment 289 Alkaline Pretreatment 289 Free Ammonia and Free Nitrous Acid Pretreatment 289 Ozone Pretreatment 289 Wet Oxidation Pretreatment 289 Calcium Peroxide Pretreatment 290

xiii

xiv

Contents

11.5.2.7 11.5.3 11.6 11.7

Triclocarban Pretreatment 290 Biological Pretreatment 290 Evaluation and Comparison of Technologies 290 Summary and Future Outlook 294 References 294

12

Recovery of Phosphorus from Wastewater and Sludge 305 Ruo-hong Li, Lin Lin, and Xiao-yan Li Introduction 305 P Recovery Technologies 306 Wet-Chemical Approach 306 Thermal Treatment 307 Chemical Precipitation 307 P Recovery Based on CEPS 307 P Recovery Based on Chemically Enhanced Membrane Bioreactors 308 Chemical Coagulation and Flocculation for Enhanced P Removal from Wastewater 309 Experimental Methods 309 Results and Discussion 310 Acidogenic Fermentation for P Release and Recovery from Sludge 312 Experimental Methods 312 Results and Discussion 312 Influence of Fe Dosage on Acidogenic Sludge Fermentation 312 Influence of Al Dosage on Acidogenic Sludge Fermentation 315 Recovery of Organic Carbon and P from the Semicontinuous Fermentation of CEPS Sludge 316 Summary 317 A Membrane Bioreactor with Fe Dosing and Sludge Fermentation for Enhanced P Removal and Recovery 317 Experimental Work 317 Results and Discussion 319 P Removal from Wastewater by Chemical Flocculation and MBR 319 Sludge Fermentation and P Recovery 321 Comparison of Acidification and Acidogenesis 325 Summary 326 Mechanisms of P Removal and Recovery from Wastewater Using an Fe-dosing Bioreactor and Co-fermentation 326 Experimental Work 326

12.1 12.1.1 12.1.1.1 12.1.1.2 12.1.1.3 12.1.2 12.1.3 12.2 12.2.1 12.2.2 12.3 12.3.1 12.3.2 12.3.2.1 12.3.2.2 12.3.2.3 12.3.3 12.4 12.4.1 12.4.2 12.4.2.1 12.4.2.2 12.4.2.3 12.4.3 12.5 12.5.1

Contents

12.5.2 12.5.3

P Speciation in the Aerobic MBR and Anaerobic Fermenters 327 Fe Speciation in the Aerobic MBR and Anaerobic Fermenters 329 12.5.4 P Extraction and Release from Sludge During Acidogenic Fermentation 330 12.5.4.1 Acidogenic Fermentation 330 12.5.4.2 Microbial Iron Reduction 331 12.5.4.3 Solubility of the Fe—P Complex 331 12.6 Conclusions 333 References 333 13

13.1 13.1.1 13.1.2 13.1.3 13.1.4 13.1.4.1 13.1.4.2 13.1.4.3 13.2 13.2.1 13.2.1.1 13.2.1.2 13.2.2 13.2.2.1 13.2.2.2 13.2.2.3 13.2.3 13.2.4 13.3 13.3.1 13.3.2

Magnetic Iron-Based Oxide Materials for Selective Removal and Recovery of Phosphorus 339 Irene Man Chi Lo, Baile Wu, and Jun Wan Introduction 339 Phosphorus Sources, Speciation, and Properties in Water 339 Phosphorus Pollution and Eutrophication 340 Phosphorus Removal and Recovery Technologies 340 Selective Removal and Recovery of Phosphorus from Water by Using Adsorption 341 Phosphate Adsorption Processes and Mechanisms 341 Current Adsorbents for Phosphate Removal 341 Selective Removal and Recovery of Phosphate from Water by Magnetic Iron Based-Oxide Materials 342 Development and Material Synthesis 343 Synthesis of Fe3 O4 Nanoparticles 343 Fe3 O4 Nanoparticles Synthesized by the Solvothermal Method 343 Fe3 O4 Nanoparticles Synthesized by the Coprecipitation Method 343 Synthesis of SiO2 @Fe3 O4 , ZrO2 @SiO2 @Fe3 O4 and ZrO2 @Fe3 O4 Nanoparticles 343 Synthesis of SiO2 @Fe3 O4 Nanoparticles 343 Synthesis of ZrO2 @SiO2 @Fe3 O4 Nanoparticles 344 Synthesis of ZrO2 @Fe3 O4 Nanoparticles 344 Synthesis of La(OH)3 /Fe3 O4 Nanocomposites 344 Synthesis of Fe0 /Fe3 O4 Composites 344 Material Characteristics 345 Characterization Methods for Magnetic Iron-Based Oxide Materials 345 Characteristics of Fe3 O4 , SiO2 @Fe3 O4 , ZrO2 @SiO2 @Fe3 O4 , and ZrO2 @Fe3 O4 Nanoparticles 345

xv

xvi

Contents

13.3.3 13.3.4 13.4 13.4.1 13.4.1.1 13.4.1.2 13.4.1.3 13.4.2 13.4.2.1 13.4.2.2 13.4.2.3 13.4.3 13.5 13.5.1

13.5.1.1 13.5.1.2 13.5.2 13.5.2.1 13.5.2.2 13.5.3 13.5.3.1 13.5.3.2 13.6 13.6.1 13.6.2 13.6.3

Characteristics of Fe3 O4 and La(OH)3 /Fe3 O4 Nanocomposites 348 Characteristics of Fe0 /Fe3 O4 Composites 350 Batch Adsorption Kinetics, Isotherms, and Affecting Factors 351 Phosphorus Removal by ZrO2 @SiO2 @Fe3 O4 and ZrO2 @Fe3 O4 Nanoparticles 351 Phosphate Adsorption Kinetics of ZrO2 @SiO2 @Fe3 O4 and ZrO2 @Fe3 O4 Nanoparticles 351 Phosphate Adsorption Isotherms of ZrO2 @SiO2 @Fe3 O4 and ZrO2 @Fe3 O4 Nanoparticles 351 Effects of pH and Zeta Potential Analysis 352 Phosphorus Removal by La(OH)3 /Fe3 O4 Nanocomposites 353 Phosphate Adsorption Kinetics of La(OH)3 /Fe3 O4 Nanocomposites 353 Phosphate Adsorption Isotherms of La(OH)3 /Fe3 O4 Nanocomposites 353 Effect of pH, Ionic Strength, and Zeta Potential Analysis 353 Phosphorus Removal by Fe0 /Fe3 O4 /Fe2+ System 355 Selective Removal and Recovery 357 Selective Phosphorus Removal and Recovery by ZrO2 @SiO2 @Fe3 O4 and ZrO2 @Fe3 O4 Nanoparticles and Their Reusability 357 Selective Phosphate Adsorption of ZrO2 @SiO2 @Fe3 O4 and ZrO2 @Fe3 O4 Nanoparticles 357 Phosphate Recovery and Reusability of ZrO2 @Fe3 O4 Nanoparticles 357 Selective Phosphorus Removal and Recovery by La(OH)3 /Fe3 O4 Nanocomposites 358 Selective Phosphate Adsorption of La(OH)3 /Fe3 O4 Nanocomposites 358 Phosphate Recovery and Reusability of La(OH)3 /Fe3 O4 Nanocomposites 360 Selective Phosphorus Removal and Recovery by Fe0 /Fe3 O4 /Fe2+ System 360 Selective Phosphate Removal of Fe0 /Fe3 O4 /Fe2+ System 360 Phosphate Recovery and Reusability of Fe0 /Fe3 O4 Composite 361 Comparison with Other Adsorbents 362 Phosphorus Removal Capacity 362 Phosphorus Removal Kinetics 363 Adsorbents Reusability and Phosphorus Recovery 364

Contents

13.7

Potential Environmental Applications and Perspectives References 366 Volume 2 Preface

xvii

14

Forward Osmosis for Nutrients Recovery from Wastewater 373

15

Removal and Recovery of Nutrients Using Low-Cost Adsorbents from Single-Component and Multicomponent Adsorption Systems 397

16

Use and Development of Biochar-Based Materials for Effective Capture and Reuse of Phosphorus 437

17

Recovery of Gold and Other Precious Metals by Biosorption 463

18

Bioelectrochemical System in Wastewater Treatment: Resource Recovery from Municipal and Industrial Wastewaters 489

19

Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater 525

20

Approaches Toward Resource Recovery from Breeding Wastewater 559

21

Resources Recovery and Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production 601

22

Recovery of Thermal Energy from Wastewater by Heat Pump Technology 635

23

Hydrocyclone-Separation Technologies for Resource Recovery and Reuse 663

24

Methane Recovery from Landfills 699

25

Resource Recovery from Electronic Waste 723 Index

755

365

xvii

v

Contents

Volume 1 Preface

xix

1

Resource Recovery and Reuse for Sustainable Future Introduction and Overview 1

2

Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy 21

3

Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England 47

4

Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production 83

5

Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management 109

6

Resource Recovery and Recycling from Livestock Manure: Current Statue, Challenges, and Future Prospects for Sustainable Management 137

7

Utilization of Microalgae and Thraustochytrids for the Production of Biofuel and Nutraceutical Products 167

8

Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective 199

9

Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater 217

vi

Contents

10

Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery 247

11

Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge 275

12

Recovery of Phosphorus from Wastewater and Sludge 305

13

Magnetic Iron-Based Oxide Materials for Selective Removal and Recovery of Phosphorus 339 Volume 2 Preface

14

14.1 14.2 14.3 14.4 14.4.1 14.4.2 14.5 14.6

15

15.1 15.2 15.2.1 15.2.2 15.3 15.4

xvii

Forward Osmosis for Nutrients Recovery from Wastewater 373 Yuanyao Ye, Wenshan Guo, Huu Hao Ngo, Nguyen Cong Nguyen, Tian C. Zhang, Rao Y. Surampalli, Hau Thi Nguyen, and Le Quang Huy Introduction 373 Forward Osmosis Technology 376 FO Systems for Nutrient Recovery 377 Recommended Systems and Key Challenges 385 Recommended Systems 385 Key Challenges 387 Future Roadmap 388 Conclusion 389 References 390 Removal and Recovery of Nutrients Using Low-Cost Adsorbents from Single-Component and Multicomponent Adsorption Systems 397 S. V. Manjunath and Mathava Kumar Introduction to Water Pollution 397 Nutrient Pollution in Aqueous Environment 398 Phosphate Pollution 398 Nitrate Pollution 400 Treatment Technologies for Removal of NO3 − and PO4 3− Nitrate and Phosphate Recovery Using Low-Cost Adsorbents 402

400

Contents

15.4.1 15.4.1.1 15.4.1.2 15.4.1.3 15.4.1.4 15.4.1.5 15.4.1.6 15.4.2 15.4.2.1 15.4.2.2 15.4.3 15.5 15.6 15.6.1 15.6.2 15.6.3 15.7 15.8

16

16.1 16.2 16.3 16.4 16.4.1 16.4.2 16.4.3 16.5

17

17.1 17.2

Factors Affecting Nutrient Adsorption 403 Effect of Initial Concentration 403 Effect of Adsorbent Dose 403 Effect of Temperature 404 Effect of Contact Time 404 Effect of pH 405 Effect of Interfering Anions 406 Kinetic and Isotherm Modeling 407 Kinetic Models 407 Isotherm Models 409 Mechanism of Nutrient Adsorption 409 Management of Spent Adsorbent 417 Removal and Recovery of PO4 3− and NO3 − Using Prosopis juliflora Weed 421 Preparation of Adsorbate and Adsorbent 421 Equilibrium Adsorption Study 421 Desorption Study for Recovery of PO4 3− and NO3 − from PJAC 425 Future Perspectives 426 Conclusions 427 References 427 Use and Development of Biochar-Based Materials for Effective Capture and Reuse of Phosphorus 437 Patrick M. Melia Introduction 437 Native Biochar Characteristics and P 439 P Recovery with Biochar 442 Modification of Biochar for Effective P Recovery 443 Cationic Impregnation of Biochar 449 Metal (Hydr)Oxide and Layered Double Hydroxide Biochar Composites 450 Magnetic Biochar Composites 454 Considerations and Outlook 455 References 456 Recovery of Gold and Other Precious Metals by Biosorption 463 Rajmohan Kunju Vaikarar Soundararajan, Sunita Varjani, Ramya Chandrasekaran, and Deepika Kandasamy Introduction 463 Industrial Applications of Gold 464

vii

viii

Contents

17.2.1 17.2.2 17.2.3 17.2.4 17.2.5 17.2.6 17.3 17.3.1 17.3.2 17.3.3 17.4 17.4.1 17.4.2 17.4.2.1 17.4.2.2 17.4.2.3 17.4.2.4 17.4.3 17.5 17.6 17.7 17.8

18

18.1 18.2 18.2.1 18.2.2 18.3 18.4 18.4.1 18.4.2 18.4.3 18.4.4

Gold as a Catalyst 464 Chemical Processing 464 Pollution Control and Monitoring 465 Fuel Cell and Sensors 465 Medical and Biomedical Application 466 Electronics 466 Technologies for Recovery of Gold and Other Precious Metals 467 Importance of Precious Metal Recycling: Gold Use in Electronic Industry 467 Recovery of Nickel and Copper 469 Recovery of Other Metals from Bioderived Materials 470 Biosorption Phenomena 472 Source and Selection of Biosorbents 473 Gold Biosorption by Industrial Biomass 473 Bacteria 474 Algae 474 Fungus 476 Plants 476 Biosorption by Commercial Sorbents 477 Sorption and Desorption Mechanisms of Gold Particles from Biomass 478 Isotherm and Various Models of Sorption Process 479 Future Perspectives 481 Summary 481 References 482 Bioelectrochemical System in Wastewater Treatment: Resource Recovery from Municipal and Industrial Wastewaters 489 Wenshan Guo, Yuanyao Ye, and Huu Hao Ngo Introduction 489 Potential Source 491 Municipal Wastewater 491 Industrial Wastewater 492 Evolution of Bioelectrochemical System 493 Technical Aspects of Bioelectrochemical System for Resource Recovery 496 Energy Recovery 496 Nutrient Recovery 498 Metal Recovery 501 Water Recovery 503

Contents

18.4.5 18.5 18.6 18.7

19

19.1 19.2 19.2.1 19.2.2 19.3 19.3.1 19.3.1.1 19.3.1.2 19.3.1.3 19.3.1.4 19.3.2 19.3.3 19.3.3.1 19.3.3.2 19.3.3.3 19.4 19.4.1 19.4.2 19.5 19.5.1 19.5.2 19.6

20

20.1 20.2

Chemical Recovery 505 Current Application of Bioelectrochemical System for Resource Recovery 507 Current Challenges and Future Perspectives 508 Conclusion 510 References 511 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater 525 Ha Manh Bui, Duy Ngoc Nguyen, Xuan-Thanh Bui, and Hien Quoc Nguyen Introduction 525 Theoretical Basis and Applications of Electron Beam 526 What Is an Electron Beam Process? 526 Applications of EB Processes for Wastewater Treatment 528 Textile Dyes in Textile Wastewater and Their Treatment Technologies 533 Textile Dye 534 Dye Classification 534 Nomenclature of Dyes 535 Dyeing Processes 535 Reactive Dyeing Mechanism 535 Textile Wastewater Characteristic 539 Practical Methods for Treating Textile Wastewater 540 Physicochemical 540 Advanced Oxidation Processes (AOPs) 544 Biological Methods 544 Electron Beam Processes for Textile Wastewater Treatment 544 Lab-Scale Tests 544 Industrial Applications 548 Economic Feasibility and Limitations of EB Processes 549 Economic Feasibility 549 The Limitations of EB Technology for Wastewater Treatment 551 Conclusion 552 References 552 Approaches Toward Resource Recovery from Breeding Wastewater 559 Huu Hao Ngo, Dongle Cheng, and Wenshan Guo Introduction 559 Characteristics of Breeding Wastewater 560

ix

x

Contents

20.2.1 20.2.2 20.2.3 20.3 20.3.1 20.3.2 20.3.3 20.3.4 20.4 20.4.1 20.4.1.1 20.4.1.2 20.4.2 20.4.2.1 20.4.2.2 20.4.2.3 20.4.2.4 20.4.3 20.4.4 20.5 20.6

Livestock Wastewater 560 Poultry Wastewater 561 Aquaculture Wastewater 562 Resources in Breeding Wastewater 563 Water 563 Nutrients 563 Bioenergy 564 Other Bioproducts 564 Approaches for Resource Recovery 565 Biological Approaches 565 Anaerobic Digestion (AD) Processes 565 Dark Fermentation (DF) 569 Physicochemical Approaches 570 Ammonia Stripping 570 Ion-Exchange and Adsorption 573 Chemical precipitation 574 Membrane Filtration 576 Plant-Based Approaches 578 Thermochemical Approaches 583 Current Application and Future Perspectives Conclusion 586 References 587

21

Resources Recovery and Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production 601 Dongyun Du and Jia Li Introduction 601 EMM Production Process and Associated Wastes 602 EMM Production Process 602 EMM Wastewater 602 Electrolytic Manganese Residue (EMR) 604 Manganese Recovery from Manganese-Bearing Wastewater 605 Wastewater Treatment Strategy 605 Onsite CO2 Emission 606 Effect of CO2 Dosage 607 Pilot Treatment System and Its Performance 609 Characteristics of Precipitates Formed 611 Thermal Stability of Formed MnCO3 613 Potential Application in Industry 614

21.1 21.2 21.2.1 21.2.2 21.2.3 21.3 21.3.1 21.3.2 21.3.3 21.3.4 21.3.5 21.3.6 21.3.7

584

Contents

21.4 21.4.1 21.4.2 21.4.2.1 21.4.2.2 21.4.2.3 21.4.2.4 21.4.2.5 21.4.2.6 21.4.3 21.4.3.1 21.4.3.2 21.4.3.3 21.4.3.4 21.4.3.5 21.4.3.6 21.5

22

22.1 22.2 22.3 22.3.1 22.3.2 22.4 22.4.1 22.4.2 22.5 22.5.1 22.5.2 22.6 22.7 22.7.1 22.7.2 22.8

Activation/Recovery of Silicon from EMR: Two Methods Characterization of EMR 615 Chemistry Activation of Silicon 616 Effect of Ball Milling on Silicon Activation 617 Effect of Mass Ratio 619 Effect of Roasting Temperature 619 Effect of Roasting Time on Silicon Activation 620 Morphology Evolution 622 Leaching of Toxic Elements 622 Bioleaching of Si from EMR 624 Bacteria Culture 624 Bioleaching Experiment 624 Growth of Silicon Bacteria Under Influence of Media and pH 625 Optimization of Parameters in Bioleaching 626 Recovery of Heavy Metals in Bioleaching 628 Transformations during Bioleaching Process 629 Conclusion 631 References 632

615

Recovery of Thermal Energy from Wastewater by Heat Pump Technology 635 Long Ni, Tao Song, and Jinyi Tian Introduction 635 Characteristics of Wastewater 636 Description of Wastewater Source Heat Pumps 637 The Principle of Wastewater Source Heat Pumps 637 Wastewater Source Heat Pumps Application 639 Wastewater Heat Exchangers 640 Classification of WWHEs 640 Style of WWHEs 641 Low-Cost Wastewater Decontamination Technologies 646 WW Filtration Technology 647 Antifouling Technology 648 Case Study 654 Challenges and Perspectives 657 Challenges 657 Perspectives 658 Summary and Conclusion 658 References 660

xi

xii

Contents

23

23.1 23.2 23.2.1 23.2.2 23.2.3 23.2.3.1 23.2.3.2 23.2.3.3 23.2.3.4 23.2.3.5 23.2.3.6 23.3 23.3.1 23.3.1.1 23.3.1.2 23.3.1.3 23.3.1.4 23.3.1.5 23.3.1.6 23.3.1.7 23.3.1.8 23.3.1.9 23.3.1.10 23.3.1.11 23.3.1.12 23.3.1.13 23.3.1.14 23.3.1.15 23.3.1.16 23.3.1.17 23.3.1.18 23.3.1.19 23.3.1.20 23.3.2 23.3.2.1 23.3.2.2 23.3.2.3 23.3.2.4

Hydrocyclone-Separation Technologies for Resource Recovery and Reuse 663 Long Ni, Jinyi Tian, and Tao Song Introduction 663 Description of Hydrocyclone-Separation Technologies 663 Working Principle of Hydrocyclones 663 Characteristic of Fluid Flow in Hydrocyclones 665 Parameters Evaluating Performance of Hydrocyclones 666 Separation Efficiency 666 Cut Size 667 Split Ratio 667 Reduced Separation Efficiency 667 Comprehensive Separation Efficiency 667 Total Static Pressure Drop 667 Enhanced-Separation Hydrocyclone Technologies 668 Optimizing Geometric Parameters 668 Cylindrical-Section Diameter 668 Cylindrical-Section Length 668 Inlet Size 668 Inlet-Section Angle 669 Inlet Shape 669 Vortex-Finder Length 669 Vortex-Finder Thickness 669 Vortex-Finder Shape 669 Overflow Diameter 670 Underflow Diameter 670 Ratio of Underflow Diameter to Overflow Diameter 671 Underflow-Pipe Shape 671 Cone Angle 671 Conical-Section Shape 671 Hydrocyclone Inclination Angle 672 Hydrocyclones with Solid Rod 672 Hydrocyclones with Inner Cone 672 Water-Injection Hydrocyclones 673 Hydrocyclones with Reflux Device 673 Multi-Hydrocyclone Arrangements 673 Optimizing Operating Parameters 674 Feed Flow Rate 674 Feed Pressure 674 Feed Density Difference 675 Feed Concentration 675

Contents

23.3.2.5 23.3.2.6 23.3.2.7 23.3.2.8 23.3.3 23.3.3.1 23.3.3.2 23.3.3.3 23.3.3.4 23.3.3.5 23.3.3.6 23.3.3.7 23.3.3.8 23.3.3.9 23.3.3.10 23.4 23.5 23.5.1 23.5.1.1 23.5.1.2 23.5.1.3 23.5.1.4 23.5.2

24 24.1 24.2 24.2.1 24.2.1.1 24.2.1.2 24.2.1.3 24.2.1.4 24.2.2 24.2.2.1 24.2.2.2 24.2.2.3 24.2.3

Feed Particle Size 675 Feed Particle Shape 675 Feed Particle Arrangement 676 Feed Fluid Viscosity/Rheology 676 Optimizing Operating Conditions 677 Electrical Hydrocyclones 677 Electromagnetic Hydrocyclones 677 Permanent Magnetic Hydrocyclones 678 Magnetic Fluids Hydrocyclones 678 Electrochemical Hydrocyclones 678 Flocculant-Assisted Hydrocyclones 679 Hydrocyclone Enhanced by Flotation 679 Hydrocyclones Enhanced by Control Particles 679 Hydrocyclones Enhanced by Adjusting Back Pressure 679 Hydrocyclones Enhanced by Monitoring and Automatic Control 680 Applications of Hydrocyclones in Resource Recovery and Reuse 680 Challenges and Perspectives 683 Challenges 683 Experimental Investigations 683 Simulation Investigations 683 Theoretical Investigations 684 Applications 684 Perspectives 684 References 685 Methane Recovery from Landfills 699 Jyoti K. Chetri and Krishna R. Reddy Introduction 699 Landfill Gas Collection 700 Site Assessment 700 Waste Characteristics 701 Depth of Landfill 701 Moisture 701 Type of Cover and Liner System 701 Estimation of Landfill Gas/Methane Emissions 701 LandGEM Model 701 Theoretical Model 702 Regression Model 703 Estimation of Landfill Gas Collection/Recovery 703

xiii

xiv

Contents

24.2.4 24.2.4.1 24.2.4.2 24.2.4.3 24.2.4.4 24.2.4.5 24.3 24.3.1 24.3.1.1 24.3.1.2 24.3.1.3 24.3.1.4 24.3.2 24.3.2.1 24.3.2.2 24.3.2.3 24.3.2.4 24.3.2.5 24.4 24.4.1 24.4.1.1 24.4.1.2 24.4.1.3 24.4.2 24.4.3 24.4.4 24.5

LFG/Methane Recovery Systems 704 Extraction Wells 704 Wellheads 704 Collection Piping System 706 Condensate Management System 707 Blower 707 Recovery of Landfill Methane 708 Pretreatment of LFG 709 Dehydration 709 Particulates filtration 711 Removal of Hydrogen Sulfide 711 Removal of Carbon Dioxide 711 Utilization of Landfill Methane 713 Direct Use 714 Electricity Generation 714 Combined Heat and Power 715 Alternate Fuels 715 Conversion to Other Forms 716 Challenges in Methane Recovery/Use 717 Technical Challenges 718 Unpredictability of Methane Generation 718 Low CH4 Production 718 Impurities 718 Economic Challenges 718 Social Challenges 719 Regulatory Challenges 719 Conclusions 719 References 720

25

Resource Recovery from Electronic Waste 723 Wenshan Guo, Huu Hao Ngo, Lijuan Deng, Rao Y. Surampalli, and Tian C. Zhang Introduction 723 Generation and Classification of E-Waste 724 Health Impacts of E-Waste 726 Benefits of E-Waste Recycling and Current Status 729 Benefits of E-Waste Recycling 729 Current Status of E-Waste Collection and Recycling 731 Recovery of the Valuable Materials from E-Waste 734 Recovery of Metals from E-Waste 734 Pyrometallurgical Processes 736 Hydrometallurgical Processes 736

25.1 25.2 25.3 25.4 25.4.1 25.4.2 25.5 25.5.1 25.5.1.1 25.5.1.2

Contents

25.5.1.3 25.5.2 25.5.3 25.5.4 25.6

Biohydrometallurgical Processes 738 Recovery of Plastics from E-Waste 740 Recovery of Lithium-Ion Batteries from E-Waste 743 Recycling Waste Solar PV Panels 745 Future Perspectives and Conclusion 747 References 749 Index

755

xv

xix

Preface The population and economic growth and the development of urbanization and industrialization have been increasing the pace of energy consumption and waste production. Currently, human beings are facing the challenges of increasing water, food, and energy demand, as well as the depletion of resources. To tackle these challenges, improve the current environment quality, and build sustainable society, we need a holistic approach toward enhanced water, energy, and waste management practices. Sustainability is defined as meeting the needs of society in ways that can continue indefinitely into the future without damaging or depleting natural resources. In other words, it requires ending the “cradle to grave” cycle of manufactured products by creating products that can be fully reclaimed or reused, and reducing waste and pollution by changing patterns of production and consumption into “cradle to cradle” design. Since organics, nutrients, and thermal heat are three major resources in waste (i.e. wastewater and solid waste), the research and development of resource recovery techniques for waste treatment and management are essential to overcome the challenges of the future and make our lives more sustainable. Water is a unique and essential resource for life on Earth. Water reclamation and reuse is being increasingly emphasized as a strategy for rational use of limited freshwater resource and as a means of safeguarding the deteriorating aquatic environment due to wastewater disposal. Over the past decades, great efforts have been made toward capturing and using the potential resources from wastewater, such as nutrient recovery and biogas production through anaerobic digestion. Since we cannot live a day without producing any waste, it is our responsibility to invent new solutions to reduce massive environmental degradation due to solid waste disposal, as well as to convert our waste into valuable products. To date, waste-to-energy conversion technologies have been considered as a very important strategy for waste management. The purpose of this book is to elucidate basic scientific principles and technological advances of current practical technologies for resource recovery (e.g. energy, nutrient, and material) from waste generated. This book also presents solutions to addressing all concerned problems associated with energy production during waste management and treatment, the feasibility of turning waste as a possible resource, health impacts due to waste disposal, and pollution prevention.

xx

Preface

This 25-chapter book includes most practical and advanced waste treatment and resource recovery technologies (e.g. thermal-chemical treatment, biological treatment, biorefinery, hydrothermal conversion, and biosorption), which aim at recovering energy and useful substances for a large variety of applications. We hope that this book will be of interest to researchers, educators, students, scientists, engineers, government officers, process managers, and practicing professionals. As an excellent state-of-the-art reference material, the book will contain rich knowledge on the principles and provide in-depth understanding and comprehensive information of current advanced technologies, their different practical applications, recent advantages and disadvantages, critical analysis and modeling of the processes, and future perspective toward research directions and development. The editors gratefully acknowledge the hard work and patience of all the authors who have contributed to this book. The views or opinions expressed in each chapter of this book are those of the authors and should not be construed as opinions of the organizations they work for. Wenshan Guo Huu Hao Ngo Rao Y. Surampalli Tian C. Zhang

1

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview Wenshan Guo 1,2 , Huu Hao Ngo 1,2 , Lijuan Deng 1 , Rao Y. Surampalli 3 , and Tian C. Zhang 4 1 Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, New South Wales, Australia 2 Joint Research Centre for Protective Infrastructure Technology and Environmental Green Bioprocess, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, New South Wales Australia and Department of Environmental and Municipal Engineering, Tianjin Chengjian University, Tianjin, China 3 Global Institute for Energy, Environment and Sustainability, Lenexa, KS, USA 4 Department of Civil Engineering, University of Nebraska–Lincoln, Omaha, NE, USA

1.1 Introduction In recent years, resource (i.e. water, raw materials, and nutrients) and energy have been subject to high pressure caused by climate change, demographic and land use changes, increase in world population, and high standards of living together with urbanization [1]. Moreover, traditional water management (i.e. take–make–waste approach) and waste management (i.e. waste dumping in landfill sites) techniques have aggravated resource scarcity and environmental, social, and economic problems [2]. Additionally, rare and precious resources (i.e. indium, silver) will be used up by traditional supplies of these elements [3]. It is predicted that the annual energy demand will reach around 23 TW worldwide by 2050 [4]. Therefore, resource and energy recovery and reuse should be realized to alleviate resource scarcity and environmental degradation, and enable economic benefits. Resource recovery can be achieved from two sources: water and waste. Current studies have focused on the recovery of heat, organic carbon, and nutrients from various types of wastewaters. The heat is recovered from household water (i.e. shower water), sewer, or wastewater treatment plants by a heat recovery system, which mainly contains a heat exchanger and a heat pump [5]. Nutrient recovery from wastewater, especially phosphorus recovery, is commonly achieved by struvite formation. Moreover, energy, nutrients, and materials can also be recovered from different kinds of wastes; one such example is the recovery of renewable energy from waste in the form of value-added products (e.g. methane containing biogas and ethanol), phosphorus from animal manures, food waste and sewage sludge, and materials in terms of heavy metals and scarce and valuable metals from mining waste, municipal and industrial waste, and e-waste [3, 6, 7]. Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

2

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview

This chapter gives a brief introduction, key drivers, current status, and future perspectives of resource and energy recovery and reuse. The chapter is divided into four sections: background, the current status for waste generation and recovery, the research needs of resource and energy recovery and reuse, and a brief review on the core ideas and key researches for each book chapter.

1.2 Background 1.2.1

Hierarchy of Resource Use

For effective resource management, an alternative “hierarchy of resource use” (HRU) has been proposed by Gharfalkar et al. [8] to clarify “prevention, preparing for re-use, re-cycling, other recovery and disposal” in the latest version of European Commission’s Waste Framework Directive 2008/98/EC and consider the “waste” as “resource.” Figure 1.1 displays the proposed alternative HRU. HRU consists of five sections as follows: (i) Replacement: rethinking, reinventing, or redesigning to remove or replace existing demand with demand for environmentally benign materials or objects and/or replace nonrenewable resources with renewable resources. (ii) Reduction: reinventing or redesigning to “reduce” use of resources, including reduction in consumption of resources, waste generation and resultant environmental degradation. (iii) Recovery involves preparing materials for reuse (a reusable material or an object can be reused by the preparing operation); reuse (reuse without any further operation, repair and reuse, refurbish and reuse, recondition on and reuse, remanufacture and reuse, any other operation and reuse); reprocessing (recycle, downcycle, and upcycle [“waste” is reprocessed into materials with the same, lower, and higher purpose/value than the original, respectively)]), and other recovery (energy recovery and/or recovery of materials to be used as fuels or for backfilling operations). (iv) Rectification (a waste treatment operation before disposal). (v) Return (disposal).

1.2.2 Analyzing the Needs for Resource and Energy Recovery and Reuse The key drivers for resource and energy recovery and reuse mainly comprise population growth, environmental impacts, resource scarcity, and economic aspects. Figure 1.2 shows the interaction among the four key drivers for resource and energy recovery and reuse. 1.2.2.1 Population Growth

In 2019, the global world population reached 7.7 billion and it has been predicted that the global population will increase dramatically up to 9.7 billion in 2050 and 10.9 billion in 2100 [9]. Less developed countries play a key role in urban growth,

1.2 Background

1 Replacement Non-waste

2 Reduction

Waste

3.1 Preparing for reuse 3.2 Reuse

3 ry ve co Re

3.3 Processing 3.4 Other recovery 4 Rectification 5 Return

Figure 1.1 Proposed alternative “hierarchy of resource use” (reverse triangle) (Source: Modified from Gharfalkar et al. [8]).

Population growth

Resource consumption

Energy consumption

Environment impact

Resource scarcity Economic aspect

Figure 1.2 reuse.

Interaction among the four key drivers for resource and energy recovery and

3

4

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview

which contributes greatly to population growth. The world energy demand mainly comes from urban demands, with more than 2/3 of the world’s energy expected to be consumed by cities from 2006 to 2030 [10]. World energy consumption is estimated to increase by approximately 23% from 2020 to 2040, reaching 820 quadrillion Btu [11]. Rapid urbanization also induces other striking problems, including land degradation, desertification, deforestation, resource (e.g. water, materials) depletion and pollution, and loss of biodiversity [12]. 1.2.2.2 Resource Scarcity

Natural resources are commonly classified into two types, namely renewable (water, land, forest, fish, etc.) and depletable (minerals, metals, oil, etc.) resources [13]. Resource scarcity is caused by the increase in population growth, economic level, standard of living, and the limited supply of resources. Although it is possible to obtain more and more energy from renewable and nuclear energy sources, the amount of generated energy is still lower than the increasing energy demands. It has been pointed out that the higher energy demand in Asia significantly induces CO2 emissions by combusting carbon-based energy sources (gas, oil, and coal), which annually increase by 2.3, 2.1, and 1.9% in India, China, and rest of Asia, respectively. It is also estimated that a large fraction of energy (> 76%) will originate from carbon-based source in 2040, which increases the diminishing rate of primary energy resources [14–16]. 1.2.2.3 Environmental Impacts

The excessive discharge of nutrients into water bodies causes algal bloom and overgrowth of plants and “dead zones” in coastal marine ecosystems [17]. Nutrients exported from urbanized river basin in 2050 are projected to be around five times the level in 2000; these mainly come from sewage, industries, and urban agriculture [18]. Drinking water, soil, fodder and food are contaminated by heavy metals from industrial waste. Furthermore, contaminated sites being important sources of pollution can lead to ecotoxicological effects on terrestrial and aquatic ecosystems (e.g. increased cell size, shortened life span, and decreased body weight) [19, 20]. Resource consumption (e.g. fossil fuel for energy) together with increased life quality and world population, as well as industrialization of developing nations, exerts adverse impacts on the environment. 1.2.2.4 Economical Aspect

Zaman [21] pointed out that per capita gross domestic product (GDP/capita/year) is positively correlated with per capita waste generation (Table 1.1). It was reported that average waste generation rates in high-income (HIC, GDP = more than $12275/cap), upper middle-income (UMIC, GDP = $3976–$12275/cap), lower middle-income (LMIC, GDP = $1006–$3975/cap), and low-income (LIC, GDP = less than $1005) countries were 2.1, 1.2, 0.79, and 0.6 kg/cap/day, respectively [23]. Although 84% of waste generated is collected in the world, only 15% is recycled. In the future, waste generation would increase because of constant economic growth,

1.3 Current Status of Resource Recovery and Reuse

Table 1.1 streams.

Total nitrogen (TN) and total phosphorus (TP) content of different waste

Wastewater

Description

TN (mg/L)

TP (mg/L)

Municipal wastewater

Sewage

15–90

5–20

Animal wastewater

Dairy

185–2636

30–727

Poultry

802–1825

50–446

Swine

1110a)–3213

310–987

Beef feedlot

63–4165

14–1195

Textile

21–57a)

1.0–9.7b)

Winery

110a)

52

Tannery

273a)

21b)

Paper mill

1.1–10.9

0.6–5.8

Industrial wastewater

Anaerobic digestion effluent

Olive mill

532

182

Dairy manure

125–3456

18–250

Poultry manure

1380–1580

370–382

Sewage sludge

427–467

134–321

Food waste and dairy manure

1640–1885a)

296–302

a) Total Kjeldahl nitrogen (TKN). b) Total orthophosphates (PO4 3 –P). Source: Cai et al. (2013) [22].

especially in the developing countries. Moreover, HIC could gain remarkable economic benefits from resource recovery and energy savings compared to other income groups. In the United States, Japan, and the European Union, the supply of raw materials influences the economy (e.g. jobs) [1].

1.3 Current Status of Resource Recovery and Reuse 1.3.1

Wastewater

Water contamination is now a serious issue due to surface water and ground water being heavily polluted by industrial and municipal wastewater, agricultural activities, and household wastes, especially in developing countries. For example, the amount of wastewater discharged in China was 68.5 billion tons in 2012 and this contained 24.2 million tons (Mt) of chemical oxygen demand (COD) and 2.5 Mt of ammonia nitrogen emission [24]. However, wastewater can be considered as a collection of resources that can be recovered, including energy, organic carbon, nutrients, and clean water. Thus, many efforts have been made to recover available resources from wastewater, mainly focusing on the recovery of nutrients, organic carbon, and heat.

5

6

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview

1.3.1.1 Nutrient Recovery

The composition of wastewater is heavily dependent on its sources. Table 1.1 shows the total nitrogen (TN) and total phosphorus (TP) content in different wastewaters. The key challenge for domestic wastewater treatment is the recovery of energy and nutrients (e.g. phosphorus, nitrogen, and potassium). There are two alternative wastewater treatment platforms to overcome these difficulties [25, 26]: 1. Low energy mainline (LEM), which enables net energy recovery through anaerobic processes (e.g. upflow anaerobic sludge blanket reactor and anaerobic membrane bioreactor) and phosphorus recovery through advanced ion adsorption techniques. However, although LEM removes nitrogen by main line anaerobic deammonification (Anammox) at low energy unit (20% of power consumption of conventional process), it cannot recover nitrogen and other elements (i.e. potassium). 2. Partition–release–recover (PRR), in which nutrients and organics are concentrated through assimilation into solids using heterotrophic or phototrophic microbes. Afterward, nutrients (nitrogen and potassium) can be recovered from the digestate subsequent anaerobic digestion by electrochemical techniques, struvite precipitation, etc. Mainline anaerobic processes are favorable to be utilized in the shorter term for phosphorus and energy recovery due to its lower energy costs and it being closer to market status. Recovery of nitrogen and potassium can be accomplished by PRR for higher strength wastewaters, or higher COD:N ratio of wastewaters [26]. In general, phosphorus can also be recovered as calcium phosphate or struvite by feeding the phosphorus-rich wastewater into a precipitation/crystallization tank with addition of calcium or magnesium salts in a mixed or fluidized state. Municipal wastewater, which is dilute in nature, contains large amounts of greywater, flush water, and even storm water. Hence, the recovery of nitrogen and phosphorus from municipal wastewater is more expensive in view of economic and energy aspects. As wastewaters from agricultural diary, brewery, and starch-manufacturing industries contain lower phosphorus levels, the recovery of phosphorus as struvite is expensive from these industrial wastewaters [27]. 1.3.1.2 Organic Carbon Recovery

COD concentration of the influent significantly affects organic energy in the form of methane. Hao et al. [28] pointed out that 53% of total energy consumption could be offset by organic energy when influent COD was 400 mg/L in a case study of a typical municipal wastewater treatment plant (WWTP) in China . Another case study was conducted by Khiewwijit et al. [29], and the proposed treatment contains bioflocculation, partial nitritation/Anammox, P recovery process, anaerobic sludge digestion, and combined heat and power. Compared to conventional activated sludge systems (no net energy yield, −0.08 kWh/m3 ), the results revealed the possibility of improving the net energy yield up to 0.24 kWh/m3 and reducing carbon emissions by 35%. Moreover, when COD increased by 20%, 26% increase in energy yield could be achieved.

1.3 Current Status of Resource Recovery and Reuse

As a renewable carbon resource, the recovery of volatile fatty acid (VFA) from different waste streams is attracting more and more attention for producing various valuable products, such as biogas, biodiesel, bioplastics (thermoplastic biodegradable polyesters), and biohydrogen. The amount of VFA generated from a given waste stream is determined by the acidification degree, which is related to the fraction of readily fermentable organics. Higher degree of acidification results in higher VFA generation. The composition of VFA (e.g. acetic acid, butyric acid, propionic acid, and others) is related to the characteristics of the organic matter contents in wastewater. The key VFA recovery methods include gas stripping with absorption, adsorption, electrodialysis, solvent extraction, nanofiltration, reverse osmosis, and membrane contractors [30]. 1.3.1.3 Heat Recovery

Thermal energy, being another large reserve of energy, can be recovered from different types of wastewaters by heat exchangers or heat pumps. It has been indicated that 1.16 kWh of thermal energy could be released by cooling one cubic meter of wastewater by 1 ∘ C for specific heat of water. A water source heat pump (WSHP) can be used to capture the heat to effectively offset the difference between the total energy consumed and energy produced by anaerobic digestion. The amount of energy obtained from wastewater is affected by pump efficiency (i.e. coefficient of performance), the flow through the pump, the change of temperature in wastewater when passing through the pump, the distance between the location to be heated and the pump, and the total area available to be heated [31, 32]. Hao et al. [28] evaluated the energy recovered from a typical municipal WWTP in China and the results suggested that considering all relevant factors (i.e. boiler heating efficiency, average power generation efficiency, heat supply grid efficiency, and power grid loss), a net energy equivalent to 0.26 kWh/m3 could be supplied by WSHP through decreasing the temperature of 1 m3 of wastewater effluent by 1 ∘ C.

1.3.2

Waste

The world generates 2.01 billion tonnes of municipal solid waste annually, with at least 33% of waste not managed in an environmentally safe manner. Worldwide, the average amount of waste generated per person per day ranges from 0.11 to 4.54 kg with an average of 0.74 kg. High-income countries generate about 34% (683 million tonnes) of the world’s waste, but they only account for 16% of the world’s population. The global waste generated is expected to grow to 3.40 billion tonnes by 2050 [33]. Depending on the economic situation of different countries, waste can be classified into different categories, including municipal solid waste (e.g. food waste, and sewage sludge), industrial waste (e.g. mining waste and construction and demolition waste), electronic waste, and hazardous waste. The disposal of waste into the environment not only leads to contamination of waterways and soil due to non-point source pollution, but also causes air pollution. The total amount of global food waste is about 1.3 billion tonnes every year, with no remarkable difference between the developed and developing countries (670 and

7

8

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview

630 million tonnes, respectively). China and India are the major contributors for FW generation from developing countries, with values of 195 and 72 million tons, respectively. In developing countries, food waste accounts for 50–55% of municipal solid waste (MSW). The worldwide urban food waste will increase by 44% between 2005 and 2025 [34]. The variable chemical composition in food waste is due to different origin of production, which contains carbohydrates, lipids, and proteins or high concentration of these constituents [35]. Current studies have explored the recovery of value-added products (i.e. proteins, polysaccharides, flavor compounds, fibers, and phytochemicals that can be used as nutritionally and pharmacologically functional ingredients) from food waste. The recovered biomolecules and by-products can be employed for food and food processing uses (e.g. as gelling agents in sweets, food additives), as well as medicinal and pharmaceutical uses (e.g. as appetite modulators). The extraction techniques include solid–liquid extraction, soxhlet extraction, ultrasound-assisted extraction, microwave-assisted extraction, pressurized fluid extraction, and supercritical fluid extraction, pulsed electric field extraction, and enzyme-assisted extraction [36]. Biosolids generated by removing the part of chemical oxygen demand during biological wastewater treatment process are often referred to as sewage sludge. It was reported that half of wastewater phosphorus (2–8 mg/L) is integrated into biosolids. The enhanced biological phosphorus removal process prompts the accumulation of phosphorus in sludge from 0.02 to 0.06–0.15 mg/g VSS (volatile suspended solids). A small proportion of the activated sludge dry mass contains 24–67 g N/kg dry mass of nitrogen and the remaining N in the form of N2 gas. Potassium accounted for 0.5–0.7% K2 O weight of dry solids of sewage sludge. The complete recovery and reuse of nutrients from 30 million tons of sludge generated annually in the world could meet the demands of 5% of phosphorus, 1.7% of nitrogen, and 0.64% of potassium [37, 38]. Mining involves various activities such as mine development, mineral beneficiation, metal extraction, smelting, refining, reclamation, and remediation. In the mining and metal extraction industry, different types of wastes are produced (e.g. waste rock, mineral beneficial tailings, metallurgical slags, wastewater, and water treatment sludge). In fact, these waste streams can be considered as secondary sources of valuable minerals and metals. The recovery of value metals from ferrous metallurgical dusts that are generated in the iron and steel manufacturing processes can be realized through pyrometallurgical and hydrometallurgical processes. Compared to the pyrometallurgical processes, the hydrometallurgical processes have more advantages of higher flexibility of operation, the required economies of scale, lower capital costs, and the minimum environmental challenges related to flue gases, dusts, and noise. Nevertheless, careful management of water, wastewater, and process solutions is required for its technical and economic feasibility. Moreover, the complex characteristics of the dust materials also restrict the wide application of the hydrometallurgical processes [39]. The high consumption of electronic equipment and their short lifespan prompt an increase in manufacture of electronic and electrical equipment and the production

1.4 Research Needs

of electronic waste (e-waste) [40]. E-waste generally comes from individual households, small business sections, original manufacturing sectors, large corporations, and institutions and governmental sectors. In recent years, approximately 20–50 million tons of e-waste are generated per year in the world with an annual growth rate of up to 5%, while around 12 million tons of them are disposed [41]. Currently, valuable metals production system is subjected to some challenges owing to the scarcity of primary resources and earth’s intrinsic limitations [42]. Therefore, as the key component of electronic devices, the waste printed circuit boards (WPCB) contain higher concentration of several precious metals than their corresponding ores. For example, concentrations of silver, gold, and palladium in minerals (< 10 g/tonne) are lower than those found in computer printed circuit boards (typically average of 1000, 250, and 110 g/tonne PCB, respectively). Besides, it is estimated that the value of copper and precious metals in waste mobile phones and WPCB accounted for up to 80% of the total value [43]. Some metals (e.g. Cu, Pb, Fe, Au, and Hg) have been successfully recovered from PCB. Spent batteries and used machines and instruments are also used for recovery of metals (e.g. Co, Li, Zn, Mn, Pb, Ni, Au, Pd, and Pt) and rare earths [41].

1.4 Research Needs 1.4.1

Development of Novel Technologies

New processes and improvements in the existing technologies for resource recovery should be carried out due to the complex characteristics of resources. Moreover, hybrid systems incorporating different kinds of technologies are suggested to be developed to enhance resource recovery and reuse from wastewater and waste, such as bioelectrochemical systems combined with current wastewater treatment plant, specific bacteria (e.g. hydrogen-oxidizing bacteria) coupled with other systems (e.g. electrochemical or bioelectrochemical system) [4], and integration of membrane-based process with anaerobic treatment. Some aspects that should be considered when developing novel technologies include: ● ● ● ●

● ● ●

Less complicated operation. High efficiency of resource recovery. Efficient and economic production of value-added products. Generation of few or several high-volume liquid transportation fuel to fulfill some national requirements of energy [44]. Production of electricity and process heat available for its own use [44]. Reduction of sludge production during wastewater treatment process. Less greenhouse gas emissions.

1.4.2

Social and Economic Feasibility of Resource Recovery and Reuse

Economic sustainability should be considered when implementing large-scale operations. Based on the principles of efficient resource recovery and economic

9

10

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview

operation, substantial research should be performed to scale up current and emerging techniques for resource recovery from waste and wastewater. Detailed techno-economic analysis of the novel technologies should be conducted, such as energy demand, operating and maintenance costs, reproducibility of lab-scale results in large scale, and market demands. Social feasibility is a key point as high acceptance of the technology could ensure its market place and wide application.

1.4.3 Development of Internationally Coordinated Framework and Strategy The cooperation among governments, researchers , and companies in developing and developed countries with regard to coordinated framework and strategy is required to ensure a global sustainable market for the application of resource recovery technologies. Rigorous legislative frameworks should also be given for wastewater and waste discharges and resource recovery to reduce environmental contamination and maximize the recovery and reuse of resource.

1.5 Book Overview The primary objective of this 25-chapter book is to elucidate basic scientific principles and technological advances of current technologies for resource and energy recovery and reuse. Food waste, an abundant bioresource, can be a potential feed to establish a sustainable supply chain of high-energy density biofuels with low carbon footprint. The urgent need for clean fuels to curb the alarming greenhouse gas emissions and the prerequisite to meet the growing energy demand have propelled the development of the hydrothermal liquefaction (HTL) technology. HTL-driven wet food waste valorization has garnered global attention for resource recovery in the form of bio-oil/bio-crude, bio-char, and other gaseous products that can be upgraded to useful platform chemicals and biofuels. Chapter 2 showcases HTL of food waste as a commercially feasible energy recovery alternative that can be highly beneficial to conventional fossil-driven sectors. In the context of ambitions for a circular economy, there is need for fundamental transition in resource recovery practices. However, while the sustainability transitions literature focuses on the transition toward a more sustainable future, practically policymakers must operationalize a goal for moving away from an unsustainable present. Historically, the United Kingdom has relied upon disposal of waste to landfill rather than viewing waste as a potential material resource and encouraging material recycling. Yet from a low base, a major successful change was made at the UK local authority level in the collection and sorting processes required to effect this sustainable transition between the then government’s waste strategy appearing in 2000 and the defeat of the government in 2010. Chapter 3 discusses this change from the perspective of local authorities, who have a pivotal role in household waste management, taking a long view of the 2000s to consider

1.5 Book Overview

the major changes in environmental policy and local government organization between 1979 and 2014. The chapter also draws on interview data from local authority waste managers in the Yorkshire and Humberside region as well as national policymakers and regulatory bodies, in addition to local authority and national waste statistics, government and industry documents. Results indicate that the historical and geographical contexts are important for understanding the ability of the local authorities to respond. It is concluded that the “landscape” level is more geographically constructed than recognized in the literature and that the regime scale, rather than being a fixture that needs to be changed, is changing continuously. As HTL has been widely applied to obtain bioenergy and high-value chemicals from biomass at moderate to high temperature (200–550 ∘ C) and pressure (5–25 MPa), Chapter 4 further focuses on the HTL conversion properties of lignocellulosic biomass of agricultural and forestry wastes. The history and development of HTL technology for lignocellulosic biomass are briefly introduced. The research status in HTL of agricultural and forestry wastes are critically reviewed, and the effects of HTL conditions on bio-oil yield and the decomposition mechanisms are summarized. The limitations of HTL of agricultural and forestry wastes are also addressed, and future research priorities are proposed. The nutrients present in human faeces, such as nitrogen, phosphorous, and potassium can be recycled if a linear, non-recycling open-ended system is discarded. A better sanitation approach would be to focus on recovery of nutrients present in human excreta and considering sanitation systems as collection and processing units of those nutrients. Chapter 5 introduces the available technologies focusing on resource recovery from human waste. The potential impacts of these systems are discussed. It is concluded that ecological sanitation solutions are much cheaper and have much lower environmental impacts although the acidification potential remains a cause of concern and requires further research. These solutions are also feasible in developing countries and require low energy and capital investments. Hence, regions that do not have adequate wastewater treatment facilities from the very beginning should concentrate on implementing technologies that are resource recovery oriented. Chapter 6 systematically outlines the current status of livestock manure production, transportation, and recycling, and proposes a future research perspective for its management from a global perspective. With the rapid development of large-size livestock farming and increasing demand of animal products, livestock and poultry farming methods have been transformed from traditional decentralized and extensive to large-scale, intensive, and specialized, as, an inevitable by-product of livestock and poultry farming, a huge quantity of animal manure (such as: pig manure, chicken manure, cow manure, duck manure, and so on) is generated all the time in each country and the resulting livestock manure has also been concentrated into the local environment, causing serious damage to the surrounding soil, groundwater, and atmosphere. Therefore, it is of great environmental and economic significance to recycle livestock and poultry manure upon a commercial model. The production, collection, transportation, and treatment of livestock and poultry manure are

11

12

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview

all currently in urgent need of research and innovation. Current treatments and technologies for livestock and poultry manure include anaerobic digestion and aerobic composting, which are all through the use different types of microorganisms under an aerobic or anaerobic environment to humify manure or convert them into natural gas, thereby achieving resource utilization. Due to the different national and economic conditions of developed and developing countries, there are some differences existing in the recycling and disposal of livestock manure. Using the theory of circular economy to recycle livestock manure has become an innovative and hot research topic, many developed countries have tried to apply the business model of circular economy to the recycling of livestock and poultry manure as well as follow the 3R principle for resource utilization. This chapter also prospects the utilization of livestock manure resources; the rational use of the theory and business model of circular economy to recycle livestock manure is an effective measure for the treatment of livestock manure in each country in the future. There are growing concerns of fossil fuel consumption and its critical negative impact on our environment, which have driven the research and development of sustainable biofuel production in the last decades. Microalgae and thraustochytrids are promising candidates for biofuel production because of their great biomass growth rate and lipid accumulation potential. However, there are still technical and economic barriers for the commercialization of microalgae/thraustochytrids-based biofuel production. Some microalgae and thraustochytrid species have shown attractive nutraceutical application potentials, which may generate great economic values to offset the biofuel production cost. Integration of the manufacture of biofuel and nutraceutical products from microalgae/thraustochytrids could probably make the microalgae/thraustochytrids-based biofuel production economically feasible. Chapter 7 presents a review of the recent advances in the production of biofuel and nutraceutical products such as docosahexaenoic acid (DHA) and eicosapentaenoic acid (EPA) from microalgae and thraustochytrids, including technical, economic, and environmental challenges as well as future perspectives. Chapter 8 identifies issues related to algal energy and bioproduct formation with a perspective on market opportunities for different bio-products. Though the dilemma of cultivating algae stems around overwhelming capital and operational costs, increasing number of published studies in the past exemplify the exponential growth of microalgae bioenergy research. Currently, the main purpose of biorefinery is to integrate the production of commodity chemicals and high-value products along with fuel and energy generation, while optimizing resources and minimizing waste. Water recyclability, co-production of high-value products along with energy generation and under highly controlled process parametric conditions with specific market value analysis is the key to make the algae biorefinery logistics highly industry feasible. Furthermore, the development of cheaper novel downstream processing options would render the algae cultivation process acceptable, while helping in managing waste and simultaneously reduce carbon footprints. The amount of sludge produced by sewage treatment is increasing rapidly, and the resource utilization of sewage sludge has been paid much attention in recent decades. Chapter 9 introduces the preparation, characteristics, and application of

1.5 Book Overview

sludge-derived carbon for wastewater biosorption in view of influencing parameters, such as activating agent, contact time, initial pollutant concentrations and modification method. The biosorption performance and mechanism of granular sludge for direct organic pollution and heavy metal treatment are also summarized. Moreover, the role of extracellular polymeric substances (EPS) during granular sludge biosorption process and the interaction between EPS and pollutions by using various spectroscopic approaches are systematically reviewed. Additionally, due to the abundant functional groups and binding sites, the environmental applications of EPS as effective bioflocculant and adsorbent in wastewater treatment are reviewed. As sewage sludge posed a serious environmental challenge, Chapter 10 reviews thermal-chemical processes, which have been considered as an emerging technology for sewage sludge management aiming for concurrent volume reduction, pathogens destruction, and energy and resource recovery. This chapter offers a comprehensive picture of the state of the art of energy and resource recovery from sewage sludge through thermal-chemical treatment, with the focus on process feasibility, cost, limitations, challenges, and solutions forward. The design, optimization, and operation of the thermal-chemical treatment processes of sewage sludge for long-term environmental sustainability and economic viability, with maximized energy and resources recovery, have also been addressed. Chapter 11 introduces the treatment of sewage sludge through anaerobic sludge digestion. The state-of-the-art technologies for enhancing methane and hydrogen productions from sewage sludge are also elucidated, including physical, chemical, and biological pretreatment. Emphasis was put on their effect on methane and hydrogen production performance, with an increase of 10–340% in methane production and an increase of 20–1300% in hydrogen production. In general, thermal pretreatment, free nitrous acid pretreatment, free ammonia pretreatment, and temperature-phased anaerobic digestion show advantages over the other pretreatment technologies. In addition, ultrasonic pretreatment (< 4400 kJ/kg total solids) will also be promising if pathogen destruction is not a main concern. In the future, various pretreatment technologies should be implemented to the same sludge source in order to avoid the bias imposed by the different sludge sources. Phosphorus (P) is a main water pollutant responsible for eutrophication and related surface water quality problems. On the other hand, P is also a limited and non-substitutable nutrient for agricultural production. The rapid depletion of natural P reserves will greatly threaten global food security. It is therefore urgent to recover and recycle P from all possible sources, including wastewater and sludge. In Chapter 12, different technologies for effective P recovery in connection to chemical P removal are introduced, including chemically enhanced primary sedimentation and chemically enhanced membrane bioreactor. The very different solubilities of Fe(III)–P and Fe(II)–P complexes and microbial transformation of Fe(III) to Fe(II) are utilized for P removal and recovery. The main pathways for P removal and recovery in chemical precipitation and acidogenic fermentation are revealed by X-ray absorption near edge structure (XANES) spectroscopy. Although P is a nonrenewable resource, excess phosphorus inevitably enters the water environment and causes eutrophication. Compared with other technologies

13

14

1 Resource Recovery and Reuse for Sustainable Future Introduction and Overview

such as chemical precipitation and biological processes, adsorption is a favorable option for its simplicity in design, easy operation, and minimal waste production. Although various adsorbents have been developed, some drawbacks limit their application in the field. Thus, Chapter 13 introduces the novel magnetic iron-based oxide materials developed in the form of core–shell structure and composites (i.e. ZrO2 @Fe3 O4 , La(OH)3 /Fe3 O4 and Fe0 /Fe3 O4 ) for phosphorus removal and recovery, which have strong selectivity, high adsorption capacity, favorable reusability, and present easy material separation by magnetic field. The results show a high potential of using magnetic iron-based oxide materials for possible phosphorus removal and recovery applications. Forward osmosis (FO), considered as a promising separation process for nutrient enrichment in wastewater, is attracting increasing interest in integration with chemical precipitation and other technologies for recovering nutrients in wastewater treatment. Chapter 14 highlights nutrients recovery processes via FO-based systems based on mechanisms and influencing factors. In addition, the key challenges related to the recovery systems discussed and some approaches are proposed to resolve these challenges. Future roadmap for future research and development on the nutrients recovery using FO-based systems are identified. Compared to aerobic FO-based systems, anaerobic FO-based processes need more investigations of their integrations’ efficiency in the context of nutrient recovery from wastewater. Emphasis is given to carry out more economic assessment and pilot- and plant-scale evolutions of the recovery systems, which makes the nutrients recovery via FO-based technologies more sustainable in wastewater treatment. Adsorption using low-cost adsorbent seems to be a promising technique for nutrient recovery. For the sustainability of the ecosystem and environmental protection, removal and recovery of nutrients from wastewater along with regeneration and reuse of spent adsorbent are essential. Many investigations have been carried out for the removal of nutrients from single-component systems whereas very few investigations have examined and modeled the nutrients removal in multicomponent adsorption systems. Chapter 15 covers the basic principles and potential application of low-cost adsorbent for removal and recovery of nutrients (PO4 3− and NO3 − ) from single-component and multicomponent adsorption systems. Furthermore, this chapter provides an overview of adsorption capability of different low-cost adsorbents for PO4 3− and NO3 − removal, factors influencing adsorption, kinetic and equilibrium modeling, mechanism of nutrient adsorption, and management of spent adsorbent. Finally, the application of modified Langmuir model for binary adsorption system (simultaneous removal of PO4 3− and NO3 − ) is discussed. Since P must be removed from wastewater, commonly to 60 ∘ C. The resulting bio-oil/biocrude was recovered from the reactor walls and the solid residue using chloroform, which was later evaporated under vacuum. The yield of high-energy density HTL products, i.e. bio-crude, solid residue, and filtrate was determined. Al-hassan and Kumar reported that HTL of solid food waste using lab-scale batch reactor yielded >80% bio-oil when treated at 225 ∘ C for 30 minutes with a food waste: solvent ratio of 1 : 10 and a catalyst ratio of 4 wt% [53]. A continuous HTL pilot-scale reactor (Figure 2.2) that consisted of a controlled feed inlet system attached to a heat exchanging system that is in turn connected to a preheating unit was reported [54]. This is followed by a plug-flow HTL reactor, which is ideally a tubular system with a constant internal diameter of 14.2 mm and length of 140 m. The HTL reactor in this case was appropriately fabricated to sustain the prevailing temperature within the reactor. The area within the reactor that is exposed to above ∼80 ∘ C is made from high nickel alloy (UNS N06025) while the parts revealed to lower temperatures are fabricated using stainless steel (SS-316). Gases

Feed flow Recirculation flow Product steam flow

Hydro cyclone

Slurry from pretreatment

HTL reactor Gravimetric separator

Feed hoper Trim heater Aqueous phase Biocrude Screw pump

Heat exchanger

Oscillator 1

Take-off system

Feed pump Oscillator 2 Cooler

Figure 2.2 Process flow in a continuous hydrothermal liquefaction (HTL) reactor. (Source: Modified from Raikova et al. [52].)

2.2 Significance of Hydrothermal Liquefaction of Food Waste

The other units connecting to HTL reactor are oscillators, take-off unit, and product collection zone. The process flow occurring in the entire HTL unit (Figure 2.2) is controlled through a programmable controller system. The feed inlet system has a feed hopper and two associated feed pumps. One pump functions to recirculate the feed and maintain homogeneous consistency and avoid feed settling. It adequately stirs the contents and provides a constant feed flow to the next feed pump under relatively higher pressure (approximately 2 bar). The next feed pump in series operates under markedly high pressure (220 bar) allowing for a flow rate of 60 L/h into the reactor. The heat exchanging unit attached to the second feed pump is a dual Inconel pipe counter-current heat exchanger that is designed to possess convolutions creating hotter and colder sides of the exchanger. Several temperature-regulating thermocouple systems are deployed around the heat exchanging unit. The counter-current heat exchanger works by effectively preheating the fresh feed using the heat from the counter flowing product stream arising from the reactor. The feed stream is again heated to the desired reaction temperature using a preheater fitted in continuity with the heat exchanger. The perfectly heated feed slurry is then allowed into the main HTL reactor where the actual conversion of feed to high-value chemicals occurs. SS heavily insulated casing is done to the hotter parts of the entire HTL unit to avoid heat loss. The feed stream after the reaction flows through the hotter side of the heat exchanger and transfers heat to the incoming cold fresh feed and consequently turns into cold product stream that flows out into the take-off system. The hydraulic oscillation system associated with the HTL is the crucial unit, which is fitted to improve mixing and heat transfer within the operating system through the turbulence developed through the two attached pistons. Turbulence in the flow is developed through the first piston placed immediately after the main feed pump and before the inlet of cooling zone of the implanted heat exchanger. The second piston lies amid the exit trail of the cooler and prior to the take-off system. These pistons continuously move in alternation such that one side is exposed to HTL contents at set reaction pressure and the other side is connected to a high pressure hydraulic oil circuit system forcing the contents to move forward and backward during their net forward movement. This increases the local velocity of the HTL contents without change in overall residence time. The product stream then enters the take-off system, which also consists of two pistons working alternatively similar to the oscillators. The product flow is channeled toward either of the pistons using attached values. The targeted piston collects the contents of HTL product stream at the same prevailing pressure and slowly retracts and disburses the contents to the HTL outlet maintained at atmospheric pressure within product collection funnel. Likewise, these pistons move in alternation to enable continuous flow of reactor contents. The product stream emptied into the product collection zone is directed first into a hydro-cyclone that fundamentally separates gaseous products from that of liquid products. This is followed by a separating funnel/gravimetric separator where the components of liquid phase, i.e. biocrude, are separated from the nonvolatile aqueous phase. Using this reactor, the biocrude yield from feed having 16% (w/w) dry matter was reported to be 26% (w/w), 33% (w/w), and 25% (w/w) for miscanthus, spirulina, and sewage sludge, respectively.

29

30

2 Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy

2.2.2

Isothermal HTL and Fast HTL

Isothermal HTL is a conventional form of HTL that is conducted at a constant temperature for several minutes. This strategy has been was found to be effective for a plethora of residues arising from agriculture as well as food processing [55]. The existing pool of knowledge on HTL mostly encompasses isothermal HTL of feed that spans over tens of minutes. On the contrary, fast HTL also referred as rapid heating HTL is a recent development where the hydrothermal liquefaction is conducted at non-isothermal temperature conditions for relatively shorter residence time (tens of seconds) than the isothermal counterpart. It is generally considered that for processes with shorter reaction time, smaller sized reactors can suffice, thus reducing both capital (CAPEX) and operating (OPEX) expenditure [56]. These two strategies are crucial HTL types that influence the concentration of metals and inorganic species present in the product stream [57]. A comparison was made between isothermal HTL of sewage sludge conducted at 673 K for 60 minutes and fast HTL carried out at 773 K for 1 minute [58]. The effect of several influencing parameters such as moisture content of feed, overall water loading in bioreactor, and type of biocrude recovery solvent, was studied in detail. It was found that water loading, which majorly regulates the pressure within the reactor, has minor influence on yield of the product. On the contrary, moisture content of the feed was observed to markedly influence the biocrude yield, which in the case of sewage sludge with a moisture content of 85% (w/w) treated through isothermal HTL was 26.8% (w/w) while that from fast HTL equivalent was 27.5% (w/w). Aliphatic long-chain hydrocarbons and related acids mostly occupied the contents of biocrude. Dichloromethane used as a biocrude recovery solvent could pool higher biocrude and energy content (approximately 50%) than other recovery solvents namely, hexane, xylenes, chloroform, methanol, ethanol, and acetone. In this case study, it was noticed that biocrude obtained from fast HTL showed a lower N/C ratio and a higher ratio of H/C and O/C than biocrude produced through isothermal counterpart. The fast HTL of microalgae resulted in maximum biocrude of ∼45% (w/w) at lower microalgae loading while isothermal HTL resulted in 37.5% (w/w) [56]. However, under these conditions biocrude with the lowest energy density (32.5 MJ/kg) was obtained, indicating that though higher amounts of biocrude may be recovered through fast HTL, the quality may be inferior to that obtained from isothermal counterpart (33.7 MJ/kg).

2.2.3

HTL Products

The HTL product streams basically include gaseous, aqueous, and solid phases. The product stream evolving from the HTL reactor is pumped into the gas separators viz., hydro-cyclones/gas–liquid extractor systems to collect the gaseous products. The aqueous–solid mixture separated through vacuum filtration units or gravimetric separators mainly consists of biochar/solid residue, nonvolatile aqueous phase and biocrude/bio-oil. Maag et al. conducted HTL of institutional FW at 300 ∘ C for 1 hour, catalyzed by either homogeneous Na2 CO3 or heterogeneous CeZrOx . The

2.2 Significance of Hydrothermal Liquefaction of Food Waste

gaseous stream was found to be composed of 80% carbon dioxide (CO2 ), 10% carbon monoxide (CO), and 10% hydrogen gas (bio-H2 ), while methane (CH4 ), ethylene (C2 H4 ), and ethane (C2 H6 ) were in negligible amounts (1–2%) among other accumulated HTL gases. The overall carbon distribution was reported to be 38.8% as bio-oil, 21.7% as aqueous phase, and 23.6% in form of biochar through HTL of FW without any catalyst. While, HTL in the presence of Na2 CO3 showed a slightly different product profile wherein 10% reduction in biochar was noticed compared to the former process, the HTL reaction catalyzed using CeZrOx yielded higher carbon content in bio-oil and biochar fractions compared to gaseous and aqueous counterparts. This result indicates that the product profile significantly varies with the type of the reactants involved during HTL. Also, the energy density of the products was found to vary according to the reaction involved. The energy recovered from food waste through CeZrOx , Na2 CO3 , and HTL without catalyst was observed to be 38.8, 27.6, and 21.3%, respectively [59]. HTL of mixed food waste, starch and casein at 315 ∘ C produced black bio-oil with high viscosity that showed energy density of 31–38.2 MJ/kg [60]. On the contrary, untreated/raw equivalents of the feed showed lower energy density viz., 17–23.6 MJ/kg which reveals an upgrade in product/oil quality after HTL. Besides, it was found that the oxygen content of the HTL bio-oil (11.3–19.7 mass %) was lowered by 50% as compared to raw feed (29.6–53.9 mass %), while, the carbon mass % markedly increased in bio-oil (71.3–75.2 mass %) relative to raw feed (38.3–47.6 mass %). In a separate study, different food wastes namely, cheese, meat, and fruit wastes when treated at 300 ∘ C showed significant variation in yield of HTL products [61]. The moisture content of the raw feed was in the range of 20.75–87.9% (w/w). In this case, the HTL reaction was initiated with dried feed of 30 g each wherein after HTL, biochar obtained was 4.94, 1.17, and 11.95 g for cheese, meat, and fruit waste, respectively. The main product, i.e. bio-oil was obtained in relatively higher amounts from cheese (5.02 g) compared to meat (1.57 g), while, fruit waste did not yield bio-oil. On the other hand, aqueous fractions were quantified to be within 152.27–164.26 g for the investigated feed. From this study, the overall yield of bio-oil was reported to be 75.8, 60.5, and 9.9% for waste from cheese, meat, and fruit processing, respectively. These reports indicate the feasibility of valorization of FW into high-heating value products through HTL technology [61].

2.2.4

Greenhouse Gas Emissions

Climatic aberrations are more prominent in recent years due to increasing urbanization, industrialization, and progressive global transport systems that increased carbon overload in the environment. The heavily polluting CO2 emissions from combustion of dwindling fossil fuels that are near to exhaustion are becoming a grave concern. Although; methane and water vapor also fall under heat-trapping gases, increasing CO2 concentration in the atmosphere can trigger irreversible changes posing substantial risk to health and the environment. Joos et al. reported that 40% of CO2 released into the atmosphere remains for 100 years while the rest can persist till 1000–10 000 years [62]. Among the total greenhouse gases (GHGs)

31

32

2 Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy

emitted, 72% is occupied by CO2 alone, which has been found to constantly increase by 1.6 times over the past decades [63]. The severe implications of atmospheric CO2 necessitate GHG reduction alternatives. On the other hand, Air Transport Action Group had set up mandates to initially stabilize net aviation CO2 emission by 2020 and later by 2050; it is set to cut emissions by 50% to those in 2005. These mandates demand a sustainable supply of low-carbon bio-jet fuels, which can be produced through upgrading the bio-oil obtained through HTL technology [64]. In this context, HTL-derived bio-oil can immensely contribute toward curtailing the burgeoning GHGs (specifically CO2 equivalent) that majorly originate from the transportation sector driven by conventional petro-fuels. Life cycle assessment (LCA) of GHG emissions from motor-fuel derived from HTL of forest waste at British Columbia using a hypothetical plant with 100 million liter/year capacity revealed that 17 g CO2 equivalent/MJ accounting for 82% CO2 reduction compared to 2005 gasoline baseline with 93 g CO2 equivalent/MJ can be achieved. Besides, on using bio-char obtained during HTL for soil improvement, GHGs can be further curtailed by 6.8 g CO2 equivalent/MJ, accounting for 89% reduction relative to petroleum fuels [65]. In a separate report, Tzanetis et al. reported suggestively lower GHG emissions (15%) are possible on combustion of bio-jet fuel derived from upgraded bio-oil produced from HTL of biomass relative to fossil-based jet fuel [64]. Connelly et al. reported 50% reduction in GHGs can be feasible on using HTL-derived algal fuels in comparison to fossil fuel counterpart [66]. Thus, these studies indicate the potential of HTL technology in curbing the GHG emissions and its associated health and environmental anomalies.

2.3 Factors Influencing HTL During FW Treatment During the course of HTL, several series of reactions namely, depolymerization of FW, decomposition by cleavage of monomers, dehydration, decarboxylation, and deamination occur followed by repolymerization and recombination of the reactive free radicals and fragments under the provided conditions [67]. Therefore, HTL-mediated transformation of wet biomass into crude oil and char is highly influenced by specific operating conditions like temperature, pressure, retention time, and catalyst or solvent used (Table 2.2). The reaction products mainly include two forms, i.e. solids (biochar) and liquids (biocrude) and often gases like H2 , CO2 , etc. are predictable depending on the conditions provided [71]. Several valuable organic compounds are reported to be identified in HTL-derived bio-oils that are inclusive of fractions of acids, esters, alcohols, ketones, aldehydes, miscellaneous oxygenates, sugars, furans, phenols, and syringols [77]. The yield of biocrude depends mainly on the process parameters showing distinctive prominence [71]. Temperature is a notable factor for HTL that leads to the degradation of the complex molecules. Under optimum temperature, solid molecules depolymerize and form free radicals favoring the radical reactions. Retention time, on the other hand, is closely coupled with temperature. As the retention time increases, temperature increases the reaction rate further allowing the secondary reactions to occur.

Table 2.2

S. No

Effect of reaction parameters and waste composition on hydrothermal liquefaction (HTL).

Waste type

Temperature (∘ C)

Time (min)

Catalyst

Pressure (bar)

Solvent

Products

Reference

1

Microalgae

310

60

–

Up to 180

Water

Nearly 34% Oil

[68]

2

Food waste

>200

10–60

–

–

Water

VFAs, HMF, furfural, ethanol, ketones

[69]

3

Meat, cheese, fruits

300

120

–

100

Water

10–75% bio oils

[70]

4

Food waste

250–370

10–120

K2 CO3 , KOH, Na2 CO3 , NaOH

100–300

Water

Carboxylic acids, alcohols, and aldehydes

[71]

5

Mixture of food waste

250–315

10–60

Na2 CO3 ,KOH

15

Water

< 40% oil

[60]

6

Waste food oils

600

1

–

–

Water

Pyrrolidines, pyrazines, fatty acid alkyl esters, and fatty acid amide

[55]

7

Food waste and coal blends

=300

60–120

–

45

Water

≈42% bio-oils

[72]

8

Food waste–woody biomass

180–260

≈60

–

24–55

Water

≈40% bio-oils

[73]

9

Food waste

300

60

Na2 CO3 , CeZrOx

–

Water

38.8% bio-oils

[59]

10

Carbohydrates, proteins

250, 350

20

–

20

Water

Piperidines, quinolones; pyrazine and its derivatives

[74]

11

Selected nutrient media

270–320

5–20

–

C-30), respectively. This widespread exploration showed that the bio-oil from FW can be considered as a prospective alternative to the fossil fuels.

2 Case Study: Tailor-Made Food Waste Blends for Improved Product Recovery Fox and co-workers [93] investigated the HTL of waste streams using tailor-made blends of municipal sewage sludge solids, poultry dissolved air flotation solids, and used cooking oil to produce sustainable fuel products. The hydrothermal processing of specific blended waste provided suppleness and scope for scalability of bio-based fuels in the near future. The HTL of waste slurry was conducted by initially heating it under subcritical conditions where the solid stream was separated. Further the reaction mixture was heated under pressurized reaction conditions to allow for HTL of waste to be transformed into value-added products. Before treatment, the blended waste slurry possessed a heating value of 8490 kJ/kg, while post-treatment the collected gas products showed 38 747 kJ/kg of energy density with 42% methane content along with renewable oil owning a heating value of 40 468 kJ/kg and 0.012% ash. This thermal treatment converted 86% of the initial carbon into energy in the form of liquid, gaseous, and solid products. The effluent stream obtained was found to be nutrient-rich with 1689 mg/L of nitrogen and about 14 000 mg/L TOC; thus it can be used as a liquid fertilizer. HTL resulted gas, liquid fuel, char, and nutrient-rich water are potential marketable products. Based on the outcomes from this study, it can be validated that HTL of blended FWs can be a feasible and sustainable technology to cater to the future energy demand.

3 Case Study: Food Waste Composition and Reaction Conditions Influence the Yield of HTL Products Aierzhati and co-workers [85] conducted HTL under elevated temperatures (280–380 ∘ C) at various reaction times (10–60 minutes) to transform different food residues that were collected from the dining hall, University of Illinois, into bio-oil. The biochemical composition analysis of different feed types revealed diverse individual protein, carbohydrate, and lipid content that ultimately influenced the product quality. The productivity of bio-oil ranged between 2 and 79% under the respective optimized HTL conditions of the studied feed type. An improved regression model focusing on both biochemical composition

2.5 Conclusions and Future Scope

and HTL reaction conditions was developed by including both descriptive HTL process energy recoveries and consumption ratios. An upgraded predictive model was developed, which precisely determined the yield of bio-oil with various FWs under diverse reaction conditions. Comprehensive experimental methodology and analyses helped to assess feasibility and scalability of HTL process in FW valorization. This study also offers a substantial basis for FW HTL with an in-depth analysis to enhance energy and nutrient recovery.

4 Case Study: Integration of HTL with Other Energy Recovery Processes Rao and co-workers [94] initially used HTL strategy to produce bio-oil from food and dairy waste. Further, the residual heat from the HTL process was utilized for membrane distillation of the aqueous effluents to produce a nutrient-rich concentrate that can be employed as a fertilizer to enhance soil quality. The drive behind this work is that the heat energy from the HTL process could be channeled toward the membrane distillation process, which eventually would improve the efficiency of the process and cut down the cost of operation. The HTL coupled polypropylene membrane system could recover around 75% water from the effluent without membrane fouling wherein the flux was positively maintained at high levels (> 10 LMH). The treated concentrate was found to consist of substantial amounts of ammonium and phosphate that can effectively serve as soil conditioner and fertilizer. Thus these in-depth studies reveal the potential of FW as a sustainable feed for bioenergy generation through HTL. Moreover, it can be comprehended that both HTL standalone and HTL integrated processes can valorize FW and its blends to significantly improve the economics of the process and decrease the selling price of the products.

2.5 Conclusions and Future Scope This chapter elucidated the viability of food waste valorization through HTL with a panorama on the availability, composition of food waste, current practices of HTL, factors affecting the product yield such as temperature, catalyst type, pressure, and retention time. The chapter also extensively discussed the present trends of HTL practices with relevant case studies. Thus, there is a momentum in the up scaling of HTL process for transformation of food waste to high-value compounds. Nevertheless, the operation of HTL at industrial scale is a quite challenging task. As the mechanism of HTL involves the accomplishment of subcritical conditions of water, i.e. high pressure and temperatures, this makes HTL exceedingly energy intensive. This is the main bottleneck of the HTL method. Through optimizing the process parameters, high yields of HTL products at lower temperature and pressure can be obtained. The employment of catalysts may favor the production of specific

39

40

2 Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy

and selective HTL products. A major challenge for HTL is instantaneous recovery of a high-value bio-oil/biocrude along with reduction in formation of low-value aqueous phase products. Novel methods need to be explored and designed for pretreatment and removal of oxygen, nitrogen, and other trace elements from food waste; therefore the plausibility of getting hydrocarbons (biocrude) would be in higher fraction. Thus, HTL can be a promising strategy for waste-to-wealth conversion inclusive of waste streams with excessive water content that are generally difficult to process through conventional waste treatment methods.

Acknowledgement Authors acknowledge funding from Council of Scientific & Industrial Research (CSIR), India in the form of project (MLP-0042; E3OW/FBR-NCP/IICT/2018RPPBDD). RPK duly acknowledges the University Grants Commission (UGC), India for providing research fellowship. AA sincerely thanks CSIR, India for financial assistance in the form of CSIR-Nehru Science Postdoctoral Fellowship (HRDG/CSIR-Nehru PDF/LS/EMR-I/04/2018). The authors wish to acknowledge the Director, CSIR-IICT (IICT/Pubs./2019/278) for kind support in carrying out this work.

References 1 Tiseo, I. (2019) Per capita food waste of selected countries worldwide in 2017. https://www.statista.com/statistics/933059/per-capita-food-waste-of-selectedcountries.html (accessed 25 July 2019). 2 Gustavsson, J., Cederberg, C., Sonesson, U. et al. (2011). Global food losses and food waste – Extent, causes and prevention. Rome: FAO. 3 Kjeldsen, P., Barlaz, M.A., Rooker, A.P. et al. (2002). Present and long-term composition of MSW landfill leachate: a review. Critical Reviews in Environmental Science and Technology 32 (4): 297–336. 4 Otles, S. and Kartal, C. (2018). Food waste valorization. In: Sustainable Food Systems from Agriculture to Industry (ed. C. Galanakis), 371–399. Academic Press. 5 Russ, W. and Schnappinger, M. (2007). Waste related to the food industry: a challenge in material loops. In: Utilization of by-Products and Treatment of Waste in the Food Industry (eds. V. Oreopoulou and W. Russ), 1–15. New York: Springer. 6 Ramachandran, S., Singh, S.K., Larroche, C. et al. (2007). Oil cakes and their biotechnological applications – a review. Bioresource Technology 98 (10): 2000–2009. 7 Pinacho, A., Garc𝚤’a-Encina, P.A., Sancho, P. et al. (2006). Study of drying systems for the utilization of biodegradable municipal solid wastes as animal feed. Waste Management 26 (5): 495–503.

References

8 Arvanitoyannis, I.S., Kassaveti, A., and Ladas, D. (2008). Food waste treatment methodologies. In: Waste Management for the Food Industries (ed. I.S. Arvanitoyannis), 345–411. New York: Elsevier. 9 Digman, B. and Kim, D.S. (2008). Review: alternative energy from food processing wastes. Environmental Progress 27 (4): 524–537. 10 Schaub, S.M. and Leonard, J.J. (1996). Composting: an alternative waste management option for food processing industries. Trends in Food Science & Technology 7 (8): 263–268. 11 Shilev, S., Naydenov, M., Vancheva, V. et al. (2007) Compost𝚤ng of food and agr𝚤cultural wastes). Utilization of by-Products and Treatment of Waste in the Food Industry (eds. V. Oreopoulou and W. Russ), 283–303. New York: Springer. 12 Venkata Mohan, S. (2009). Harnessing of biohydrogen from wastewater treatment using mixed fermentative consortia: process evaluation towards optimization. International Journal of Hydrogen Energy 34 (17): 7460–7474. 13 Wang, Y.H., Li, S.L., Chen, I.C. et al. (2009). Starch hydrolysis characteristics of hydrogen producing sludge in thermophilic hydrogen fermentor fed with kitchen waste. International Journal of Hydrogen Energy 34: 7435–7440. 14 Tawfik, A., Salem, A., and El-Qelish, M. (2011). Two stage anaerobic baffled reactors for biohydrogen production from municipal food waste. Bioresource Technology 102 (18): 8723–8726. 15 Tawfik, A. and El-Qelish, M. (2012). Continuous hydrogen production from co-digestion of municipal food waste and kitchen wastewater in mesophilic anaerobic baffled reactor. Bioresource Technology 114: 270–274. 16 Marone, A., Massini, G., Patriarca, C. et al. (2012). Hydrogen production from vegetable waste by bioaugmentation of indigenous fermentative communities. International Journal of Hydrogen Energy 37: 5612–5622. 17 Pasupuleti, S.B., Sarkar, O., and Venkata Mohan, S. (2014). Upscaling of biohydrogen production process in semi-pilot scale biofilm reactor: evaluation with FW at variable organic loads. International Journal of Hydrogen Energy 39 (14): 7587–7596. 18 Cavinato, C., Giuliano, A., Bolzonella, D. et al. (2012). Bio-hythane production from food waste by dark fermentation coupled with anaerobic digestion process: a long-term pilot scale experience. International Journal of Hydrogen Energy 37 (15): 11549–11555. 19 Yeshanew, M.M., Frunzo, L., Pirozzi, F. et al. (2016). Production of biohythane from food waste via an integrated system of continuously stirred tank and anaerobic fixed bed reactors. Bioresource Technology 220: 312–322. 20 Sarkar, O. and Venkata Mohan, S. (2017). Pre-aeration of food waste to augment acidogenic process at higher organic load: valorizing biohydrogen, volatile fatty acids and biohythane. Bioresource Technology 242: 68–76. 21 Chen, H. and Wu, H.Y. (2010). Optimization of volatile fatty acid production with cosubstrate of food wastes and dewatered excess sludge using response surface methodology. Bioresource Technology 101: 5487–5493.

41

42

2 Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy

22 Jiang, J., Zhang, Y., Li, K. et al. (2013). Volatile fatty acids production from FW: effects of pH, temperature, and organic loading rate. Bioresource Technology 143: 525–530. 23 Wang, K., Yin, J., Shen, D. et al. (2014). Anaerobic digestion of FW for volatile fatty acids (VFAs) production with different types of inoculum: effect of pH. Bioresource Technology 161: 395–401. 24 Dahiya, S., Kumar, A.N., Sravan, J.S. et al. (2018). Food waste biorefinery: sustainable strategy for circular bioeconomy. Bioresource Technology 248: 2–12. 25 Sarkar, O., Kumar, A.N., Dahiya, S. et al. (2016). Regulation of acidogenic metabolism towards enhanced short chain fatty acid biosynthesis from waste: metagenomic profiling. RSC Advances 6 (22): 18641–18653. 26 Li, Y., Su, D., Feng, H. et al. (2017). Anaerobic acidogenic fermentation of food waste for mixed-acid production. Energy Sources, Part A: Recovery, Utilization, and Environmental Effects 39 (7): 631–635. 27 Walker, K., Vadlani, P., Madl, R. et al. (2013). Ethanol fermentation from food processing waste. Environmental Progress & Sustainable Energy 32 (4): 1280–1283. 28 Huang, H., Singh, V., and Qureshi, N. (2015). Butanol production from FW: a novel process for producing sustainable energy and reducing environmental pollution. Biotechnology for Biofuels 8 (1): 147. 29 Takakuwa, N. and Saito, K. (2010). Conversion of beet molasses and cheese whey into fatty acid methyl esters by the yeast Cryptococcus curvatus. Journal of Oleo Science 59 (5): 255–260. 30 Chi, Z., Zheng, Y., Jiang, A. et al. (2011). Lipid production by culturing oleaginous yeast and algae with food waste and municipal wastewater in an integrated process. Applied Biochemistry and Biotechnology 165 (2): 442–453. 31 Papanikolaou, S., Dimou, A., and Fakas, S. (2011). Biotechnological conversion of waste cooking olive oil into lipid-rich biomass using Aspergillus and Penicillium strains. Journal of Applied Microbiology 110 (5): 1138–1150. 32 El Bialy, H., Gomaa, O.M., and Azab, K.S. (2011). Conversion of oil waste to valuable fatty acids using oleaginous yeast. World Journal of Microbiology and Biotechnology 27 (12): 2791–2798. 33 Pleissner, D., Lam, W.C., and Sun, Z. (2013). Food waste as nutrient source in heterotrophic microalgae cultivation. Bioresource Technology 137: 139–146. 34 Castanha, R.F., Mariano, A.P., and Morais, L.A.S. (2014). Optimization of lipids production by Cryptococcus laurentii 11 using cheese whey with molasses. Brazilian Journal of Microbiology 45 (2): 379–387. 35 Reddy, M.V. and Venkata Mohan, S. (2012). Influence of aerobic and anoxic microenvironments on polyhydroxyalkanoates (PHA) production from food waste and acidogenic effluents using aerobic consortia. Bioresource Technology 103 (1): 313–321. 36 Amulya, K., Jukuri, S., and Venkata Mohan, S. (2015). Sustainable multistage process for enhanced productivity of bioplastics from waste remediation through aerobic dynamic feeding strategy: process integration for up-scaling. Bioresource Technology 188: 231–239.

References

37 Gui, M.M., Lee, K.T., and Bhatia, S. (2008). Feasibility of edible oil vs. non-edible oil vs. waste edible oil as biodiesel feedstock. Energy 33: 1646–1653. 38 Ratledge, C. (2002). Regulation of lipid accumulation in oleaginous microorganisms. Biochemical Society Transactions 30: 1047–1050. 39 Economou, C.N., Aggelis, G., and Pavlou, S. (2011). Single cell oil production from rice hulls hydrolysate. Bioresource Technology 102: 9737–9742. 40 Subhash, G.V. and Venkata Mohan, S. (2011). Biodiesel production from isolated oleaginous fungi Aspergillus sp. using corncob waste liquor as a substrate. Bioresource Technology 102 (19): 9286–9290. 41 Amulya, K., Reddy, M.V., and Venkata Mohan, S. (2014). Acidogenic spent wash valorization through polyhydroxyalkanoate (PHA) synthesis coupled with fermentative biohydrogen production. Bioresource Technology 158: 336–342. 42 Alburquerque, J.A., Fuente, C., Ferrer-Costa, A. et al. (2007). Assessment of the fertilizer potential of digestate from farm and agroindustrial residues. Biomass & Bioenergy 40: 181–189. 43 Pleissner, D., Lam, W.C., Han, W. et al. (2014). Fermentative polyhydroxybutyrate production from a novel feedstock derived from bakery waste. BioMed Research International https://doi.org/10.1155/2014/819474. 44 Chen, W.T., Zhang, Y., Lee Timothy, H. et al. (2018). Renewable diesel blend stocks produced by hydrothermal liquefaction of wet biowaste. Nature Sustainability 1: 702–710. 45 Areeprasert, C., Ma, D., Prayoga, P. et al. (2016). A review on pilot-scale applications of hydrothermal treatment for upgrading waste materials. International Journal of Environmental Science and Development 7 (6): 425–430. 46 Matsumura, Y. (2015). Hydrothermal gasification of biomass. In: Recent Advances in Thermochemical Conversion of Biomass (eds. A. Pandey, B. Thallada, M. Stöcker and R. Sukumaran), 251–267. Elsevier. 47 Patel, B., Guo, M., Izadpanah, A. et al. (2016). A review on hydrothermal pre-treatment technologies and environmental profiles of algal biomass processing. Bioresource Technology 199: 288–299. 48 Cantero, T., Cantero, A.D., Celia, M.M. et al. (2018). Characterization of the solid products from hydrothermal liquefaction of waste feedstocks from food and agricultural industries. The Journal of Supercritical Fluids 133: 665–673. 49 Peterson, A.A., Vogel, F., Lachance, R.P. et al. (2018). Thermochemical biofuel production in hydrothermal media: a review of sub- and supercritical water technologies. Energy and Environmental Science 1: 32–65. 50 Jindal, M.K. and Jha, M.K. (2016). Effect of process parameters on hydrothermal liquefaction of waste furniture sawdust for bio-oil production. RSC Advances 6 (48): 41772–41780. 51 Wagner, L.J., Le, D.C., Ting, P.V. et al. (2017). Design and operation of an inexpensive, laboratory-scale, continuous hydrothermal liquefaction reactor for the conversion of microalgae produced during wastewater treatment. Fuel Processing Technology 165: 102–111. 52 Raikova, S., Smith-Baedorf, H., Bransgrove, R. et al. (2016). Assessing hydrothermal liquefaction for the production of bio-oil and enhanced metal recovery from

43

44

2 Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy

53

54

55

56

57

58 59 60

61

62

63

64

65

66

microalgae cultivated on acid mine drainage. Fuel Processing Technology 142: 219–227. Alhassan, Y. and Kumar, N. (2015). Catalytic hydrothermal liquefaction of solid food waste for light bio-oil production: process optimisation. Journal of Biofuels 6 (1): 38–43. Anastasakis, K., Biller, P., Madsen, B.R. et al. (2018). Continuous hydrothermal liquefaction of biomass in a novel pilot plant with heat recovery and hydraulic oscillation. Energies 11: 2695. Gollakota, A. and Savage, P.E. (2018). Hydrothermal liquefaction of model food waste biomolecules and ternary mixtures under isothermal and fast conditions. ACS Sustainable Chemistry & Engineering 6 (7): 9018–9027. Faeth, J.L., Savage, P.E., Jarvis, J.M., and McKenna, A.M. (2016). Characterization of products from fast and isothermal hydrothermal liquefaction of microalgae. AIChE Journal 62 (3): 815–828. Jiang, J. and Savage, P.E. (2017). Metals and other elements in biocrude from fast and isothermal hydrothermal liquefaction of microalgae. Energy & Fuels 32 (4): 4118–4126. Qian, L., Wang, S., and Savage, E.P. (2017). Hydrothermal liquefaction of sewage sludge under isothermal and fast conditions. Bioresource Technology 232: 27–34. Maag, R.A., Paulsen, D.A., Amundsen, J.T. et al. (2018). Catalytic hydrothermal liquefaction of food waste using CeZrOx. Energies 11: 564. Zastrow, J.D. and Jennings, A.P. (2013). Hydrothermal liquefaction of food waste and model food waste compounds. AIChE Annual Meeting, Paper No. 336978. Kostyukevich, Y., Vlaskin, M., Borisova, L. et al. (2018). Investigation of bio-oil produced by hydrothermal liquefaction of food waste using ultrahigh resolution Fourier transform ion cyclotron resonance mass spectrometry. European Journal of Mass Spectrometry 24 (1): 116–123. Joos, F., Roth, R., Fuglestvedt, J.S. et al. (2013). Carbon dioxide and climate impulse response functions for the computation of greenhouse gas metrics: a multi-model analysis. Atmospheric Chemistry and Physics 13 (5): 2793–2825. Hosseini, S.A., Lashgari, H., Choi, J.W. et al. (2013). Static and dynamic reservoir modeling for geological CO2 sequestration at Cranfield, Mississippi, USA. International Journal of Greenhouse Gas Control 18: 449–462. Tzanetis, F.K., Posada, A.J., and Ramirez, A. (2017). Analysis of biomass hydrothermal liquefaction and biocrude-oil upgrading for renewable jet fuel production: the impact of reaction conditions on production costs and GHG emissions performance. Renewable Energy 113: 1388–1398. Yuhao, N. and Bi, X. (2018). Life-cycle assessment of transportation biofuels from hydrothermal liquefaction of forest residues in British Columbia. Biotechnology Biofuels 11: 23. Connelly, E.B., Colosi, L.M., Clarens, A.F., and Lambert, J.H. (2015). Life cycle assessment of biofuels from algae hydrothermal liquefaction: the upstream and downstream factors affecting regulatory compliance. Energy & Fuels 29 (3): 1653–1661.

References

67 Gollakota, A.R.K., Nanda, K., and Gu, S. (2017). A review on hydrothermal liquefaction of biomass. Renewable and Sustainable Energy Reviews 81: 1378–1392. 68 Coma, M., Martinez, H.E., Abeln, F. et al. (2017). Organic waste as a sustainable feedstock for platform chemicals. Faraday Discussions 202: 175. 69 Usman, M., Chen, H., Chen, K. et al. (2019). Characterization and utilization of aqueous products from hydrothermal conversion of biomass for bio-oil and hydro-char production: a review. Green Chemistry 21: 1553–1572. 70 Yury, K., Mikhail, V., Ludmila, B. et al. (2017). Investigation of bio-oil produced by hydrothermal liquefaction of food waste using ultrahigh resolution Fourier transform ion cyclotron resonance mass spectrometry. European Journal of Mass Spectrometry https://doi.org/10.1177/1469066717737904. 71 Deniel, M., Haarlemmer, G., Roubaud, A. et al. (2016). Energy valorisation of food processing residues and model compounds by hydrothermal liquefaction. Renewable and Sustainable Energy Reviews 54: 1632–1652. 72 Saqib, N.U., Sarmah, K.A., and Baroutian, S. (2019). Effect of temperature on the fuel properties of food waste and coal blend treated under co-hydrothermal carbonization. Waste Management 89: 236–246. 73 Wang, T., Zhai, Y., Li, H. et al. (2018). Co-hydrothermal carbonization of food waste-woody biomass blend towards biofuel pellets production. Bioresource Technology 267: 371–377. 74 Fan, Y., Hornung, U., Dahmen, N.K. et al. (2018). Hydrothermal liquefaction of protein-containing biomass: study of model compounds for Maillard reactions. Biomass Conversion and Biorefinery 8 (4): 909–923. 75 Yang, J., Corscadden, K., and Niu, H. (2019). Advanced models for the prediction of product yield in hydrothermal liquefaction via a mixture design of biomass model components coupled with process variables. Applied Energy 233: 906–915. 76 Castello, D., Pedersen, T., and Rosendahl, L. (2018). Continuous hydrothermal liquefaction of biomass: a critical review. Energies 11 (11): 3165. 77 Milne, T.A., Evans, R.J., and Abatzaglou, N. (1998). Biomass Gasifier “Tars”: Their Nature, Formation, and Conversion (No. NREL/TP-570-25357; ON: DE00003726) (ed. National Renewable Energy Laboratory). Golden, CO. 78 Parshetti, G.K., Suryadharma, M.S., Pham, T.P.T. et al. (2015). Heterogeneous catalyst-assisted thermochemical conversion of food waste biomass into 5-hydroxymethylfurfural. Bioresource Technology 178: 19–27. 79 Biswas, B., Kumar, A.A., Bisht, Y. et al. (2017). Effects of temperature and solvent on hydrothermal liquefaction of Sargassumtenerrimum algae. Bioresource Technology 242: 344–350. 80 Liu, H.M., Feng, B., and Sun, R.C. (2011). Acid–chlorite pretreatment and liquefaction of cornstalk in hot-compressed water for bio-oil production. Journal of Agricultural and Food Chemistry 59 (19): 10524–10531. 81 Barreiro, L.D., Gómez, R.B., Ronsse, F. et al. (2016). Heterogeneous catalytic upgrading of biocrude oil produced by hydrothermal liquefaction of microalgae: state of the art and own experiments. Fuel Processing Technology 148 (2016): 117–127.

45

46

2 Hydrothermal Liquefaction of Food Waste: A Potential Resource Recovery Strategy

82 Tran, K.Q. (2016). Fast hydrothermal liquefaction for production of chemicals and biofuels from wet biomass–the need to develop a plug-flow reactor. Bioresource Technology 213: 327–332. 83 Mazaheri, H., Lee, K.T., Bhatia, S. et al. (2010). Sub/supercritical liquefaction of oil palm fruit press fiber for the production of bio-oil: effect of solvents. Bioresource Technology 101: 7641–7647. 84 Liu, Z. and Zhang, F.S. (2008). Effects of various solvents on the liquefaction of biomass to produce fuels and chemical feedstocks. Energy Conversion and Management 49 (12): 3498–3504. 85 Aierzhati, A., Stablein, J.M., Wu, E.N. et al. (2019). Experimental and model enhancement of food waste hydrothermal liquefaction with combined effects of biochemical composition and reaction conditions. Bioresource Technology 284 (2019): 139–147. 86 Braguglia, C.M., Gallipoli, A., Gianico, A. et al. (2018). Anaerobic bioconversion of food waste into energy: a critical review. Bioresource Technology 248: 37–56. 87 Li, F., Liu, L., An, Y. et al. (2016). Hydrothermal liquefaction of three kinds of starches into reducing sugars. Journal of Cleaner Production 112: 1049–1054. 88 Zhu, Y., Biddy, M.J., Jones, S.B. et al. (2014). Techno-economic analysis of liquid fuel production from woody biomass via hydrothermal liquefaction (HTL) and upgrading. Applied Energy 129: 384–394. 89 Zhang, L., Xu, C.C., and Champagne, P. (2010). Overview of recent advances in thermo-chemical conversion of biomass. Energy Conversion and Management 51 (5): 969–982. 90 Toor, S.S., Rosendahl, L., and Rudolf, A. (2011). Hydrothermal liquefaction of biomass: a review of subcritical water technologies. Energy 36: 2328–2342. 91 Hammerschmidt, A., Boukis, N., Hauer, E. et al. (2011). Catalytic conversion of waste biomass by hydrothermal treatment. Fuel 90 (2): 555–562. 92 Chen, W.H., Lin, Y.Y., Liu, H.C. et al. (2019). A comprehensive analysis of food waste derived liquefaction bio-oil properties for industrial application. Applied Energy 237: 283–291. 93 Fox, J.T., Zook, A.N., Freiss, J. et al. (2019). Thermal conversion of blended food production waste and municipal sewage sludge to recoverable products. Journal of Cleaner Production 220: 57–64. 94 Rao, U., Posmanik, R., Hatch, L.E. et al. (2018). Coupling hydrothermal liquefaction and membrane distillation to treat anaerobic digestate from food and dairy farm waste. Bioresource Technology 267: 408–415.

47

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England Pauline Deutz and Anne Kildunne Department of Geography, Geology and Environment, University of Hull, Hull, UK

3.1 Introduction Rising interest in the concept of the circular economy sees increasing policy and industry efforts toward keeping resources in productive use for as long as possible. This includes strategies such as capturing pre- and post-consumer residues for incorporation into other products or processes [1, 2]. Significantly, circular economy also involves the intentional design of products and infrastructure to facilitate resource recovery and prevent waste from leaking out of the economy into the environment, i.e. “loop closing” in terminology drawing concepts from ecology [3]. The current attention to the long-term environmental impact of waste materials such as plastics indicates the relevance of this effort [4]. However, whilst academic definitions of a circular economy can be aspirational (e.g. [2, 5]), policy makers have the prosaic task of adapting current practice [6]. The UK national government is among the many in the EU and elsewhere examining how to incentivize changes in the practice of private companies, which, filtered through consumer practices, will influence the task for local authorities with a statutory responsibility for household waste management. Local authorities (LAs) have long had this responsibility, with little or no influence over the nature of the waste stream presented to them by the public [7]. Even if national governments exercise upstream control over waste generated at the local level, LAs will remain a critical pivot point between product distribution and resource recovery [8]. The role of LAs therefore provides a useful perspective from which to explore the complexity of interwoven issues and relationships involved in implementing loop closing within a circular economy. In this chapter, we consider the significant change in waste management practice in England1 between 2000 and 2010 with the aim of providing a geographically and historically contextualized understanding of how this was brought about, focusing on the role of LAs. At a national scale, landfill dependency for household waste 1 As waste is a devolved policy area in the UK, we refer to “England” as the national level of government, as some policies or arrangements different between the different countries comprising the UK. Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

48

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

reduced from 79% to roughly 43% [9] during the decade following the publishing of the then government’s Waste Strategy in 2000 [10]. This required not just technological change but an ontological adjustment in the understanding of waste from something requiring cheap disposal, i.e. primarily landfill, to a resource with potential value [11, 12]; the responsibilities of LAs changed from safely managing waste to establishing resource recovery operations. Conceptualizing this change in practice in English waste management as a sustainability transition, we use the latter literature as a theoretical framework for the study. Researchers have defined a transition as, e.g., an observed period of change between different socio-technical systems (i.e. combinations of technology and institutional factors around its usage) [13, 14]; or periods with an opportunity for change [15]. Bailey and Wilson [16] referred to periods of relative policy constancy (when certain approaches/priorities are paradigmatic, to use Hall’s expression [17]), interspersed by times when the policy situation is more fluid and changes can more readily occur albeit constrained within boundaries of social acceptability. These interpretations of transitions form a starting point for interrogating the case study, drawing on more recent work that emphasizes policy processes and the agency of actors and policy [18–20]. As Giddens [21] noted, a key purpose of social science is to understand change – no less important if a further purpose is to help bring change about. This chapter combines a geographically defined case study covering various LAs within Yorkshire and Humberside, with a long view of the 2000s. The study draws on semi-structured interviews, extensive document analysis, and the voluminous LA waste statistics collected by the UK government to ask: (i) What were the European and national-scale influences on LAs in England, before and during the key period? (ii) How did authorities in the case study region respond to those influences and with what outcomes? (iii) What have been the initial effects of the changing context since? We begin by reviewing the sustainability transitions literature as the theoretical framework for the study and outlining the waste literature. We explain our research design and methods, then analyze first the EU and then the wider political context of the 2000 Waste Strategy. We then report on how a geographically bounded group of English LAs has worked within these policy constraints to affect the transition from “waste management” to “resource management,” particularly through their changing relationships with the public, i.e. householders.

3.2 Sustainability Transitions Literature Sustainability transitions research seeks to understand how societies can become more sustainable given increasing economic and population pressures, rising global carbon emissions and limited environmental resources [22–24]. The core assumption of the transition studies approach is that technological innovations are crucial for sustainability but insufficient in and of themselves, instead requiring interaction with ongoing social and institutional processes to deliver change

3.2 Sustainability Transitions Literature

[20, 22]. It acknowledges the multiple stakeholders, technological innovations, and institutions involved in achieving progress toward sustainability [22–25]. Early transitions research often overemphasized the importance of technological innovations in driving change, rather than long-term changes in behavior on the part of the public [26] or policy innovations [27]. However, the field has blossomed with no less than 96 frameworks or theorizations recently identified [28]. A comprehensive review of all these theories is beyond the scope of this chapter, but we highlight the multilevel perspective or MLP [13, 20, 22, 23]. The MLP [22, 23, 29, 30] uses a systems perspective, attempting to provide a whole-picture explanation. It is particularly appropriate here as the management of materials involves both long-established complex infrastructure enmeshed with legislation, industries, markets, institutions, and societal habits. It is a framework to explain how change can come about within what is termed the regime, with a focus on the adoption of technological innovation [14, 29]. The MLP distinguishes between protected areas of activity (niches) where new technologies may gain a foothold, the regime (comprising the stakeholders, institutions, and existing practice relating to the activity in question), and the landscape – seen as exogenous long-term factors beyond the scope of individual actors. These factors frame attitudes and behavior in the regime [13, 14, 21, 23, 31] (see Table 3.1). The upper, landscape, level in the multilevel model is the least well defined. This comprises factors considered exogenous to the system including “cultural and normative values” that may impact on the regime to bring about change [29, p. 1260]. A change in cultural values might support stakeholders with agency in the perception that a change would be in their interests. An example of that could be the rise in the electoral strength of avowedly green parties bringing about the inclusion of environmental issues in the manifestos of mainstream political parties. Markard and Truffer [32, p. 606] interpret the landscape to refer to “the set of residual factors that have an impact on innovation and production processes without being influenced by the outcome of innovation processes on a short to mid-term basis.” Table 3.1 Definition of terms used for the levels in the multilevel perspective and processes by which change is held to occur. Definition

Landscape

Regime

Niche

Change processes

Macroeconomic trends, e.g. globalization, climate change ● Cultural patterns, e.g. public attitudes to the environment and its perceived value Formal and informal rules, e.g. guiding principles, relationships, regulations, policies

Gradual long-term change beyond the scope of individual actors

Small-scale innovative technologies and changes in practices

Disruptive change based on differing expertise, new networks, and expectations

●

Incremental changes based on existing knowledge, networks, and expectations

Source: Giddens [21]; Affolderbach and Schulz [23]; Geels and Kemp [30]; Kern [31].

49

50

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

This is implicitly following a systems perspective, placing the landscape outside the system. But such boundaries are artificial, in social systems [33, 34]. Taking a dialectical, or critical realistic, approach, relationships can be examined without postulating a boundary between different elements, and furthermore allows the uncovering of underlying causal factors that may not be apparent from an analysis of directly observable characteristics of a predefined system. The landscape is the least thoroughly examined and theorized aspect of the MLP. Transitions occur through interactions between these levels. Defining the regime as the institutions involved in the operation of the dominant system, almost by definition regime members have a vested interest in the maintenance of the status quo. The regime actors include both policy makers who could propose and enforce changes, and manufacturers/services providers related to the regime who might lobby against change. The resistance to change may account in part for periods of relative stability when change is only incremental [18, 35]. However, a change at the landscape level, e.g. prioritization of climate change, can cause disruptions in the regime, which offer “windows of opportunity” for new niche technologies and practices to gain both social acceptance and financial viability [24]. The MLP approach has been applied in a variety of sectors and contexts (inter alia transport [13]; aviation [36]; energy [23, 24, 37, 38]; urban sustainability issues [22]). Particular criticism of the MLP framework has noted the lack of consideration of political processes and the agency of actors and policy to bring about change [18, 19]. Avelino and Wittmayer [19] for instance consider the relative power of different stakeholders, categorized as state, market, or community, arguing that change may be brought about by shifts in the balance of power between these different interest groups. Meadowcroft [18] also sees this balance of power in political terms. Neither a spatial nor a hierarchical scale, “regime” in the MLP relates to a particular sector and its existing routines and institutions across scales of governance. The niche level includes many activities conducted by individual actors in a particular context, but is not necessarily limited to a particular local area. Local authorities might be considered part of the regime of governance, but their role also interacts closely with aspects of the landscape, in particular householders’ attitudes toward the environment and the extent to which waste is perceived as something that individuals can control. Geographic and governance scales, not addressed in earlier MLP literature, have recently been considered more explicitly too [22, 25]. Meadowcroft [18], for example, points to the need for the regional variability and governance structures of Canada to be considered within any national framework. The multi-scalar approach to the MLP remains, however, a relatively underexplored avenue [25]. This study therefore builds on recent work by considering the political context of the waste transition, using a regionally bounded case study to explore the context and response of stakeholders in depth. Furthermore, we are not simply looking at the regional to local scale. We adopt a multi-scale governance (MSG) framework as part of the ontology of the project permitting local and sub-national exploration with greater focus on the policy and process mechanics of how transitions occur. Thus the LA scale is considered in national and supranational context, in contrast to existing studies of waste-related transitions [30]. Arguably, a transition in waste

3.3 Waste Management in England

management comparable to the one studied in the UK herein had already been accomplished by the Netherlands by 2000 [30, 39], i.e. before the UK seriously engaged with waste diversion from landfill. The Dutch experience indicated how landscape shifts associated with scandals over waste management in the 1980s and increased environmental awareness on the part of the public were the backdrop to policy efforts in the regime in bringing about a shift away from landfill reliance [30, 39]. An important part of this transition was the creation of a national waste management council and the reorganization of local waste sites into four larger waste areas. The resulting larger quantities of waste and consequent economies of scale enabled large multinational waste management firms to invest in incineration. Thus, although an associated increase in recycling occurred, to a large extent this was a swap from one form of disposal to another but accompanied by energy recovery. Missing from the analysis, however, is an international dimension: i.e. what was the relationship between events in the Netherlands and the wider context of the EU? As discussed below, in the case of the UK that wider context is critical, not least due to the inspiration and political leverage that came from the example of the Netherlands [10]. With its efforts to grasp the interrelationships between stakeholders, institutions, and technology in the establishment of sustainability innovations, the sociotechnical transition offers a suitable theoretical perspective to shed light on developments in waste in the UK. In turn, a multi-scalar analysis of changes in waste practice in the UK offers a different lens by which to analyze transition theory.

3.3 Waste Management in England During the 2000s, a substantial body of literature appeared variously addressing the issues associated with the emerging transition in waste management. Some of these concerned the attitudes and behavior of the public in response to LA initiatives [40–42], noting for example how LA’s attempts to educate, persuade, or more firmly encourage residents to change behavior, and how responses vary among different sections of the population. Indeed changes in the role of waste services from disposal to diversion have adjusted the boundaries of public and private responsibility; increases in the recycling rate and other measures have only been possible because the public have been prepared to adjust their behavior [43, 44]. Jalil et al. [8] identified a symbiotic relationship between LAs and the local population; the services provided need to work to the convenience of both parties to this relationship. LAs, however, are dependent on the vicissitudes of national government preferences to determine both the scope of their obligations and the manner in which they can go about fulfilling them [45]. Progress in waste diversion in the England was made in the context of changes to regional-scale governance, including the planning regime [46], and the appearance of a range of policy instruments used by national governments (including the extensive imposition of targets) to steer regional/local government to achieve the ends imposed from the EU [47]. These studies situate changes in English waste management within a governance

51

52

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

perspective, acknowledging the centralized model of UK2 governance (so-called Westminster model), which, given limited confidence in the longevity of funding and tax incentives, limits both local authorities and firms from co-operating on a long-term basis [48, 49]. UK commentators have largely framed the waste transition as a success story noting significant improvements in the level of recycling, but focus predominantly on waste as an urban issue [e.g. 50]. Regional variations in the England have been noted quantitatively [e.g. 51], but have been subject to less qualitative analysis. Moreover, these studies were undertaken before the effectiveness of the structures could be known; let alone the impacts of later changes. Ehnert et al. [52, p. 107] for instance refers to the “hollowing out” of UK local government since 2010, particularly in terms of funding cuts. Other research focusing on operational changes by LAs to increase recycling levels often fails to consider the wider political and legislative context. Waite et al. [53] for example analyze 48 local authorities between 2008 and 2013 and find that a smaller bin for residual waste combined with food waste collection increased recycling rates. Farmer et al. [54] notes how moves toward incineration rather than landfill constitute an incremental move up the waste hierarchy, which conflicts with resource recovery and utilization through recycling activities. The technological solution of incineration fails to shift the long-term sub-optimal pattern of resource utilization requiring no changes in public attitudes toward waste or consumption patterns and local authorities remain tied into existing contract relationships with waste firms. Recycling is not an optimum solution from a circular economy perspective [55]. There are concerns for a rebound effect, whereby if recycling is seen as environmentally adequate, it may remove a check on increasing consumption [53, 56]. Gharfalkar et al. [55] call for more policy attention to address circular concepts such as reduce and reuse, which are not clearly defined in EU legislation. Notwithstanding the limitations of recycling in a circular economy vision, from a policy stakeholder perspective, it is important to understand how the challenges of increasing recycling rates were overcome. Insight from the waste transition can serve as a guide for future developments around this and other policy areas (including for plastics, for which recycling rates are stubbornly low) and in countries where a comparable transition has yet to occur.

3.4 Research Design and Methods This chapter employs a critical realist approach, which seeks to examine the underlying causal factors behind observed events and perspectives [33, 57]. Furthermore, we approach this with a dialectical understanding of the interrelationships between entities, which are routed in spatial and temporal context, whilst being continuously subject to change [34]. In common with recent ST literature which 2 This is characterized by reliance on central government hold on funding, which directs local authorities into a model that favors market instruments and also tends to be short-term/tied into national the political cycle.

3.4 Research Design and Methods

emphasizes agency [52, 58], and the importance of politics [18, 23], but tends to focus on that of national government, [52] we also recognize the reciprocal nature of the relationship between local authorities and other governance structures [20]. We have selected LAs as the entry point through which to examine the political process, whilst also drawing on the perspectives of other stakeholders. The socio-technical regime for waste in England, as doubtless elsewhere, is a complex web of organizations and, critically, the multifarious array of individuals and households comprising “the public.” In waste, as in other policy areas, national government relies on the actions of LAs to fulfill its policy commitments, using an array of policy instruments to “steer” their activities, whilst leaving considerable latitude for local decisions about service delivery [e.g. 48].

3.4.1

Research Design

Geographically the study focuses on the Yorkshire and Humberside region of England (Figure 3.1). Bounding the study geographically provides the opportunity to assess the behavior and performance of an environmentally and politically diverse range of local authorities and to cover a limited area in some depth in a range of organizational contexts (including urban/rural/unitary/waste collection/disposal authority, different waste companies, etc.). This selection provides a microcosm of the England nation state, as this area was almost exactly at the average in terms of performance compared to other English local authorities (see Figure 3.5). Moreover, the Yorkshire and Humberside area covers urban areas such as Leeds and Sheffield, but also largely rural areas. Taking a time period as a case study is also problematic, as it implies starting and finishing points, which, though, inevitably subjective to some extent, are nonetheless necessary to frame the boundary of the study [34]. To compensate for that, we are working with porous boundaries, taking 2000 as a starting point, but glancing backward, and 2010 as an ending point but cognizant of changes presently in motion. The year 2010 is significant for both policy and political reasons, as will be discussed below.

3.4.2

Methods

Our research utilized a combination of primary interview data (largely through 24 face-to-face interviews conducted between June 2010 and August 2011) supported by secondary data sources. The use of interview data is an established methodology within the sustainability transitions literature [19, 53] and is also drawn on heavily within policy studies [59]. Drawing on secondary data sources allows us to triangulate data and improves the robustness of the research. A key feature of UK national policy toward local authorities from 1995+ has been the collection of staggering amounts of data on performance across all areas of activity, including waste. We have also consulted a wide range of documents including EU, national, regional, and local policy statements, guidance, reports, and evaluations.

53

54

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

Figure 3.1 Case study authorities in regional, national, and international context. Maps made by Graham Ferrier.

3.4.3

Selection of Interviewees

A purposive sampling strategy was utilized, i.e. the targeted selection of interviewees representing different roles within local authorities of different types. Table 3.2 shows the five groups of interviewees: local authority waste managers, waste industry personnel, policy makers; regulators, and academics. In several instances, we were able to interview more than one person from a given organization, which provides perspectives from different operational areas, as well as varying levels of seniority/experience [60]. Other interested parties including regulators and academics were also interviewed, adapting the topics to their particular expertise

3.4 Research Design and Methods

Table 3.2

Interviewees and their organizational context. Organization’s waste responsibility

Individual’s position (years’ experience)

Unitary Authority

Collection and disposal

Service Manager Waste infrastructure; former secretary of NAWDO; 10 years

East Riding of Yorkshire

Unitary Authority

Collection and disposal

Head of waste national role in CIWM

City of Hull

Unitary Authority

Collection and disposal

Asst Head of Waste >12 years

North Lincolnshire

Unitary Authority

Collection and disposal

Waste and recycling services manager >11 years

North Lincolnshire

Unitary Authority

Collection and disposal

Head of Waste >20 years

North East Lincolnshire

Unitary Authority

Collection and disposal

>20 years’ experience of waste and regional LARAC rep

City of York

Unitary Authority

Collection and disposal

Sets up and runs the recycling program

North Yorkshire

County Council

Disposal

Waste partnership officer

Scarborough

Borough Council in North Yorkshire

Collection

City of Leeds

Metropolitan district council

Collection and disposal

Implementing waste strategy 1 year; previously at EA

Industry A

Large private waste company

Landfill management, collection/disposal contracts

Executive director, business development; retired >20 years

Industry A

Large private waste company

Landfill management, collection/disposal contracts

Landfill management; retired

Industry A

Large private waste company

Landfill management, collection/disposal contracts

Permitting; >20 years

Industry B

LA owned (“arms’ length”) company

Landfill management, collection/disposal contracts

Planning and permitting issues 6 years

Industry B

LA owned (“arms’ length”) company

Landfill management, collection/disposal contracts

Business development >20 years’ experience of waste

Industry C

Large private waste company

Landfill management, collection/disposal contracts

Communications >20 years’ experience of waste)

Organization

Type

East Riding of Yorkshire

(continued)

Recycling manager 5 years

>20 years

55

56

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

Table 3.2

(Continued)

Organization

Type

Organization’s waste responsibility

Individual’s position (years’ experience)

Industry D

International engineering consultancy

EIA, operation, research, modeling relating to landfill

Hydro-geologist >20 years’ experience of waste

Industry E

Trade association for waste industry and professional body individuals

Lobbies on behalf of waste industry; training and profession development for individuals

Senior executive in both organizations

Defra

National environmental policy maker

Devises the UK waste strategy and selects policy instruments

Environmental management – landfills

Defra

National environmental policy maker

Devises the UK waste strategy and selects policy instruments

Waste diversion – LATS

Environment Agency

Regulator for operations including LATS

Advisor to PPC Inspectors

Permitting; 15 years’ experience

Environment Agency

Regulator for operations including LATS

PPC Inspector

PPC officer

University A

Research, consultancy, teaching

Research and consultancy

Research scientific aspects of waste; advisory body to EA; served DEFRA’s waste Foresight committee

University B

Research, consultancy, teaching

Teaching and industrial liaison

Former regional co-ordinator of CIWM, lectures in waste

>20 years’ experience of waste

because of their wider “landscape” experience of the contextual development of the legislation and particular expert knowledge [61, 62]. As interviewees were promised anonymity, interviews will be referenced by code, e.g. LA (local authority) 1, WI (waste industry) 2, etc. These local authorities are socially and geographically contrasting (Table 3.3), ranging from the dense urban area of Hull to the relatively sparsely populated rural county of North Yorkshire and farmland of the East Riding of Yorkshire. There is significant variation in size of area to be served, as well as population density. Socially and economically there is also a wide diversity between these locations, which is summed up by the deprivation ranking. The latter is based on the proportion of an authority’s population living in the most deprived areas in England, based on the Index of Multiple Deprivation. In 2010 this was based on a weighted average of the following indicators: income (22.5%); employment (22.5%);

Table 3.3

Contextual information on local authorities included in interviews and/or analysis of waste statistical data.

Local authority

Authority type

Population 2001a)

Population 2011a)

Area km2

Population density 2011 (people/km2 )

Deprivationb) Rank among sub-county LAs in England, where 1 is most deprived and 294 would be the least deprived

East Riding of Yorkshire

Unitary

314 900

334 200

2 479

135

171

City of Hull

Unitary

249 900

256 400

28

9 157

12

North Lincolnshire

Unitary

153 000

167 400

846

198

109

North East Lincolnshire

Unitary

158 000

159 600

192

832

37

City of York

Unitary

181 300

198 000

272

728

178

North Yorkshire

County Council

570 200

598 400

5 571

107

(ranked 125/149 on county scale)

Leeds

Metropolitan borough (within West Yorkshire)

715 600

751 500

552

1 361

59

Scarborough

Borough district of North Yorkshire

106 200

108 800

817

133

99

a) Taken from UK census data https://webarchive.nationalarchives.gov.uk/20160108133329/http://www.ons.gov.uk/ons/rel/mro/news-release/census-resultshows-increase-in-population-of-yorkshire-and-the-humber/censusyorkandhumbernr0712.html (29 July 2019). b) Rank based on extent of deprivation: extent is defined as the proportion of an area’s population living in the most deprived Lower-layer super output area (LSOA) in the country. https://docs.google.com/spreadsheets/d/1x5hE_5UW_VuLOV1h-c0CX1eiAchBPts72FFNzVVQOWw/edit?hl=en&hl=en#gid=0 (29 July 2019).

58

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

health and disability (13.5%); education, skills, and training (13.5%); barriers to housing and services (9.3%); crime (9.3%); and living environment (9.3%). Again there is a notable contrast between the City of Hull, which was on this measure the 12th most deprived place in the UK and the East Riding, which was the 171th. These figures obviously do not reveal variations in population density and deprivation within authorities.

3.4.4

Secondary Data

This study makes extensive use of the Waste Data Flow statistics provided by LAs to the UK government Department of Environment Food and Rural Affairs (Defra). Up to 2012, these statistics comprised part of the system of performance monitoring of local authorities and collectively contribute to the UK’s reporting to the EU with respect to waste targets. As the monitoring system changed in style and priorities between 1995 and 2012, and was then discontinued, the precise indicators collected also change (e.g. between weight of waste per household, to per capita). Definitions of municipal waste have also varied, leading to new expressions such as “local authority collected waste.” It is frustratingly difficult therefore to construct time series across the whole period. A further change in definitions in 2013/2014 means that comparisons before and after that time cannot readily be made [9]. For the case study LAs we have used data accessed through Defra’s waste statistics, which relate to England. To provide an international comparison EU, we are using data from Eurostat, which presents data at the UK scale.

3.5 Results and Discussion Between 2002 and 2009 the UK rapidly transitioned from landfilling 90% of domestic waste, whilst total waste production was still increasing, to 50% recovered (not landfilled) and total waste on a downward trend (Figure 3.2). The diversion-based (i.e. from landfill) approach to waste management referred to by Bulkeley [47] was firmly established, and the UK successfully met the EU’s 2010 targets for the diversion of biodegradable waste from landfill [63]. The following sections analyze these influences from the perspective of LAs, whose century-old, statutory responsibility for domestic waste collection and disposal continued throughout the introduction of modern waste and environmental regulations in the 1970s and beyond (Table 3.4), i.e. a long-standing institution was able to effect significant change.

3.5.1

English Waste in the Context of the EU

The key signal for change in practice to LAs was the government’s 2000 Waste Strategy [10], strongly reinforced by the 2002 follow-up [64]. Although broader in scope than the requirements of the then new EU Landfill Directive [65] the recovery targets (primarily recycling) set in England announced in the strategy

3.5 Results and Discussion 40 000 35 000

Waste amounts in 000s tonnes

30 000 25 000 20 000 15 000 10 000 5 000 0 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Year Tot waste

Landfill

Incinerated (w and w/o energy recovery)

Recycling

Composting and digesting

Diverted

Figure 3.2 Total waste generation (in 000 tonnes) and management routes for the UK. “Diverted” comprises the sum of all routes except landfill. Source: Data from http://appsso .eurostat.ec.europa.eu/nui/show.do (4 September 2020); graph made by authors.

were nonetheless inspired by the greater performance of other EU countries [64]. Alongside domestic concerns such as decreasing landfill capacity, the strategy prominently refers to the much higher recycling rates in countries such as the Netherlands and Austria [64]. It observes the economic and environmental advantages of resource recovery and notes also the 10- to 15-year time frame needed to accomplish a transition to a recovery-based residue management. Routed in a pro-growth approach to sustainable development, the challenge for the Labour government elected in 1997 was decoupling economic growth from environmental impact. The waste hierarchy, introduced by the 1975 Waste Framework Directive, became and remains axiomatic to UK waste policy at both the national and local levels [10, 65–67]. However, prior to 2000, there had been little change in attitude to waste since the 1982 mission statement of the then Department of the Environment that included “To dispose of wastes of all kinds in an acceptable way.” [68, p. 62]. As shown in Figure 3.3, in 1995 the UK was far behind the then other members (EU 15) in term of landfill reliance, having framed convenient and cheap landfill in empty quarry sites as a socially and sustainable beneficial waste management approach [10]. The landscape change that paved the way for the promotion of resource recovery in the EU was the globally heightened awareness of environmental impacts in the 1970s, which happened to coincide with the UK’s accession to the EU. However, the early years of UK membership were marked by tension and adjustment between the UK’s more flexible, local approach to environmental regulation, and the EU’s more rigidly enforced principles and nascent approach to environmental

59

Table 3.4 Overview of historical contextual factors for local authorities, including primary EU waste-related regulation, English waste policy initiatives, and other major events in the organization of local government (made by the authors from EU and UK/English regulations and policy documents).

Era

Time

EU regulation

Waste-related policy and related changes in England

Public health agenda

1870s+

1875: LA gain responsibility to remove and dispose of household waste; incineration widely used

Austerity agenda

1930s

Early incinerators phased out; increasing reliance on landfill: cheap and argued to be beneficial (restoring quarried landscapes)

1948 Environmental protection and resource security Rising awareness internationally of environmental impact; UK pollution incidents

1970s

Efficiency drive

1980s

Relevant changes to local government

Urban and rural sanitary authorities established under direction of the Local Government Board

Planning authorities UK joined the EU 1975 EU Waste Framework Directive – waste hierarchy introduces idea that “waste” is not homogeneously destined for disposal 1989 Groundwater Directive: landfill operators (and others) liable for groundwater pollution

Waste regulations follow high profile pollution-related incidents: 1974 Control of Pollution Act – introduces definition of waste

1974 Waste disposal and collection authorities created: Plan for quantity and type of waste, and license disposal sites; recovery optional 1979: Conservative government Privatization of LA services: compulsory competitive tendering – based on cost; Establishment of “arm’s length” operations; Reporting to national government required on indicators on multiple indicators

Low carbon agenda 1990s 1991 Waste Framework International agreements Directive translated in EU/national targets 1994: first Producer Responsibility Directive, for Packaging

1992: designation of unitary authorities 1990 Environmental Protection Act Establishes legal definition of waste in the UK under the Local govt act 1992: Private finance initiative to Duty of Care regulations: holder of waste facilitate and later require private retains responsibility for its safe funding contribution to public disposal/recovery after passing to another infrastructure operator 1994 Waste management licensing regulations 1995 Environment Act Requires national waste strategy (implementing EU requirement)

1999: Landfill Directive: Targets for biodegradable waste diversion; operational standards

(continued)

1996: Newly formed Environment Agency takes over regulation of waste facilities from local authorities 1996: Landfill tax UK’s first environmental tax 1998: Waste Minimization regulations: collection/disposal authorities can invest in measures to reduce waste generation, with possibility of funds from government

Table 3.4

(Continued)

Era

Time

EU regulation

Waste-related policy and related changes in England

Stepping up sustainability 2000: Waste Incineration Directive – sets emissions limits, ending lifetime of incinerators too old to retrofit

2008: Waste Framework Directive: mandates separate collection of recyclable waste streams; recycling targets 50% by 2020

2000 Waste Strategy: frames waste as sustainable development issue; need to delink from economic growth and avoid carbon emissions associated with landfill. Not just national government responsibility: Sets performance targets for local authorities; public cooperation required; also forewarns industry of producer responsibilities. DEFRA formed 2001 2002 Waste strategy: LAs no longer just safe keepers of waste but need strategies for diversion with targets to meet. Performance targets for recycling/composting customized to existing performance levels. Waste as a resource; notes high recovery performance in EU. 2004 Landfill Allowance Trading Scheme for biodegradable waste 2003: Waste Implementation Fund, including advice and funding for LAs and new technology demonstration fund 2006: Waste Infrastructure Delivery Programme – aid in large project procurement 2007: Waste strategy for England – ramping up efforts on earlier strategy 2010 Waste management licenses replaced by Environmental Permits

Relevant changes to local government

1997 “New” Labour Government 1999: Best Value – procurement based on performance as well as cost; eight waste indicators, some used as targets A range of regional bodies and processes created e.g.: Regional Development Agencies Regional Spatial Strategy – co-operation on waste planning between authorities Regional Assemblies Government Offices in the regions

2008–2010 National indicators – reduced list of categories; data still collected, but not as targets

Austerity agenda

2010 Industrial emissions?

2011: Kerbside collection of recyclables required; “separate” collection of two material streams 2012 Waste strategy for England 2007 2013 Abolition of Landfill Allowance Trading Scheme for biodegradable waste

Circular economy

2015 2015: Circular Economy strategy

2010 Coalition government (Conservative and Liberal Democrat) Funding cuts, including government support for PFI: Cuts to DEFRA, EA and LAs 2013 Regional Development Agencies abolished, along with other regional scale initiatives 2015 Conservative government

2016: “Brexit” vote with ongoing 2018: Resources and Waste consultation opens: uncertainty 2018: Waste Framework Directive, references “circular suggests homogenization of LA collection systems; and match between economy”

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

700 kg waste per capita

600 500 400 300 200 100 2017

2016

2014 2015

2011

2012 2013

2010

2009

2008

2007

2005 2006

2004

2001

2002 2003

1999

2000

1997 1998

1996

0 1995

64

Year UK waste to landfill

UK waste incinerated

UK waste recycled

UK organic waste recovered

EU 15 waste to landfill

EU 15 waste generated

Figure 3.3 UK management for municipal waste shown as stacked columns; height of column comprises total per capita waste in kg arising by year and can be compared to the upper line which shows the corresponding figure averaged for EU 15 (i.e. countries who were EU members in 1995). Thus since 2006 per capita municipal waste generation in the UK has been less that of the EU 15 average, though the difference has not varied much since 2012. The UK per capita landfill figure had converged with the EU 15 average by 2015. According to the Europa website: “Municipal waste consists mostly of waste generated by households, but may also include similar wastes generated by small businesses and public institutions and collected by the municipality.” Source: Data from http://appsso.eurostat.ec .europa.eu/nui/show.do (4 September 220); graph made by authors.

governance [68, 69]. The UK House of Lords finally accepted in 1984 that the EU had a legal right to pass environmental laws [67] but the UK continued to develop its own approach [70]. Following the transition analyzed herein, the UK landfill rate had markedly dropped to 50% by 2009 [63]. In comparison however by this point Austria, Germany, the Netherlands, and Sweden were sending less than 10% of their domestic waste to landfill (albeit three of the four had higher rates of per capita waste generation than the UK) [63]. In addition, the more ambitious 2008 EU WFD recycling targets had been established, requiring a combined total recycling and composting of municipal waste at 50% by 2020 [71]. Increasingly stringent targets for resource recovery, amendments to Producer Responsibility targets covering the most damaging industrial waste streams, and the more ambitious circular economy policy [2] collectively emphasize that further progress is needed.

3.5.2

Influences in the UK Context for LAs

Against the backdrop of the UK’s evolving relationship with the EU, and the increasingly stringent environmental targets for resource recovery, we now consider the impact on English local authorities in the national context. We note how both waste

3.5 Results and Discussion

services and LA operations significantly altered between the UK joining the EU (1973) and the publication of the 2000 strategy, due to important changes in the institutional context of waste management including the contracting out of waste services and the formation of unitary authorities [47, 72, 73]. Combined with the effects of improved waste technologies and a strengthening of environmental institutions such as the Environment Agency, English LAs were facing a continually changing institutional landscape (Table 3.4). By the late 1990s, after a decade of compulsory competitive tendering of waste services,3 there was a greater diversity of institutional arrangements as well as geographic circumstances, between LAs. Contracting out of waste services (i.e. the retention of responsibility by the LA but the opening of a previously public service to private sector delivery) had not been uncommon prior to the 1980s (Industry interviewee A) [72]. However, forcing local authorities to justify not contracting out meant that they entered into contractual relationships with private sector companies on an unprecedented, but not universal, scale [74]. Whilst some LAs remained responsible for their own waste collection or disposal (as the case may be), albeit via an arm’s-length company (owned by the LA but treated as a separate entity); others contracted with private companies. This variability in service delivery was superimposed on different levels of financial resources, across diverse geographical and demographic challenges between high-density urban landscapes and widely dispersed rural populations. The experience to the householder was not essentially altered, although one local authority representative s suggested that the threat of external competition brought a new focus to considerations of service quality. Local government responsibilities were dramatically redrawn with the formation of unitary authorities in 1992. Most of the interviewed authorities are of this type, which combine the responsibilities of waste disposal authorities (originally county councils) and waste collection authorities (district or metropolitan boroughs within the jurisdiction of a county). Waste authorities were defined in the 1970s with the legal ability, but not requirement, to recover value from waste. Geographically, LA boundaries also shifted in the mid-1990s: for example, the county of Humberside (with a recycling rate of just 2% in 1995), subdivided in 1996 to the Kingston upon Hull, East Riding of Yorkshire, and North and North East Lincolnshire. This reshaping, together with the formation of unitary authorities, meant a redefinition of services, with a disposal contract inherited from before the change, but collection services requiring reorganization. Interviewees appreciated the ability that their unitary status gave them to design services (whether in house or not) without having to coordinate with other district councils or a separate disposal authority. Some authorities chose to work together, as in the cases of York (UA) and North Yorkshire (waste disposal authority), and Hull and East Riding (neighboring unitary authorities). Compulsory reporting on a range of statistics (whether offered separate collections, total percent recycled, landfilled or energy from waste) began 3 Policy of the Thatcher government to open public services to private sector delivery; local authorities had to package services for external bids, albeit they could bid themselves via privatized versions of their own service departments. These “arm’s length” companies could be fully owned by the LA but were treated as separate entities (see Davoudi [74] and Davies [73]).

65

66

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

in 1995 (potentially to comply with the EU requirement, and then used as the base line for subsequent targets) but introduced an element of transparent competition and emphasized the importance of demonstrating “value for money” between the often very different LAs. The statistics’ importance was re-emphasized by, and also enabled, the UK’s first environmental tax, on landfill, introduced in 1996, and the Landfill Allowance Trading Scheme (LATS) which we discuss in more detail in Section 5.3. Yet despite these actions, landfill amounts continued to rise until 2001 (Figure 3.2). Besides targets for diverting biodegradable waste, the EU Landfill Directive also instituted operational standards for landfills, which were ahead of the UK norm and ended the practice of co-disposal of hazardous and nonhazardous waste. This built on the introduction of waste management licensing (1994), originally under the remit of waste disposal authorities (who were sometimes regulating their own arm’s-length companies) but later shifted to the EA. Higher technical and regulatory standards led to a rapid consolidation of landfill sites as smaller sites became unviable (Industry A). In addition, the Waste Incineration Directive (passed in 2000, and replaced in 2014 by the Industrial Emissions Directive) imposed emissions limits that were beyond the technical compliance capacity of the UK’s older energy from waste facilities. These examples illustrate how the UK’s long-standing practices were being adjusted by the EU’s perspective of the waste landscape, particularly encouraging a professionalization of the UK waste industry. The increased professionalization and technological improvements could not have been achieved without capital investment, frequently through use of the Private Finance Initiative. Introduced in 1992, this policy first allowed and then required private investment in public infrastructure, facilitating the development of capital-intensive technology such as energy from waste. Although a detailed examination of the growth of incineration as an alternative to landfill4 and the resource implications of this technology choice is not possible here, it is notable that there is a sociocultural and historic influence to responses within a given landscape. Local authorities and the waste industry had to deal with that filtered landscape through the reality of planning applications and public response, which was hostile to energy from waste. The later withdrawal of central government support from PFI under the austerity regime of the Coalition government (2010–2015) caused complications for interviewees in both the public and private sector involved in contract negotiations at the time. Expensive, long-term, investment decisions are therefore vulnerable to national political decisions.

3.5.3

Implementation of the 2000 Waste Strategy

Into this tumultuous scenario the 2000 Waste Strategy arrived, bringing further change for local authorities. To promote the government’s aim to decouple economic growth from environmental impact, LAs were for the first time required to 4 Conversely, some EU countries have taxed incineration as well as landfill in order to drive up recovery.

3.5 Results and Discussion 100 90 80

Per cent

70 60 50 40 30 20 10 0 2000 01 2001 02 2002 03 2003 04 2004 05 2005 06 2006 07 2007 08 2008 09 2009 10 2010 11 2011 12 2012 13 2013 14 Year ERYC

Hull

NE Lincs

N Lincs

York

England

Figure 3.4 Rates of landfilling of household waste in the case study authorities. Data source: Defra Waste Data Flow statistics; graph made by authors. Source: Data from Defra Waste Data Flow statistics; graph made by authors.

devise strategies to divert from landfill and increase recovery of both biodegradable and so-called dry recycling (including plastic, metals, paper, glass, textiles, etc.). Defra was looking for LAs “to put in place local strategies for sustainable management of municipal waste” [64, p. 45], to be backed by a range of policy instruments (Table 3.4). These were largely financial instruments allowing some flexibility of approach, with a trust in adjustments to market signals rather than direct requirements (e.g. taxing landfill, rather than banning other than materials mandated by the EU). 3.5.3.1 LA Implementation of Waste Policy

The changes to landfilling rates for the case study authorities over this time are shown in Figure 3.4. The dramatic drop in NE Lincolnshire during 2004 is due to the energy from a waste plant coming online, with rates of landfilling continuing to drop to near-negligible rates by 2013−2014. The other authorities show more gradual rates of decline across this time period, with N Lincolnshire’s and York’s landfill rates reaching low points in 2009–2010 and 2011–2012 respectively and increasing slightly since. The combined recycling and composting rates over time for the same authorities is shown in Figure 3.5. Overall, they mirror the national performance, transitioning from barely 10% of household waste being recycled in 2000−2001, to (in four out of the five cases) exceeding 40% recycling by 2010/2011. However, only East Riding of Yorkshire and Hull remain on an upward trajectory at the end of this time period. York reached a level ahead of the national average by 2007−2008, but has not progressed significantly since. North East Lincolnshire and North Lincolnshire peaked in 2010−2011 and have both since seen rates decrease. North Lincolnshire was the highest achieving of the four throughout most of this period, and still met the national average even when their rates stagnated. However, NE Lincolnshire has

67

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

70 60 50 Per cent

68

40 30 20 10 0 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 01 02 03 04 05 06 07 08 09 10 11 12 13 14 Year ERYC

Hull

NE Lincs

N Lincs

York

England

Figure 3.5 Rates of household combined waste recycling and composting for the interviewed unitary authorities compared to the rate for England. Source: Data from Defra Waste Data Flow statistics; graph made by authors.

never reached the national average and has from 2010 to 2011 onward been diverging from it. For most years it is possible to separate the data for non-organics recycling from composting. Comparing Figures 3.6 and 3.7 it can be seen that these two management techniques account for roughly similar proportions of household waste in these authorities. However, ERYC (rural authority) and N Lincs are the best performing for organic waste, whilst Hull the highest performing for recycling. The failure to maintain the increase in combined wet and dry recycling in N Lincs is shown by Figure 3.6 to primarily be the result of a reduction in the proportion of household waste being composted. Given the overlap in the incidence of the various policy mechanisms over this time period (Table 3.4), it is very difficult to separate out influences simply by looking at the figures. The rate of progress increased from 2003 to 2004, coincident with or just before the beginning of LATS in 2004, which may have encouraged more authorities to undertake separate collection of organic waste. Unfortunately, composting and dry recycling rates were not collected separately for 2005−2006, but by the following year compositing rates were on an upward trajectory in all these authorities, including those for which rates had previously remained very low (i.e. ERYC, York). One can compare Hull (dense urban population with significant elements of deprivation) with East Riding (relatively low density, rural population with on average more wealthy population). For East Riding, composting accounts for more than half the combined recycling rate. For Hull composting rate has fluctuated at around 16% since 2010−2011, while dry recycling has increased from 25 to 34% over this time (supported by introduction of alternate weekly collections of residual waste in 2013) and accounts for 2/3 of the combined. Conversely ERCY dry recycling

3.5 Results and Discussion

35.0 30.0

Per cent

25.0 20.0 15.0 10.0 5.0 0.0 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 01 02 03 04 05 06 07 08 09 10 11 12 13 14 Year ERYC

Hull

NE Lincs

N Lincs

York

Figure 3.6 Rates of household waste composting for the interviewed unitary authorities. Source: Data from Defra Waste Data Flow statistics; graph made by authors. 40 35

Per cent

30 25 20 15 10 5 0 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 01 02 03 04 05 06 07 08 09 10 11 12 13 14 Year ERYC

Hull

NE Lincs

N Lincs

York

Figure 3.7 Rates of household waste recycling for the interviewed unitary authorities. Source: Data from Defra Waste Data Flow statistics; graph made by authors.

stagnated from 2006−2007 to 2011−2012, and then jumped up again in 2013/2014 following the introduction of alternate weekly collections of residual waste. Evidently the plateauing of the English recycling average is not simply a question of reaching technological limits, but is masking a divergence between authorities who continued to progress and those who have not been able to keep up rates achieved under the more supportive national conditions prior to 2010. In the coming sections, therefore, we draw on the interview date to understand the influences.

69

70

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

3.5.3.2 Targets

Changes to LA data reporting requirements were a significant part of the national government’s drive to monitor and improve waste management. Although the Labour government elected in 1997 did not attempt to reverse the previous governments’ reform of local government (and continued the program of privatization), they did make adjustments which reflected a new determination for implementing sustainable development (1999; albeit criticized subsequently for being stronger on rhetoric than action [2004]). The introduction of the principle of best value (1999) continued the theme that public services should offer value for money, but removed the assumption that value for money would be best served by the private sector (taking the compulsory out of CCT), and allowed criteria such as performance and community opinions to be considered [73, 75]. Whilst this restored some discretion to LAs in terms of service delivery, the best value program was very much a target-driven initiative. Data reporting from LAs to central government was considerably expanded and some indices, including the waste-related targets in the 2000 strategy, were given the regulatory force of statutory targets. In 2007 the National Indicators were introduced (Audit Commission), replacing the previous Best Value Performance Indicators. This marked a reduction in the number of indicators to be reported to national government from eight to three, with the those remaining being central to meeting Defra’s policy goals (per capita waste production; waste landfilled and a combined figure for recycling and composting) [76, 77]. The targets no longer had a legal implication, but were used until 2012 to assess performance against the Landfill Allowance Trading Scheme (see below). Reporting of waste data continues (partly in response to “user demand”), and continues to evolve, which somewhat limits the utility of what at first sight is an extraordinary 25-year time series relating to household and local authority waste practices. The interviewed authorities were pragmatic about the reporting requirements, as their departments were not responsible for funding the associated positions within their authority. They could obtain the necessary information fairly readily from contractors or companies to whom recyclates were sold (commercialization of the waste sector here being beneficial); and also used the data for monitoring themselves. Hull and East Riding set themselves a voluntary higher target in response to public enthusiasm for recycling (at least in the abstract). This focused the minds of local politicians, who did not want to miss their own targets, and may have fostered a sense of local competition as seen in the case of Italian local waste management [78]. Cost implications via the landfill tax (discussed below) proved to be far more significant than the targets per se. According to a Defra interviewee (1), the targets were seldom enforced by more than the minister explaining the importance of the targets. 3.5.3.3 Financial Instruments

Against this backdrop of national government steering LAs with a more flexible approach, two very firm market-based policies must be noted, both explicitly focused on waste. First, Landfill Tax, introduced in 1996 at £7/tonne for active waste [79]

3.5 Results and Discussion

which was aimed explicitly at incentivizing waste diversion from landfill, as well as raising funds for the treasury. Local authorities were explicitly liable for the tax [79]. In 1999 the then government introduced the “landfill tax escalator” which initiated the annual increase in the tax.5 At this time, explicit reference was made to the Landfill Directive target for diverting biodegradable waste, in addition to the pending UK national waste strategy [79]. The further acceleration of the escalator in 2008 came as a surprise to LAs, and was no longer offset by other funding mechanisms [80]. The Local Government Association saw this as a change to a purely fund-raising mechanism, as they felt there was insufficient time to develop alternative technologies [80]. Conversely, some in the waste industry would have welcomed the introduction of high tax levels much sooner, to increase the financial viability of alternatives to landfill (WI 8). The drive to avoid landfill thus arguably took insufficient account of the financial realities of developing alternatives, such as energy from waste, or anaerobic digestion and the timescales required to raise finance and obtain planning permission for changes in land use. At the time, circular solutions, which are now receiving research investment to industry and academia, depended on the less certain influences of the producer responsibility directives [81]. The second financial mechanism introduced was the Landfill Allowance Trading Scheme (LATS) which began April 2004 and lasted until 2012. Along the lines of carbon trading schemes, the idea was that LAs who did not use their full allowance for landfilling biodegradable waste could sell the excess to local authorities who exceeded theirs. The allowances were set at levels so that the UK as whole would meet its EU targets. Doubts were expressed about the effectiveness of LATS to bring about the development of non-disposal technologies [47]; it gave a significant advantage to LAs who included energy from waste in their waste management portfolio. Whilst on the one hand waste targets were being made aspirational, rather than statutory, on the other, LATS imposed strict limits (rather than targets) which came with a punitive mechanism for those not complying. Notably, LATS refers only to biodegradable waste, while the Best Value Performance Indicators and National Indicators included other forms of recycling. Opinions were divided among interviewees as to whether LATS or the Landfill Tax or the targets were the most effective mechanism – perhaps reflecting at what rate their preparations to divert waste from landfill had progressed. According to Defra 2, “when [LATS] came it was really good, it gave a kick-start to local authorities but yes I mean it’s now been overtaken by the landfill tax, local authorities see that more as the sort of driver for them to reduce waste going to landfill. But certainly the impression we’ve been given by local authorities is that LATS at the time was a sort of . . . .real kick-start that they needed.” The termination of LATS was more an indication that the policy was no longer needed, rather than that the intention no longer applied. Reliance on “soft reinforcement” and market-based instruments 5 The tax increased by £1 per year up to and including 2004; subsequently increased to £3 per year up to and including 2007; the rate of £8 introduced in 2008 has been continued by the new government, with the pledge that this will last until 2014, with the then rate of £80 per tonne being the minimum until at least 2020.

71

72

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

typifies the UK’s reliance on these policies and the involvement of private industry in this sector. Significantly, the policies in place were not only about setting challenges for LAs, measuring progress and potentially punishing. The support and guidance in place were also extensive. Interviewed authorities reported helpful collaboration with WRAP (government funded body established to promote recycling, e.g. by research into building markets for recycled material). WRAP started its “Love Food Hate Waste” campaign in 2007, as an early example of an effort to reduce waste rather than tidy up after it. Interviewees commented on how publicity campaigns were funded by grants from central government, with the key period from 2000 to 2010 characterized by a constructive relationship whereby central government was prepared to support the achievement of higher standards with relevant financial and organizational support. 3.5.3.4 Regional Governance

Another Labour Government policy discontinued by the post-2010 Coalition Government was a regional scale of governance. Of particular relevance to waste management was the regional spatial strategy, later the single regional strategy [82]. These plans, to be produced in co-operation with local government and other agencies, were designed to co-ordinate regional needs and provision across a wide range of policy areas with an ambitious mandate to incorporate the economic, social, and environmental pillars of sustainable development. LAs thus had to co-operate in the production of these regional plans, into which their own more specific plans had demonstrably to fit. The demise of this scale of governance in 2012 met with a mixed response from interviewees. However, whilst one LA representative stated “I can’t believe they’ve done it” (LA 6), a more common view was that the ideal of co-operation had not been achieved in practice. It was seen as difficult for LAs to abrogate the interests of their own area to the benefit of a co-ordination at a larger scale. There was more concern for the reducing presence of national government in the regions. The national “Government Office” in the region has been used as a means to bring national government bodies, such as Defra, into contact with LA representatives. This was seen as a useful source of information on policy developments: “and we used to get heads up on potential policy changes and it was an opportunity for us to feedback so that worked quite well” (LA 3). However, the institutional context for this forum has disappeared emphasizing the contingency of the regional scale, which is not a dependable long-term proposition for sustainability initiatives [83].

3.5.4

Local Authorities and the Public

Public participation is essential for the recovery of post-consumer waste. In this section we look at how that has changed and the constraints on LAs imposed by the public. As stated above, although the quality of the service may have varied over time, from 1875 to the early 1990s there was little significant change in the relationship

3.5 Results and Discussion

between LA and residents. Residents put out items they did not want; the LA took them away – even to the point of going to the rear of houses to retrieve items. The first change to this was the appearance of the wheeled bin, introduced in the case study authorities in the mid-1990s. As wheeled bins did not require lifting, householders rather than the collection staff could be expected to move them from their storage location to the kerb edge and back. This reduced the time taken to complete collections, as well as minimizing the contact of waste collectors with rubbish. Wheeled bins were introduced on a gradual basis, at first meeting with resistance, but subsequently demanded by residents who had yet to be given one. Although the motivation for introducing the wheeled bin was to increase the speed of collection, once established, there were both positive and negative implications for waste diversion. Wheeled bins shift some of the burden of the waste collection service on to the residents, but also give LAs a mechanism by which to limit the amount of waste collected, by ensuring that residents had a uniform receptacle. Some LAs have taken a very firm line on not collecting waste additional to the capacity of the bins provided, insisting bin lids must be shut. There is a perception, at least, among LA staff that residents will fill whatever size of bin they have – the size of bin effectively marking both the maximum and minimum volume of waste likely to be presented for collection. A step beyond strictly enforcing the capacity of the bin, or even replacing them with smaller bins, is reducing the frequency with which the bins are emptied. The withdrawal of the weekly bin collection has brought the wrath of local residents upon the local authorities that attempted it. Residents saw the weekly collection as a service they had paid for, and denying it was portrayed as a health hazard. Typically, however, the displeasure is short-lived and recycling rates significantly increase [84]. This has also been the experience of the interviewed authorities. One reported “previously if you launched a new scheme and you still had weekly collection they’d do it for the first two months and then kind of get out of the habit, because it’s easier not to bother.” (LA 5). Reducing the frequency of collection, however, sets a new benchmark for acceptable levels of waste. Coupled with the cost savings implied by reducing the frequency of residual waste collection, the switch to alternate weekly collections becomes commonplace. None of the LA representatives interviewed made any reference to national government pressure on them to switch to alternate weekly collections. The Defra representative emphasized LA autonomy in waste decision-making. There was, however, a plethora of guidance on waste matters from national-level institutions and LAs were generally positive in their opinions of the support available. Another important change to the LA–resident relationship has been the instigation of kerbside recyclable collections. Kerbside collection of recyclables had been at the discretion of local authorities, but from 2010 they have been obliged to offer at least two kerbside collections, i.e. residual waste and one recyclable material, or residual waste and comingled recyclables. In this way an informal practice has been made a requirement. The national government was for the first time stipulating a collection practice and thereby in theory directly impacting on the relationship

73

74

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

between LAs and their residents. In practice, however, there was little impact on the LA–resident relationship, as most LAs were already exceeding the requirements of the law the time it came into force in 2010 [63]. The reduction in frequency of residual waste collections (i.e. nonrecyclable, and/or potentially recyclable but unsegregated waste) helps to support the cost of kerbside recycled waste collection. If recycling is an accepted social norm, kerbside recycling is a significant improvement to waste collection services, i.e. compared to householders having to take their own bottles to a bring site. But if recycling is not considered a desirable for its own sake, this appears as an imposition by the LA on residents. Residents were encouraged, though not required, to segregate their waste at source. In the authorities studied, as others the length and breadth of the UK, much effort has gone into trialling different schemes (different combinations of materials collected, together, separately, hand or machine sorting, size of bins) with accompanying publicity campaigns and levels of support for householders up to and including door-to-door visits [41, 42]. Presorting of waste is highly influential in determining its value, and having the public do that is cheaper than employing staff or investing in technology. Thus, the public can be faced by a complexity of different colored bins to be placed by the kerbside at fortnightly to monthly intervals. Hull and East Riding experimented with different approaches before settling on something very similar with comingled dry recycling, green waste, and residual waste collections. Another case study authority experimented with charging for the collection of organic waste as an extra service. Uptake was low, and with customers widely distributed over a rural area the service was not viable. No other LAs interviewed had attempted charging (suggested as an option by the 2002 strategy). This is too far out of the norm for UK public, and overambitious as a waste policy, given the risk of residents illegally dumping waste (fly tipping) to avoid charges.

3.5.5

Legacy of the Strategy

The initial view of the 2000 Waste Strategy would have to be that it was successful – given the dramatic fall in waste to landfill and increase in recycling. Rates have largely, but not invariably, held up since the removal of most of the policies that supported them, suggesting either that the landfill tax sent such a strong signal other measures are insignificant, or that recycling has impressed itself on the public. Even through the austerity era since 2010, recycling services have been maintained. Vulnerability of the standards is shown by comment of senior figure at one of the (on average) more prosperous unitary authorities who commented that the statutory targets and financial mechanisms helped him defend waste services against other council priorities. However, rates of recovery have not increased significantly since the time period studied herein, though the highest achieving authorities in the UK had reached excess of 60% by 2017 [84]. It has to be considered, though, that recycling rates presently include organics, which are relatively straightforward to process compared to other materials, e.g. plastic. Notwithstanding the tremendous improvement to recycling, the government’s drive away from landfill is the most successful element of the policy. Landfill rates

3.6 Conclusions

continue to fall, whilst recycling rates have stagnated, even with the inclusion of organic waste [9, 12, 84]. Interviewees (urban UA and urban collection authority) referred to the diminishing returns on investments to improve recycling rates once a 40% level had been reached. UK had met its biodegradable waste diverted from landfill targets for 2020 by 2015 [84]. It has met or exceeded targets for packaging, batteries, although there are signs of strain as packaging waste recycling in 2014 was less than 2013 [85]. Over the last couple of years at the national level, as well as one or two of the authorities herein, recycling has taken a slight backward step and the prospects for the 50% 2020 target are looking slim (depending on the extent to which packaging figures feed into that target) [9, 85]. A step change in technology, accompanied by a more “circular” vision, may be required rather than the previous more incremental changes. There is a view (particularly from the waste industry) that the UK has achieved as much as it can with current waste infrastructure. Something more radical will be needed which could include advanced technology and may be spurred on by more stringent approaches and research attention to plastics [4]. The landfill diversion as mainstay of policy appears to have run its course and the 2018 waste and resources strategy may have come just in time with consultation around targets for inclusion of recycled plastic, uniformity of waste collection, and ensuring that labels match service availability [4, 6]. The implications for policy makers are that policies are most effective when there is some alignment across different areas of activity. Support and guidance are needed to back up targets and financial instruments. The public can become adapted in a surprisingly short amount of time, although not everyone has become an avid recycler.

3.6 Conclusions This chapter has analyzed the transition in household waste practices associated with the 2000 Waste Strategy from the perspective of local authorities responsible for collecting and/or disposing of household waste. We suggest that a transition is a social change that with hindsight can be analyzed in historical and geographical context in order to understand the causal factors involved and the underlying influences, which may not be immediately apparent. The time span over which a transition occurs can only be identified in retrospect and the selection of period of study will involve some arbitrarily defined boundaries. However, a transition is a useful concept to identify an event worthy of explanation, and through which to learn lessons for the future. The causal factors emerging, though, will depend on how/where you look in terms of spatial/temporal scale chosen for the study. Thus, using our extended timeframe the regime emerges as far less fixed than implied by previous studies of transition, and subject to continuous change on a political timescale rather than comprising an institutional barrier to change. Although embedded in the cultural/social appreciation of the environmental landscape of the time, from the LA perspective, the landscape comes strongly filtered by government policy. The landscape cannot be considered to be outside of the

75

76

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

“system.” One could say that LAs had undergone two decades of reform prior to the publication of the strategy, but it would be better to say that reform is continuous, as it has continued up to the present. The UK in the 1970s was experiencing the effects of deindustrialization in the context of globalization, without the benefit of hindsight to understand the magnitude of economic restructuring implied (and continuing today). The whirlwind of the Conservative government’s (1979−1997) neoliberal response caught up LAs and effectively created the present waste industry. The Labour government, taking a somewhat more engaged approach to sustainable development, in enacting the policies herein were prepared to invest some of the gains from positive economic performance of the time into environmental measures (with the assumption of protecting future growth). These fluctuations played out against a scenario of convergence with the EU and rising evidence for and political support for Climate Change. Policies serve to harness and direct that ability to change in a certain direction, and the range of support mechanisms, as well as financial signals, were a critical aspect of changing practices, as also seen for bringing about changes to energy practices in the UK [31]. But at least for the LA actors involved change (e.g. responding to signals from government) is normal; for national government figures bringing about change is their job. A narrow consideration of waste policies would not capture the institutional changes, or the challenges of reorganization. The starting point of the Waste Strategy is just a convenient analytical tool. Focusing solely on 2000–2010 would be entirely within the frame of one party that could simply be termed “the government.” With our longer more porous temporal framework, the difference in emphasis between the Conservatives, Labour, and the Coalition governments emerge setting different policy signals despite the common backdrop of increasing environmental standards. The entire period considered could be seen as one anti-landfill policy “paradigm” [17], but nonetheless with significant variations within it. And although the 2010 change in government brought some notable policy shifts, it is probably too soon to tell whether it marks a significant moment. The political focus on climate change and the perception of the circular economy as a potential solution [5] may mean that the plateau in recycling rates since 2009 proves short-lived. Either way, LAs will continue to face changing requirements and contexts, coping with which appears to be a routine matter. This study therefore indicates that the political element in transitions referred to by Meadowcroft [18] is significant, and furthermore taking a long view of the time span and broad view geographically is necessary to reveal the underlying influences. Too prescriptively defined boundaries would predetermine where one looked for factors and therefore what was found [34]. As well as missing regional to local variations, a national focus to transition studies gives a false impression of the absoluteness of the landscape. Typically, the landscape is the least examined part of the MLP. Understanding the wider context of English waste policy as in this study introduces supranational variations. Even given the common rise in concern for the environment, increasing understanding and political acceptance of climate change, a couple of decades were required for the UK to embed within the EU practice and approach to policy [54]. The social and

References

political construction of the landscape is revealed by the wider analysis. The impact of Brexit on the UK’s view of the environmental landscape (presently dominated by concerns for biodiversity and plastic pollution) remains to be seen. Further research needs to take a similarly historical and geographically routed view of other stakeholders in the evolution of waste management; and indeed, of other forms of transition. What is the experience of change for national government, the waste industry and the public? As Ollman [34] discussed, different spatial and temporal scales of analysis reveal different underlying factors as causal mechanisms. Analysis of waste management practices other than spatial and temporal scales may also reveal other significant causal factors in bringing about or hindering change (perhaps differently in different places). What has been the influence of the local or supranational scales on transitions that have been identified at national levels? Furthermore, as attention in resource management shifts to preventative strategies in the context of a circular economy [5], the definition of relevant stakeholders is much wider. Industry becomes illuminated as an influential agent (not the passive producer of goods that the public purchase and subsequently hand LAs to dispose of). The national context becomes more problematic, given the international nature of supply chains. If goods designed with “built in circularity” rather than “built in obsolescence” emerge, the resulting drop in household waste production may render recycling rates much less noteworthy.

Acknowledgements Many thanks to all interviewees and others who have given their time to this project. Thanks to Graham Ferrier for the maps. We also thank the reviewers for their comments. This chapter was partially supported by EPSRC grant EP/S025537/1 “Evolving a circular plastics economy.”

References 1 Ellen MacArthur Foundation (2012). Towards the Circular Economy Vol. 1: An Economic and Business Rationale for an Accelerated Transition. https://www .ellenmacarthurfoundation.org/publications/towards-the-circular-economy-vol1-an-economic-and-business-rationale-for-an-accelerated-transition (accessed 15 December 2019). 2 European Commission (2015). Communication from the Commission to the European Parliament, the Council, the European Economic and Social Committee and the Committee of the Regions. Closing the Loop – An EU Action Plan for the Circular Economy. COM/2015/0614 final. EC, Brussels. 3 Tibbs, H.B. (1991). Industrial Ecology – An Environmental Management Agenda for Industry. Arthur D Little, Inc. 4 European Commission (2018). Communication from the Commission to the European Parliament, the Council the European Economic and Social

77

78

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

5 6 7

8

9 10 11 12 13

14 15

16

17 18 19

20

21 22

Committee and the Committee of the Regions. A European Strategy for Plastics in a Circular Economy. COM/2018/028 final. EC, Brussels. Geissdoerfer, M., Savaget, P., Bocken, N. et al. (2017). The circular economy: a new sustainability paradigm? Journal of Cleaner Production 143: 757–768. Department for Environment, Food and Rural Affairs (2018). Our Waste, Our Resources: A Strategy for England. London: Defra. Deutz, P., Neighbour, G., and McGuire, M. (2010). Integrating sustainable waste management into product design: sustainability as a functional requirement. Sustainable Development 18: 229–239. Jalil, E., Grant, D., Nicholson, J., and Deutz, P. (2016). Reverse logistics in household recycling and waste systems: a symbiosis perspective. Supply Chain Management 21 (2): 245–258. Department for Environment, Food and Rural Affairs (2016). Digest of Waste and Resource Statistics. London: Defra. Department of Environment, Transport and the Regions (2000). Waste Strategy 2000 for England and Wales. London: DETR. Deutz, P. and Frostick, L. (2009). Reconciling policy, practice, and theorisations of waste management. The Geographical Journal 175 (4): 247–250. Levidow, L. and Raman, S. (2018). Metamorphosing waste as a resource: scaling waste management by ecomodernist means. Geoforum 98: 108–122. Geels, F.W. (2004). From sectoral systems of innovation to socio-technical systems: insights about dynamics and chance from sociology and institutional theory. Research Policy 33 (6-7): 897–920. Geels, F.W. (2010). Ontologies, socio-technical transitions (to sustainability), and the multi-level perspective. Research Policy 39 (4): 495–510. Marsden, T. (2013). From post-productionism to reflexive governance: contested transitions in securing more sustainable food future. Journal of Rural Studies 29: 123–134. Bailey, I. and Wilson, G.A. (2009). Theorising transitional pathways in response to climate change: technocentrism, ecocentrism and the carbon economy. Environment and Planning A 41 (10): 2324–2341. Hall, P. (1993). Policy paradigms, social learning and the state: the case of economic policy in Britain. Comparative Politics 25: 275–296. Meadowcroft, J. (2016). Lets get this transition moving! Canadian Public Policy 42: S10–S17. Avelino, F. and Wittmayer, J. (2016). Shifting power relations in sustainability transitions: a multi-actor perspective. Journal of Environmental Policy and Planning 18 (5): 628–649. Geels, F.W., Schwanen, T., Sorrell, S. et al. (2018). Reducing energy demand through low carbon innovation: a sociotechnical transitions perspective and thirteen research debates. Energy Research and Social Science 40: 23–35. Giddens, A. (1984). The Constitution of Society: Outline of the Theory of Structuration. USA: University of California Press. Markard, J., Raven, R., and Truffer, B. (2012). Sustainability Transitions: an emerging field of research and its prospects. Research Policy 41: 955–967.

References

23 Affolderbach, J. and Schulz, C. (2016). Green Building Transitions, Regional trajectories of Innovation in Europe, Canada and Australia. Luxembourg: Springer. 24 Geels, F.W. (2014). Regime resistance against low-carbon transitions: introducing politics and power into the multi-level perspective. Theory, Culture & Society: Explorations in Critical Social Science 31 (5): 21–40. 25 Hansen, T. and Coenen, L. (2015). The geography of sustainability transitions: review, synthesis and reflections on an emergent research field. Environmental Innovation and Societal Transitions 17: 92–109. 26 Bergek, A. and Berggren, C. (2014). The impact of environmental policy instruments on innovation: a review of energy and automotive industry studies. Ecological Economics 106: 112–123. 27 Shove, E. and Walker, G. (2010). Governing transitions in the sustainability of everyday life. Research Policy 39: 471–471. 28 Sovacool, B. and Hess, D. (2017). Ordering theories: typologies and conceptual frameworks for sociotechnical change. Social Studies of Science 47 (5): 703–750. 29 Geels, F.W. (2002). Technological transitions as evolutionary reconfiguration processes: a multi-level perspective and a case-study. Research Policy 31 (8–9): 1257–1274. 30 Geels, F.W. and Kemp, R. (2007). Dynamics in socio-technical systems: typology of chance processes and contrasting case studies. Technology in Society 29: 441–455. 31 Kern, F. (2012). The discursive politics of governing transitions towards sustainability: an analysis of the Carbon Trust in the UK. International Journal of Sustainable Development 15 (1–2): 90–106. 32 Markard, J. and Truffer, B. (2008). Technological innovation systems and the multi-level perspective: towards an integrated framework. Research Policy 37: 596–615. 33 Sayer, A. (1984). Method in Social Science: A Realist Approach. London: Hutchinson University Library. 34 Ollman, B. (2003). Dance of the Dialectic. Urbana, IL: University of Illinois Press. 35 Abernathy, W.J. and Utterback, J.M. (1978). Patterns of industrial innovation. Technology Review 80 (7): 40–47. 36 Nakamura, H., Kajikawa, Y., and Suzuki, S. (2013). Multi-level perspectives with technology readiness measure for aviation innovation. Sustainability Science 8: 87–101. 37 Hofman, P. and Elzen, B. (2010). Exploring system innovation in the electricity system through sociotechnical scenarios. Technology Analysis and Strategic Management 22: 653–670. 38 Fuenfschilling, L. and Truffer, B. (2014). The structuration of socio-technical regimes – conceptual foundations from institutional theory. Research Policy 43: 772–791. 39 Kemp, R. and van Lente, H. (2011). The dual challenge of sustainability transitions. Environmental Innovation and Societal Transitions 1: 121–124.

79

80

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

40 Bulkeley, H. and Gregson, N. (2009). Crossing the threshold: municipal waste policy and household waste generation. Environment and Planning A 41: 929–945. 41 Mee, N., Clewes, D., Phillips, P., and Read, A. (2004). Effective implementation of a marketing communications strategy for kerbside recycling: a case study from Rushcliffe, UK. Resources, Conservation and Recycling 42: 1–26. 42 Martin, M., Williams, I.D., and Clark, M. (2006). Social, cultural and structural influences on household waste recycling: a case study. Resources, Conservation and Recycling 48 (4): 357–395. 43 Van Vliet, B., Chappells, H., and Shove, E. (2005). Infrastructures of Consumption: Environmental Innovation in the Utility Industries. London: Earthscan. 44 Bulkeley, H. and Askins, K. (2009). Waste interfaces: biodegradable waste, municipal policy and everyday practice. Geographical Journal 175 (4): 251–260. 45 Atkinson, H. and Wilks-Heeg, S. (2000). Local Government from Thatcher to Blair. Cambridge: Polity Press. 46 Davoudi, S. and Evans, N. (2005). The challenge of governance in regional waste planning. Environment and Planning C: Government and Policy 23: 493–517. 47 Bulkeley, H., Watson, M., and Hudson, R. (2007). Modes of governing municipal waste. Environment and Planning A 39 (11): 2733–2753. 48 Hall, P.A. and Soskice, D. (eds.) (2001). Varieties of Capitalism: The Institutional Foundation of Comparative Advantage. London: Oxford University Press. 49 Hawkey, D., Tingey, M., and Webb, J. (2014). Local Engagement in UK Energy Systems: A Pilot Study of Current Activities and Future Impact. Loughborough: Energy Technologies Institute. 50 Uyarra, E. and Gee, S. (2013). Transforming urban waste into sustainable material and material usage: the case of Greater Manchester. Journal of Cleaner Production 50: 101–110. 51 Abbott, A., Nandeibam, S., and O’Shea, L. (2011). Explaining the variation in household recycling rates across the UK. Ecological Economics 70 (11): 2214–2223. 52 Ehnert, F., Kern, F., Borgström, S. et al. (2018). Urban sustainability transitions in a context of multi-level governance: a comparison of four European States. Environmental Innovation and Societal Transitions 26: 101–116. 53 Waite, S., Cox, P., and Tudor, T. (2015). Strategies for local authorities to achieve the EU 2020 50% recycling, reuse and composting target: a case study of England. Resources, Conservation and Recycling 105: 18–28. 54 Farmer, T., Shaw, P., and Williams, I. (2015). Destined for indecision? A critical analysis of waste management practices in England from 1966-2013. Waste Management 39: 266–276. 55 Gharfalkar, M., Court, R., Campbell, C. et al. (2015). Analysis of waste hierarchy in the European waste directive 2008/98/EC. Waste Management 35: 305–313. 56 Ma, B., Li, X., Jiang, Z., and Jiang, J. (2019). Recycle more, waste more? When recycling efforts increase resource consumption. Journal of Cleaner Production 206: 870–877.

References

57 Fletcher, A. (2017). Applying critical realism in qualitative research: methodology meets method. International Journal of Social Research Methodology 20 (2): 181–194. 58 Wittmayer, J., Avelino, F., van Steenbergen, F., and Loorbach, D. (2017). Actor roles in transition: insights from sociological perspectives. Environmental Innovation and Societal Transitions 24: 45–56. 59 Lawhon, M. and Murphy, J. (2012). Sociotechnical regimes and sustainability transitions: insights from political ecology. Progress in Human Geography 36 (3): 354–378. 60 Pinsonneault, A. and Kraemer, K. (1993). Survey research methodology in management information system; an assessment. Journal of Management Information Systems 10 (2): 75–105. 61 Gillham, B. (2005). Research Interviewing: The Range of Techniques. Buckingham: Open University Press. 62 Tansey, O. (2007). Process tracing and elite interviewing: a case for non-probability sampling. Political Science and Politics 40 (4): 765–772. 63 Department for Environment, Food and Rural Affairs (2011). Waste Data Overview. London: Defra. 64 Cabinet Office (2002). Waste Not, Want Not: A Strategy for Tackling the Waste Problem in England. https://webarchive.nationalarchives.gov.uk/ 20080718162249/http://www.cabinetoffice.gov.uk/strategy/work_areas/waste .aspx (accessed 6 December 2019). 65 European Council (1999). Council Directive 1999/31/EC of 26 April 1999 on the landfill of waste. OJ L 182, 16.7.1999, 1–19. EC, Brussels. 66 Department for Environment, Food and Rural Affairs (2007). Waste Strategy for England 2007. London: Defra. 67 Department for Environment, Food and Rural Affairs (2009). Waste Strategy: Annual Progress Report 2008/09. London: Defra. 68 Royal Commission on Environmental Pollution (1984). Tenth Report: Tackling Pollution: Experience and Prospects. https://webarchive.nationalarchives.gov.uk/ 20110322144015/http://www.rcep.org.uk/reports/10-pollution/1984-10pollution .pdf (accessed 31 July 2019). 69 Sharp, R. (1998). Responding to Europeanisation; a governmental perspective. In: British Environmental Policy and Europe (eds. P. Lowe and S. Ward). London: Routledge. 70 Royal Commission on Environmental Pollution (1988). Twelfth Report: Best Practicable Environmental Option. https://webarchive.nationalarchives.gov.uk/ 20110322144236/http://www.rcep.org.uk/reports/12-bpeo/1988-12bpeo.pdf (accessed 31 July 2019). 71 European Council (2008). Council Directive 2008/98/EC of 19 November 2008 on Waste and Repealing Certain Directives. OJ L 312, 22.11.2008, p. 3–30. EC, Brussels. 72 Domberger, A., Meadowcroft, S.A., and Thompson, D.J. (1986). Competitive tendering and efficiency: the case of refuse collection. Fiscal Studies 7: 69–87.

81

82

3 Coping with Change: (Re) Evolution of Waste Management in Local Authorities in England

73 Davies, S. (2007). Politics and Markets: the case of UK municipal waste management. School of Social Sciences, Cardiff University, Working Paper 95. 74 Davoudi, S. (2000). Planning for waste management: changing discourses and institutional relationships. Progress in Planning 53: 165–216. 75 Stoker, G. (2004). Transforming Local Governance. Basingstoke, UK: Macmillan Palgrave. 76 Audit Commission (Undated). Performance Indicators 2002/03: Council Services Compendium for England. London: Audit Commission Publications. 77 Communities and Local Government (2008). National Indicators for Local Authorities and Local Authority Partnerships: Handbook of Definitions. London: CLG. 78 Morone, P., Lopolito, A., Anguilano, D. et al. (2016). Unpacking landscape pressures on socio-technical regimes: insights on the urban waste management system. Environmental Innovation and Societal Transitions 20: 62–74. 79 Seeely, A. (2009). Landfill Tax: Introduction and Early History. London: Parliamentary Briefing Paper, House of Commons Library. 80 Local Government Association (2013). Wealth from Waste: The LGA Local Waste Review. London: LGA. 81 Deutz, P. (2009). Producer responsibility in a sustainable development context: ecological modernisation or industrial ecology? Geographical Journal 175 (4): 274–285. 82 Davoudi, S. (2009). Scalar tensions in the governance of waste: the resilience of state spatial Keynesianism. Journal of Environmental Planning and Management 52 (2): 137–156. 83 Lyons, D. and Deutz, P. (2010). Regional sustainable development: making development work in politically contingent space. Sustainable Development 18: 183–186. 84 Williams, I.D. and Cole, C. (2013). The impact of alternate weekly collections on waste arisings. Science of the Total Environment 445–446: 29–40. 85 Department for Environment, Food and Rural Affairs (2017). Digest of Waste and Resource Statistics. London: Defra.

83

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production Huihui Chen, Gang Luo*, and Shicheng Zhang* *Shanghai Key Laboratory of Atmospheric Particle Pollution and Prevention (LAP3 ), Department of Environmental Science and Engineering, Fudan University, Shanghai, China

4.1 Introduction Biomass generally refers to nonfossil organic matter, e.g. animals, plants, microorganisms, and organics from their excretion and metabolisms, including agricultural and forestry residues, organic wastes, and aquatic plants from industry and urban life [1–6]. The worldwide consumption of biomass energy ranks no. 4 among all energy sources, behind only three conventional energy sources, namely petrochemical, coal, and natural gas. Around 14% of world’s primary energy consumption is accounted for by biomass [7]. For instance, China annually produces as much as 740 million tons of agricultural straw, the heat of which amounts to 317 million tons of standard coal [8]. Therefore, making the most of biomass energy can significantly reduce the consumption of nonrenewable fossil fuels. Based on the Global Forest inventories and field observations, the existing total biomass estimate is about 400 Gt C. Biomass can be classified into lignocellulosic biomass (e.g. forests, wood chip, agricultural straw, and by-products of agricultural practices) and non-lignocellulosic biomass (e.g. algae, sewage sludge, food wastes). Tropical forests are the largest reservoir of biomass on land, containing 66% of the terrestrial biomass. Temperate and boreal forests are of significance. However, each source accounts for only ∼20% the quantity of the tropical biomass. Forest lignocellulosic biomass accounts for 92% of global biomass; thus the distribution of forests is considered equal to that of biomass. Lignocellulosic biomass energy is superior to fossil energy in the following aspects [9–12]: (i) lignocellulosic biomass is renewable since it derives from plants, which draw energy from sunlight. Large quantities of plants die annually, while massive new organisms grow at the same time. (ii) As a result of its low contents of sulfur and nitrogen, lignocellulosic biomass generates less SOx and NOx during combustion as fuel, which makes it an environmentally friendly energy

Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

84

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

source. During its growth and combustion, the amount of released carbon dioxide equals that consumed [9–12]; therefore, zero emission of carbon dioxide could be achieved during its life span. (iii) Biomass is universally available from sea and land (including mountains and plains). Thus, as a kind of renewable energy, biomass energy is more attractive than the others (e.g. wind, geothermal, solar, and tidal energy) [10]. Considering the high worldwide production of agricultural and forestry biomass each year, lignocellulosic biomass is an effective and significant succedaneum to fossil fuels. A joint study, often referred to as the “Billion Ton Study,” by the US Department of Energy and the US Department of Agriculture, found that the United States annually generated approximately 1.42 billion tons of non-grain biomass feedstock that were renewable and could be put into biofuel production. These feedstock are mainly from agricultural and forestry resources (3.68 billion tons) [13]. Hydrothermal liquefaction (HTL) is an environmentally friendly thermochemical processing technology. A wide range of studies on HTL have been conducted for its potential in producing bio-oil [14–16]. HTL converts heteroatoms in biomass into harmless by-products, different from combustion that causes air pollution. As a chemical reforming process, HTL uses high-temperature pressurized water (or other solvents) as both reaction medium and reactant. Afterward, lignocellulosic biomass is depolymerized into bio-oil, biochar, biogas, and water-soluble matter (Figure 4.1) [14, 18]. Prepared bio-oils are able to serve as fuels for burners, boilers, turbines, and stationary diesel engines [15, 16]. Bio-oils can be further refined into transport fuels (gasoline and diesel) and other valuable products, such as aromatics, asphalt, lubricants, and polymers [15]. This chapter aims to provide a clear description of the research status in HTL of lignocellulosic biomass, particularly focusing on bioenergy products under different HTL conditions and the decomposition pathways of main components in biomass. In addition, the limitations of HTL are discussed, and future research priorities are proposed.

Gas phase

Water phase

Lignocellulosic biomass

CO2 removal Biorefinery

Inhibiting organics

HTL processing Bio-oil

Solid residue

Catalytic upgrading

Modification

Anaerobic digestion, et. al

High quality oil

Carbon materials

Figure 4.1 Products distribution of lignocellulosic biomass using HTL processing. Source: Cao et al. [17].

4.2 Composition of Lignocellulosic Biomass and their Degradation in HTL Processes

4.2 Composition of Lignocellulosic Biomass and their Degradation in HTL Processes 4.2.1

Composition of Lignocellulosic Biomass

Agricultural and forestry biomass, specifically called lignocellulosic biomass, includes plants and plant-based materials. Plant cell walls are comprised of carbohydrate polymers (pectin, hemicellulose, and microfibrils of cellulose) and noncarbohydrate polymers, e.g. proteins and lignin (composed of phenylpropane units). Lignocellulosic biomass samples are mainly composed of hemicellulose, cellulose, lignin (Figure 4.2) [19], and macromolecules composed of hydrogen, carbon, and oxygen atoms. The average contents of cellulose (38–50%), hemicellulose (23–32%), and lignin (15–25%) were summarized the in most promising lignocellulosic biomass, such as wheat straw, barley straw, corn stover, hardwood, softwood, Miscanthus, switchgrass, and rice straw [20]. Moreover, lignocellulosic biomass also contains a small number of pectin, nitrogen compounds, and inorganic components. The ratios of lignin, hemicellulose, and celluloses in different agricultural and forestry biomass are summarized in Table 4.1.

4.2.2

Brief Review on the Development of HTL Technology

In the early 1970s, the well-known PERC (Pittsburgh Energy Research Center) technology was developed by Appell et al., pioneers in the field [41]. Due to energy shortage and environmental pollution, considerable attention has been paid to biomass conversion to biofuel through HTL [42]. Around the year 2000, more efforts were made to dehydrate low-rank coal through hydrothermal technology on account of high water content [43]. In recent years (2010–2013), wet biomass, in particular,

Hemicellulose 15–30%

Hemicellulose

Cellulose 35–60% Lignin Lignocellulosic biomass

Figure 4.2

15–30%

Main components of woody biomass and their chemical structures.

85

86

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

Table 4.1 Compositions of different typical agricultural and forestry biomass (wt% dry basis). Cellulose (wt%)

Hemicellulose (wt%)

Lignin (wt%)

Refs.

Rice husk

34.7

17.4

25.5

[21]

Wheat straw

41.2

27.7

18.5

[22]

Corn stover

38.8

23.5

20.2

[23]

Corn cobs

44

36.4

18

[24]

Rapeseed straw

33.9

18.2

15.3

[25]

Sugarcane bagasse

56.03

36.36

[26]

Raw materials

Agricultural residues

4.59

Sunflower stalks

34.1

26.2

26.8

[27]

Sweet sorghum bagasse

36.2

24.6

13.1

[28]

White poplar

42.3

20.7

21.0

[29]

Hybrid Poplar

51.30

20.16

17.62

[30]

Aspen

47.14

19.64

22.11

[31]

39.74

21.39

25.74

[32]

40.2

18.9

25.1

[33]

43.8

20.8

28.83

[34]

51.5

11.7

34.5

[35]

25.59

19.29

19.33

[36]

Newspapers

44.21

17.84

26.84

[37]

Recycled paper

60.8

14.2

8.4

[38]

Distiller’s grains

12.63

16.9

—

[39]

Brewer’s spent grain

18.5

26.5

19.1

[40]

Hardwood

Eucalyptus globulus Eucalyptus Softwood Spruce Pinus radiata Herbaceous Bermuda Paper waste

Industrial by-products

algal biomass, has become the main processing object of HTL [44, 45]. The high lipid content of algae leads to high energy content in the bio-oil product, which is close to fossil reference values. Moreover, catalysts can be applied to increase HTL conversion rate or upgrade bio-oil [46]. In the last two years, many biological processes have been used for the production and separation of biochemicals, the use of by-products, and the development of environmental energy [8, 22, 47–50].

4.2 Composition of Lignocellulosic Biomass and their Degradation in HTL Processes

4.2.3 Main Components Degradation of the Lignocellulosic Biomass During HTL 4.2.3.1 Cellulose and its Degradation in HTL Processes

Cellulose accounts for the largest proportion of lignocellulosic materials, while hemicellulose and lignin take the second place. Figure 4.3 shows the representative structures of the three main components in the lignocellulosic biomass. Cellulose, general formula being (C6 H10 O5 )n (n ≈ 10 000), is formed by the β-1,4 glycosidic linkage of D-glucopyranose units. The glucose monomers of cellulose are bonded together by hydrogen bonds. According to Sakaki, cellulose begins to degrade into hexose and oligosaccharides when temperature exceeds 230 ∘ C, and nearly decomposes completely at 295 ∘ C [51]. As is reported by Savage et al., hydrolysis of cellulose occurred in the range of 150–250 ∘ C; at nearly 150 ∘ C and above, hydrogen-bonded structures were broken down by water and glucose monomers were released [52]. The conversion of decomposed glucose into furan derivatives occurs at the peak HTL temperature of 250 ∘ C, whereas aldehyde, carbonic acid, and acetic acid are formed when the temperature further rises [53–55]. However, at the critical point of water, high-rate conversion of cellulose is achieved with glucose or oligomer as the main product, thereby preventing further decomposition. Figure 4.3 summarizes the main degradation pathways of cellulose in HTL process. Cellulose hydrothermal degradation pathway is closely tied with the final hydrothermal temperature and the heating rate [17, 56]. Extreme heating rate (e.g. 2.2 ∘ C/s) could promote cellulose degradation to produce more oil and gas and reduce solid residue formation [57]. With a low heating rate, the reaction occurs mainly under heterogeneous conditions. These compounds may decompose

Cellulose degradation to chemicals Cellulose Temperature control Dissolution

Oligosaccharides Depolymerization Fermantation Glucose

Ethanol Hydrogenation Polylol

Figure 4.3

Humins Dehydration

HMF

Hydrolysis

Degradation pathways of cellulose in HTL processing.

Levulinic acid Formic acid

87

88

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

further to water-insoluble residues at high temperatures. At higher heating rates (>2.2 ∘ C/s), reactions (hydrolysis and decomposition) can occur in a homogeneous phase. It is expected that a homogeneous phase can be achieved by applying a high heating rate, which will increase oil and gas production. 4.2.3.2 Hemicellulose and its Degradation in HTC Process

Hemicellulose, representing 15–30% of biomass, is a mixture of various sugars, especially xylose, glucose, and mannose. Hemicellulose has different compositions depending on the type of biomass. The hemicellulose of herbaceous plants has a high proportion of xylan, whereas that of woody plants is made up mostly of glucose, mannose, and chitosan. Owing to branched chains and poor structural regularity, hemicellulose shows feebler crystallinity compared to cellulose. But hemicellulose can be easily hydrolyzed by acid and base over 180 ∘ C. Mok et al. found that hydrolysis efficiencies of hemicelluloses in both woody and herbaceous plants could reach up to 100% at 230 ∘ C for two minutes [58]. Garrote et al. reported the yields of sugar and polymer products increased from 65 to 82% by the hydrolysis of hemicellulose [59]. The decomposition of sugars occurred simultaneously with the hydrolysis of hemicellulose. Sasaki et al. studied the effect of subcritical and supercritical condition on the decomposition of D-xylose in water (360–420 ∘ C, 25–40 MPa) [51]. The experimental results suggested the occurrence of reverse aldol condensation is the main reason for D-xylose decomposition, but the dehydration reaction could be prevented in supercritical water. The ethanol aldehyde, glycerin aldehyde, and hydroxyacetone as the main decomposition products were detected. Increasing reaction temperature could be beneficial for the production of small molecules from ´ dehydration and acetylation of xylose. Pinkowska reported that the degradation of xylan could produce carboxylic acids, furfurals, and aldehydes at high temperatures (235–300 ∘ C) [60]. 4.2.3.3 Lignin and its Degradation in HTC Processes

Lignin is one of the renewable carbon sources of biomass and it accounts for 30% of the organic carbon. p-hydroxybenzoic, vanillin and syringic aldehyde are the three primarily monomers of lignin (Figure 4.4). The three-dimensional aromatic biopolymer of lignin is an attractive feedstock for producing high value-added chemicals in consideration of its considerable quantities from agricultural and forest residues [61]. Furthermore, lignin as the natural source of aromatic compounds has a higher energy density than that of polysaccharide polymers [62]. The amount and exact structure of lignin in the plant cells and wood tissues are mainly dependent on the source of biomass [63]. For example, softwood contains high content of lignin (25–30%), while hardwood contains 18–25% of lignin by weight [64]. Different reaction conditions and reactors can be optimized to improve the yield of high value-added monomers, biogases, and bio-oils, which could be further converted into low-molecular weight chemicals via hydrogenation or dehydrogenation [65, 66]. Lignin is the aromatic biopolymer made up of p-hydroxyphenyl propanoid units connected by C—C and C–O–C bonds. Therefore, the cleavage of C—C or

4.2 Composition of Lignocellulosic Biomass and their Degradation in HTL Processes CHO

CHO

CHO

H3CO

H3CO OH

OH

p-hydroxybenzoic

Figure 4.4

OCH3 OH

Vanillin

Syringic aldehyde

Primary lignin monomers.

C—O—C bonds in lignin is of great significance for the production of low-molecular weight products in the HTL process. During the HTL of lignin, the cleavage of C—O and C—C bonds, alkylation, demethoxylation, and repolymerization occurred simultaneously. Therefore, the high selective production of aromatic monomers is frustrated by the side reactions. During the liquefaction of lignin, the modified or unmodified structure of lignin, which was subjected to various pretreatments, is a determining factor for the high yield of phenolics. The harsh extraction process could lead to the cleavage of ether bond and repolymerization of C—C bond, which is difficult to break in subsequent conversion of lignin. Despite the abundance and advantages of lignin, its utilization toward bio-based chemicals and fuels is still in the initial stage due to the lack of effective depolymerization technologies and difficulties in the product separation. For the depolymerization of lignin, chemical methods were conducted to provide effective, environmentally sound, and simple strategies for the recovery of aromatic monomers [67]. Rahimi et al. demonstrated the depolymerization of oxidized lignin could achieve 60 wt% yield of low-molecular mass aromatics in aqueous formic acid [61]. Another example illustrated the promotional effect of formic acid in lignin depolymerization as shown in Figure 4.5. The oxidized lignin may enable the benzylic carbonyl group to polarize the C—H bond and reduce the energy barrier OH HO

β

OH

O 4

Lignin

OH

HO O

OMe

O OMe MeO O

OX

C–O and/or C–O cleavage

Low-molecular mass aromatic compounds

OMe O Lignin OX

HO

Figure 4.5 Lignin structure and strategies for depolymerization of representative structure of β-O-4 unite.

89

90

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

for the rate-limiting elimination reaction. Both base (formate) to remove the proton and acid (formic acid) to remove the leaving group (formate) are involved in those processes. Therefore, the high selective cleavage of β-O-4 bonds in lignin structure can be achieved in this approach. Shuai et al. also demonstrated that the addition of formaldehyde (FA) in the lignin fraction can contribute to the high yield of guaiacyl and syringyl monomers in the subsequent hydrogenolysis of Klason lignin for beech and transgenic poplar [62]. For the depolymerization of hemicelluloses, cellulose, and lignin separately, monomer yields were around 76–90 mol%. More importantly, FA could inhibit the repolymerization of active intermediates in the depolymerization process. The formation of C—C bond was hindered by two pathways: forming a stable sixmembered 1,3-dioxane (acetal) structure on lignin side chains to block the highly reactive benzylic cations and forming hydroxymethyl groups to block electron-rich positions at ortho- or para-positions to methoxyl groups on the aromatic ring. As expected, the yield of phenol-derived monomers improved significantly via the protection of reactive position in the lignin structure.

4.3 Research Status in HTL of Lignocellulosic Biomass 4.3.1

Products Description

4.3.1.1 Bio-oil

Bio-oil is the dark, viscous, and high energy-content liquid with an energy content equal to 70–95% that of fossil fuel [68, 69]. HTL process enables full-component utilization of biomass components (hemicellulose, cellulose, lignin, proteins, and lipids) to produce bio-oils. The physical and chemical properties of bio-oil are mainly dependent on the types of feedstock and conversion conditions. Typically, the composition of bio-oil is very complex and the components have a wide range of molecular weights. For quality analysis, gas chromatography (GC) and gel permeation chromatography were applied to determine the exact chemicals and molecular weight. However, most of the high-molecular mass compounds could not be eluted into the column and thus further studies are needed to prove the their exact structures and composition [68, 70]. 4.3.1.2 Solid Residue

The hydrothermal conversion of biomass can be divided into HTL, hydrothermal gasification, and hydrothermal carbonation based on the composition and proportion of the aim products [71]. The ash and char as the main solid by-products from liquefaction were studied due to their special properties. There are high contents of C, H, and N in solid residues from the liquefaction of agricultural and forestry wastes. Therefore, some nutrients in the residual ash may potentially serve as soil amendments. For hydrochar, it is typically considered as a by-product from the hydrothermal process to obtain bio-oil and chemicals. But relatively less focus was put on the environmental application of solid residue (hydrochar) due to its less aromatic

4.3 Research Status in HTL of Lignocellulosic Biomass

structure, low surface area, thermal recalcitrance, and poor porosity, thereby preventing effective exploitation [72]. Currently, hydrochar is considered as an energy carrier inferior to bio-oils and biogas. Due to the stable and nontoxic nature of hydrochar, it has many advantages and is easier to handle and store for further application. Hydrothermal carbonation may be a feasible choice for the production of functional carbonaceous materials owing to its considerably higher heating value than that of the raw material [73]. Hydrochar or modified carbonaceous can be used for carbon sequestration [74], low-cost adsorbent [71], magnetic carbon materials [75], and carbon catalyst [76]. Furthermore, hydrochar has drawn much attention as a carbon material, which could increase the efficiency of fuel cells due to its cost efficiency [77]. 4.3.1.3 Other By-products

The gaseous fraction could achieve nearly 5–10% yields of the original organic compounds from biomass feedstock [66, 78]. Among gaseous products of HTL, CO2 is the dominant product, followed by other three main components CO, H2 , and CH4. In our previous work [68], we detached the gaseous products of leaves and branches from eight selected greening plants, of which the mean yields were 8.22 and 9.17%, respectively. The volume percent of CH4 , CO, H2 , and CO2 were 2, 10, 16, and 72%, respectively. Apparently, the results showed that more amount of gases were produced from the branches than the leaves, and the amount of CO2 was much higher than other components in both kinds of biomass. From the reported literature, the low amount of CO is mainly because of decarboxylation that occurred during oxygen removal in HTL, instead of decarbonylation [79], or that the readily formed CO could react easily to form CO2 and H2 by the water–gas shift reaction [80]. The aqueous phase is also an important by-product of HTL. Affected by temperature, pressure, duration time, catalysts selected, and different types of biomass, the yields of the aqueous phase products are in the range of 20–50% related to the feedstock biomass [81, 82]. During the HTL process, about 10–50% C and 50–70% N are released into the aqueous phase [81, 83, 84]. By analyzing the composition of aqueous phase in various lignocellulosic feedstock, the main compounds are characterized as organic acids (with acetic acid and glycolic acid having the highest concentrations) and alcohols (mainly ethanol and methanol). Inorganic metallic substances were also detected, including sodium, silicon, and sulfur [85]. Moreover, numerous ketones like acetone and cyclopenta have been observed. Table 4.2 shows the representative organic compounds identified in HTL aqueous phases of different kinds of biomass.

4.3.2

Operating Parameters for Bio-oil Production by HTL

As the main product through HTL of agricultural and forestry wastes, the bio-oil part deserves thorough research to determine the optimal operating parameters that contribute to maximal bio-oil productivity. Therefore, the operating parameters for bio-oil production from HTL of lignocellulosic biomass are reviewed and discussed in this section.

91

92

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

Table 4.2

Typical organic compounds identified in aqueous phases of HTL.

Lignocellulosic biomass

Major compounds into HTC-AP

Refs.

Woody biomass (e.g. aspen, pine, sawdust, poplar)

Phenols, furfurals, glycolic, acetic acid, alcohols, and cyclopentenone

[30, 86–88]

Switchgrass

VFAs, phenols, furfurals, sugars, alcohols, and cyclopentenone

[89]

Crop straw and husk

VFAs, capronic acid, lactic acid, furfural, sugars, alcohols, and cyclopentenone

[23–25, 90–93]

Orange pomace

Acetic acid, HMF, furfurals, ethanol, acetone, butanone, and alkyl derivatives

[94]

Sugar cane bagasse

VFAs, phenols, furfural, sugars, alcohols, and cyclopentenone

[95]

Newspapers

VFAs, furfural, sugars, alcohols, and phenols cyclopentenone

[96]

Spent grain

Cyclopentenones, carboxylic acids, pyrazines and ketones

[40]

4.3.2.1 Bio-oil

The native type of biomass feedstock is a significant parameter affecting the HTL process [8, 97, 98]. For different biomass components, their thermal decomposition intervals, as well as the difficulty of liquefaction, differ greatly. Even if the same operating conditions are used, the HTL of biomass will show differences in product yield and composition. Lignocellulosic biomass owns outstanding potential to become economical and renewable feedstock for various purposes. For instance, Feng et al. studied the liquefaction of three kinds of feedstock, namely, white spruce bark, white pine bark, and birch bark, under identical technological conditions [99]. They found that the bio-oil production rates were 58, 36, and 67%, respectively. Karagoz et al. studied the HTL behavior of rice husk, sawdust, cellulose, and lignin at 280 ∘ C for 15 minutes, and found that the conversion efficiency of cellulose was higher than that of lignin [100]. The possible reasons related to this phenomenon are listed as follows: among all the major components, lignin molecular chain is complex, which is the most difficult component to break in nature; and it shows a tendency to be cooked during the process of liquefaction. In addition, the physical structure of hemicellulose and cellulose is much simpler with relatively weak hydrogen bonding and stacking interactions, which makes them show poor thermal resistance and thus higher degradability in HTL. Theoretically, the higher content the hemicellulose and cellulose take in biomass, the larger is the bio-oil yield through HTL. But, there are studies that reported opposite results for the influence of lignin content on liquefaction reaction. Chan et al. found that the content of lignin in pomace, pulp fiber, and palm kernel shell was 18.6, 30.6, and 33.5%, respectively, but the bio-oil yields under supercritical conditions

4.3 Research Status in HTL of Lignocellulosic Biomass

were, 37.39, 34.32, and 38.53%, respectively [97]. Our previous studies on bio-oil production from different green landscaping wastes through HTL also showed that the yield of hydrothermal bio-oil and biochar varies according to the different contents of lignin, hemicellulose, and cellulose. The synergistic and adverse effect of the three components as well as the difference in feedstock would influence the output of bio-oil and biochar [68]. 4.3.2.2 Temperature

Reaction temperature acts as the pivotal factor, playing a decisive role in, determining the output and quality of bio-oil in HTL of biomass. At temperatures below the critical point of water (374 ∘ C), as reaction temperature increases, the chemical bond of each component in biomass gradually breaks down and decomposes into small molecules. When the concentration of free radicals rises, the possibility of polycondensation between small molecules increases, thus improving the yield of bio-oil. At temperatures close to or slightly higher than the critical point of water, with the increase of reaction temperature, the macromolecules of bio-oil polymerize into coke or decompose into gas products, resulting in the increase of solid residue and gas production, and thereby reducing the production of bio-oil [68, 78]. Zhu et al. investigated the HTL of barley straw at different temperatures, and found that the maximum bio-oil yield reached up to 34.9% at 280–300 ∘ C [93]. When the temperature rises continually to 400 ∘ C, the bio-oil yield decreased 15.3%, whereas the solid residue yield increased 5.1%. Low temperature is likely to play a positive role in the degradation of biomass into bio-oil, while high temperature will reduce the yield of bio-oil as a result of molecular polymerization or depolymerization. Our group also observed a similar trend in which higher temperatures had a negative impact on bio-oil production from rice straw through HTL [68]. Proper heating can improve the production of bio-oil, and meanwhile promote bio-oil quality and reduce energy loss, so as to get the optimum liquefaction effect [101]. Yang et al. studied the effect of reaction temperatures (five temperatures in the range of 200–300 ∘ C at 25 ∘ C intervals) on the yield and properties of the resulting crude bio-oil [102]. They found the highest yield of the crude bio-oil (47.3% mass fraction) was achieved at 275 ∘ C, reaction time of 10 minutes, and a water/feedstock mass ratio of 20 : 1, with an initial pressure of 2.0 MPa. The calorific value of product oil is 31.0 MJ/kg, much higher than that of the raw material (20.2 MJ/kg). 4.3.2.3 Heating Rate

The heating rate is also one of the factors that are able to affect the biomass HTL; but research on the influence of heating rate is still limited. Akhtar and Amin reported that heating rate had no significant effect on the distribution of pyrolysis products [103]. The reason may be the dissolution and dispersion stability of fragment molecules in high-pressure water within the HTL process. However, Zhang et al. found that when the heating rate significantly increased from 5 to 140 ∘ C/min in the HTL of the perennial herb, the yield of bio-oil increased 13% [104]. They reported the following correlation formula: YB = (0.0042 ln R 0.5514) × 100%

93

94

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

where YB and R are the bio-oil yield and the rate of temperature rise (∘ C/min), respectively. Therefore, spontaneous heating (i.e. very fast heating rate) can increase the yield of bio-oil, while a lower rate stimulates the polymerization and increases the output of coke. The higher rate is able to trigger the decomposition reactions instantly and promote the bond cleavage and macromolecule breakdown, thus producing a large amount of energy-rich gaseous products. Therefore, in-depth studies of the effects of heating rates is urgently required in order to deeply understand the pathways of HTL reactions in the future. 4.3.2.4 Residence Time

The reaction time of HTL at the target temperature is called residence time, which excludes heating and cooling time. The level of bio-oil production is directly affected by the length of residence. Too short reaction times will lead to incomplete degradation and polymerization, while too long reaction times will lead to polymerization among intermediate products, resulting in low yield of bio-oil. Wang et al. studied the HTL of Litsea cubeba seeds after 30–120 minutes.With a time increase from 30 to 60 minutes, the recovery rate of bio-oil increased from 53.5 to 56.9% [105]. During the duration of residence, the adverse effects of polymerization and cracking reactions on the recovery of bio-oil were further increased. However, time does not affect solid residue output. It was 15.3% at 30 minutes and 13.2% at 60 minutes, decreasing a little (i.e. 25.6% at 120 minutes). Eboibi et al. took Dunaliella salina as the feedstock and found a positive effect of extended reaction time on bio-oil production at 310 ∘ C [106]. When the temperature reached 350 ∘ C, the production of bio-oil decreased with the increase of residence time, but the yields of bio-oil and natural gas increased. The results showed that the recovery rate of bio-oil was different under different reaction temperatures. Therefore, residence time and temperature have a great influence on the compound reaction. Under the same conditions, Yu et al. achieved a maximum yield of 39.4% at 280 ∘ C and 120 minutes, while Garcia et al. achieved a maximum yield of 49.4% at 5 minutes and 375 ∘ C [79, 107]. The results show that at higher reaction temperature, the polymerization reaction between liquefied products can be reduced by shortening the residence time. On the contrary, the conversion rate of feedstock and bio-oil can be improved by appropriately increasing the residence time at lower reaction temperature. In addition, at the same temperature, different residence time will also lead to different products. Therefore, different reaction conditions should be considered comprehensively to determine the optimal residence time and maximum temperature. 4.3.2.5 Pressure

Reaction pressure is a significant factor that influences the performance of high-temperature superconductivity. Under subcritical and supercritical conditions, high pressure can make water exist in a single phase, which will prevent the phase transition of water and reduce energy requirement, thus reducing energy consumption. Similarly, high pressures increase water density, increase water permeability, and increase biomass decomposition and conversion. In supercritical

4.3 Research Status in HTL of Lignocellulosic Biomass

state, the effect of pressure on water or solvent properties will decrease rapidly, and the HTL reaction will also be weakened [104]. Ngamprasertsith et al. investigated supercritical liquefaction of waste biomass. The researchers found that the yield of the bio-oil would remain constant when the biomass was under pressure of 7–12 MPa [108]. Kabyemela et al. studied supercritical liquefaction of cellobiose and the result showed that under pressure of 30–40 MPa, the hydrolysis rate of biomass remained basically unchanged, while pyrolysis rate slowed down [109]. It is speculated that the local solvent density increases due to high pressure, thus forming a cage effect on the C—C bond, which inhibits the cleavage of the C—C bond and leads to the stagnation of biomass pyrolysis. Moreover, temperature and pressure are two closely related factors in liquefaction reaction. The increase of temperature leads to the increase of pressure, which leads to the polymerization reaction among the liquefaction products and increases the yield of coke. On the contrary, the decrease in temperature leads to a decrease in pressure, which leads to incomplete liquefaction and reduces the yield of the bio-oil. 4.3.2.6 Catalysts

The purpose of adding a catalyst is to inhibit the side reactions in the HTL reaction process, lower the reaction pressure and temperature, improve the reaction rate, reduce solid residue generation, and improve the activity and quality of the bio-oil. Homogeneous and heterogeneous catalysts are widely used in the HTL process. Homogeneous catalysts including alkali catalysts and acid catalysts have been widely used in the HTL of biomass. Phosphoric acid, acetic acid, perchloric acid, hydrochloric acid, sulfuric acid, etc. are commonly used acid catalysts. Experimental results indicated that during HTL process, the bio-oil mainly acted as a solvent, but its oxygen content is higher. Although strong acids have good catalytic properties, their corrosiveness is an obstacle to their large-scale industrial application. This has led to the wide application of basic catalysts in many researches. There are eight catalysts, including Na2 CO3 , K2 CO3 , NaOH, and KOH. Minowa et al. evaluated the effect of catalysts on the liquefaction of 27 cellulose species in subcritical water. Under the action of catalyst, the recovery rate of bio-oil increased obviously and the coke yield decreased obviously. The experimental results showed that the activity of bio-oil increased by 43% after adding Na2 CO3 catalyst compared with Ni catalyst. Jindal and Jha studied the effect of alkali on hydrothermal liquefaction of wood chips of waste furniture in a 15-minute intermittent reactor at 280 ∘ C under subcritical water reaction conditions [110]. The results showed that in the presence of 1.0 m K2 CO3 , the total recovery rate of bio-oil was the highest (34.9 wt%) and the recovery rate of solid residue was the lowest (6.8 wt%). According to the biomass reaction and yield evaluation, the order of reactivity of the base was K2 CO3 > KOH > Na2 CO3 > NaOH. Singh et al. studied the effect of alkaline catalysts on water hyacinth HTL, and the experimental results showed that the total bio-oil activity was significantly increased when mixed catalysts of K2 CO3 and KOH were used [111]. It is speculated that the addition of alkali inhibited dehydration reaction of the biomass while promoted decarboxylation and water-gas conversion reactions during HTL process. Moreover,

95

96

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

Singh et al. found that the amount of catalyst also affected the effectiveness of HTL [111]. Anastasakis and Ross reached similar conclusions by studying the effect of KOH catalyst load on the products in HTL. Proper amount of catalyst is beneficial to the HTL process; very little catalyst cannot ensure the reaction speed is fast enough, while too much catalyst will inhibit the reaction [111]. Heterogeneous catalysts mainly include two kinds: metal catalysts and supported catalysts. The metals widely studied in recent years include Pt, Pd, Ni, Mo, Ru, Co, and some common carriers such as SiO2 , Al2 O3 , and zeolites. In the reaction process of HTL, the use of multiphase catalysts is most likely to lead to poisoning, internal diffusion limitation, sintering dissolution, inactivation, and other phenomena that inhibit the reaction. All, reducing gases are introduced to prevent catalyst deactivation. Although this results in a certain improvement in the recovery rate of the bio-oil, the sensitivity of its elemental composition and calorific value to the catalyst remains [46, 112]. Besides, exogenous catalysts play a very important part in lowering nitrogen content of liquid products and improving the quality of bio-oil. However, no clear conclusion has been reached yet. 4.3.2.7 Liquid-to-Solid Ratio

Water plays an important role in the dispersion of biomass, improving the quality of bio-oil, inhibiting the polymerization of intermediate products, protecting free radicals, and providing hydrogen source. It is a crucial medium in the process of biomass liquefaction. Jindal and JHA found that the ratio of water to sawdust had little effect on biomass conversion [113]. Nevertheless, as the proportion of water to sawdust increased from 2 to 6, the yield of total bio-oil (TBO) increased from 3.7 to 12.7 wt%, and further increased to 10 wt%, and then decreased to 9.6 wt%. The presence of active hydrogen is the main reason for the increase of bio-oil production. It made liquefaction intermediates more stable and reduced their formation of solid residues. In the case of low and high ratio of water to sawdust, the reasons for low bio-oil production may also be different. At low proportion of water to sawdust, as a result of the insufficient amount of water and the uneven mixture of water and sawdust suspension, the liquefaction reaction is thus hindered; while when the ratio increased from 6 to 10, the production of bio-oil decreased, but the production of solid residue (SR) and water soluble (WS) increased, which is possibly caused by the self-condensation of bio-oil components formed solid residue. Singh et al. investigated the effect of the liquid-to-solid ratio on the liquefaction of water hyacinth [110]. With the liquid-to-solid ratio increasing from 3 : 1 to 12 : 1, the conversion rate and yield of bio-oil also increased. When the ratio of liquid to solid was 3 : 1, the solid residue was as high as 47%, and the yield of bio-oil was only 6%. When the ratio was 6 : 1, the output of bio-oil was greatly increased, and the yield of solid residue reduced to 31%. The result shows that the yield of bio-oil can be significantly improved under the appropriate liquid-to-solid ratio. Wang et al. claimed that with the increase of liquid–solid ratio, the dissolution of biomass components improved, while the secondary reaction was inhibited, thus the production of solid residue and gas was reduced [105]. Lower liquid-to-solid ratio reduced the solubility

4.4 Limitations and Prospects for Bioenergy Production from Lignocellulosic Biomass by HTL

of small molecular products and intermediates in water, and improved the yield of the solid residue [103].

4.4 Limitations and Prospects for Bioenergy Production from Lignocellulosic Biomass by HTL Compared with other liquefaction technologies, the HTL technology has special advantages. The production of bio-oil through HTL can make full use of abundant biomass resources for effective energy production. However, several problems need to be solved [114, 115].

4.4.1

Poor Quality of Crude Bio-oil

The crude bio-oils from direct HTL may need to be catalyzed with reducing gas (H2 ) to convert this substance into a mixture of liquid hydrocarbons, and make it more suitable for storage, transportation, and consumption as liquid transport fuel because it is often viscous and tar like and may contain a large amount of oxygen, nitrogen, and sulfur. However, the use of reducing gas further increases the production cost. In the presence of Pt/C catalyst, the heating value (42 MJ/kg) and acid value of supercritical water were lower than that of crude bio-oil. The O and N content of the improved oil is also low, and basically sulfur-free, which can be an ideal medium to improve the efficiency of heavy oil resources.

4.4.2

Aqueous By-products Utilization

It has been reported that during the HTL process about 20–50% of the organics are transferred into the water phase (HTL-AP) [22]. In addition, there are many variable or uncertain factors in HTL process, including the natural complexity of biomass, diversity of raw materials and reaction conditions, various reaction products, intermediate products and secondary polymeric compounds (such as cyclopentenones, saccharides, volatile fatty acids, alcohols, phenols, amino acids, ammonia, and melanin). These factors lead to a very complex composition of HTL-AP. Direct emissions from HTL-AP may also harm the environment. Some HTL-AP products (such as furans, pyridines and phenols) may have toxic and inhibitory effects on marine organisms, and have an adverse impact on the fertility of land and water organisms [17, 116]. Therefore, reducing dissolved organic carbon and nutrients in HTL-AP and/or recovering these compounds through complementary processes should be considered as an important step for future employment of HTL [117]. Using HTL-AP to produce useful products is important for achieving the overall economic sustainability of the HTC process [73, 118–125]. Previous studies have attempted to recover resources from HTL-AP before they are discharged into environment by algae cultivation, extracting valuable compounds, recycling them into HTL processes and anaerobic digestion (AD) or supercritical water gasification [8, 22, 126–130]. However, there are still many challenges in using HTL-AP

97

98

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

efficiently. For example, HTL-AP extracted from HTL of algae contains toxic compounds, and even if the dilution factor is large, the growth of algae is still greatly inhibited in the culture of algae [116]. The toxic and refractory compounds in HTL-AP have also led to inefficient biogas production from AD [22].

4.4.3

Prospects

Reducing operation costs, optimizing reaction conditions, and improving the yield and quality of bio-oil are keys to the development of high-temperature biomass liquefaction technology. Based on the aforementioned factors and the knowledge of this article, future studies should try to achieve the following: ●

●

●

●

The development of new materials requires breakthroughs. It is of great significance to develop an advanced reactor with high pressure, high temperature, corrosion resistance, and pollution resistance. The development of a high-performance and economical hydrogen donor instead of the reducing gas can reduce the reaction cost and make great contribution to the practical application of HTL. In order to shorten the reaction time and lower the temperature and pressure to reduce energy consumption, a multifunctional catalyst with excellent green engineering performance and long life should be found. The effects of and approach to denitrification by homogeneous catalysts are also significant research directions. In order to maximize the benefits, more advanced technologies and methods for bio-oil refining and upgrading need to be developed to promote the comprehensive utilization of various groups in bio-oil, especially high-economic value chemicals. To further study the dynamics and reaction pathways of complex reactions, mathematical modeling and intelligent optimization techniques can be adopted to analyze operating factors and establish an optimized process model. Based on pilot-scale experiments and guided by the mechanistic models, through simulated calculation, optimal operating conditions can be determined to achieve the highest yield and best quality of bio-oil, along with promoting large-scale industrial applications of HTL technology.

4.5 Conclusion and Future Work HTL processing for converting agricultural and forestry waste biomass into biofuel is a promising technology, but further investigation is required. This chapter summarizes and discusses the effects of liquefaction conditions on the yield of bio-oil and the decomposition pathways of the main components in lignocellulosic biomass. The important factors affecting the production of HTL bio-oil are reaction temperature, lignocellulosic biomass composition, residence time, and catalyst. The combination of HTL process with subsequent conversion technologies, for instance, catalytic conversion, activation modification, and anaerobic digestion can be adopted to lower the cost of the integrated process.

References

References 1 Chen, W., Zhang, Y., Zhang, J. et al. (2014). Hydrothermal liquefaction of mixed-culture algal biomass from wastewater treatment system into bio-crude oil. Bioresource Technology 152: 130–139. 2 Duan, P., Zhang, C., Wang, F. et al. (2016). Activated carbons for the hydrothermal upgrading of crude duckweed bio-oil. Catalysis Today 274: 73–81. 3 Gardner, D.W., Huo, J., Hoff, T.C. et al. (2015). Insights into the hydrothermal stability of ZSM-5 under relevant biomass conversion reaction conditions. Acs Catalysis 5: 4418–4422. 4 Hadhoum, L., Balistrou, M., Burnens, G. et al. (2016). Hydrothermal liquefaction of oil mill wastewater for bio-oil production in subcritical conditions. Bioresource Technology 218: 9–17. 5 Yim, S., Quitain, A., Yusup, S. et al. (2016). Metal oxide-catalyzed hydrothermal liquefaction of Malaysian oil palm biomass to bio-oil under supercritical condition. Journal of Supercritical Fluids 120: 384–394. 6 Yu, I.K.M. and Tsang, D.C.W. (2017). Conversion of biomass to hydroxymethylfurfural: a review of catalytic systems and underlying mechanisms. Bioresource Technology 238: 716–732. 7 Haarlemmer, G., Guizani, C., Anouti, S. et al. (2016). Analysis and comparison of bio-oils obtained by hydrothermal liquefaction and fast pyrolysis of beech wood. Fuel 174: 180–188. 8 Chen, H., Zhang, C., Rao, Y. et al. (2017). Methane potentials of wastewater generated from hydrothermal liquefaction of rice straw: focusing on the wastewater characteristics and microbial community compositions. Biotechnology for Biofuels 10: 140–156. 9 Mahmood, R., Parshetti, G.K., and Balasubramanian, R. (2016). Energy, exergy and techno-economic analyses of hydrothermal oxidation of food waste to produce hydro-char and bio-oil. Energy 102: 187–198. 10 Durak, H. and Aysu, T. (2015). Structural analysis of bio-oils from subcritical and supercritical hydrothermal liquefaction of Datura stramonium L. Journal of Supercritical Fluids 108: 123–135. 11 Raikova, S., Smith-Baedorf, H., Bransgrove, R. et al. (2016). Assessing hydrothermal liquefaction for the production of bio-oil and enhanced metal recovery from microalgae cultivated on acid mine drainage. Fuel Processing Technology 142: 219–227. 12 Yan, W., Duan, P., Wang, F. et al. (2016). Composition of the bio-oil from the hydrothermal liquefaction of duckweed and the influence of the extraction solvents. Fuel 185: 229–235. 13 Simmons, B., Loque, D., and Blanch, H.W. (2008). Next-generation biomass feedstocks for biofuel production. Genome Biology 9: 242–242. 14 Gan, J., Yuan, W., Johnson, L. et al. (2012). Hydrothermal conversion of big bluestem for bio-oil production: the effect of ecotype and planting location. Bioresource Technology 116: 413–420.

99

100

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

15 Ramos-Tercero, E., Bertucco, A., and Brilman, D.W.F. (2015). Process water recycle in hydrothermal liquefaction of microalgae to enhance bio-oil yield. Energy & Fuels 29: 2422–2430. 16 Shanmugam, S., Adhikari, S., and Shakya, R. (2017). Nutrient removal and energy production from aqueous phase of bio-oil generated via hydrothermal liquefaction of algae. Bioresource Technology 230: 43–48. 17 Cao, L., Zhang, C., Chen, H. et al. (2017). Hydrothermal liquefaction of agricultural and forestry wastes: state-of-the-art review and future prospects. Bioresource Technology 245: 1184–1193. 18 Wen, S., Li, S., Jia, J. et al. (2013). Highly efficient conversion of cellulose to bio-oil in hot-compressed water with ultrasonic pretreatment. Industrial & Engineering Chemistry Research 52: 586–593. 19 Katalin, B. and Ford, P. (2014). Catalytic conversion of nonfood woody biomass solids to organic liquids. Accounts of Chemical Research 45: 1503–1512. 20 Gu, T. (2012). Green Biomass Pretreatment for Biofuels Production. Netherlands: Springer. 21 King-Sern, H., Rajni, H., Farook, A. et al. (2016). Conversion of rice husks to polyhydroxyalkanoates (PHA) via a three-step process: optimized alkaline pretreatment, enzymatic hydrolysis, and biosynthesis by Burkholderia cepacia USM (JCM 15050). Journal of Chemical Technology & Biotechnology 92: 100–108. 22 Chen, H., Wan, J., Chen, K. et al. (2016). Biogas production from hydrothermal liquefaction wastewater (HTLWW): focusing on the microbial communities as revealed by high-throughput sequencing of full-length 16S rRNA genes. Water Research 106: 98–107. 23 Katsimpouras, C., Christakopoulos, P., and Topakas, E. (2016). Acetic acid-catalyzed hydrothermal pretreatment of corn stover for the production of bioethanol at high-solids content. Bioprocess & Biosystems Engineering 39: 1415–1423. 24 Le, W., Ming, Y., Fan, X. et al. (2011). An environmentally friendly and efficient method for xylitol bioconversion with high-temperature-steaming corncob hydrolysate by adapted Candida tropicalis. Process Biochemistry 46: 1619–1626. 25 Chen, B., Zhao, B., Li, M. et al. (2017). Fractionation of rapeseed straw by hydrothermal/dilute acid pretreatment combined with alkali post-treatment for improving its enzymatic hydrolysis. Bioresource Technology 225: 127–133. 26 Robl, D., Costa, P., Rabelo, S. et al. (2016). Use of ascomycete extracts in enzymatic cocktail formulations increases sugar cane bagasse hydrolysis. Bioenergy Research 9: 1–7. 27 Hesami, S., Zilouei, H., Karimi, K. et al. (2015). Enhanced biogas production from sunflower stalks using hydrothermal and organosolv pretreatment. Industrial Crops & Products 76: 449–455. 28 Rohowsky, B., Häßler, T., Gladis, A. et al. (2013). Feasibility of simultaneous saccharification and juice co-fermentation on hydrothermal pretreated sweet sorghum bagasse for ethanol production. Applied Energy 102: 211–219.

References

29 Zhang, L., You, T., Tian, Z. et al. (2016). Synergistic effect of white-rot fungi and alkaline pretreatments for improving enzymatic hydrolysis of poplar wood. Industrial Crops & Products 86: 155–162. ˇ nová, ˇ 30 Kuˇcerová, V., Výbohová, E., Ca I. et al. (2016). The effects of both insoluble lignin and the macromolecular traits of cellulose on the content of saccharides within solids during hydrothermal pretreatment of hybrid poplar wood. Industrial Crops & Products 91: 22–31. 31 Pedersen, T., Grigoras, I., Hoffmann, J. et al. (2016). Continuous hydrothermal co-liquefaction of aspen wood and glycerol with water phase recirculation. Applied Energy 162: 1034–1041. 32 Gallina, G., Cabeza, Á., Biasi, P. et al. (2016). Optimal conditions for hemicelluloses extraction from Eucalyptus globulus wood: hydrothermal treatment in a semi-continuous reactor. Fuel Processing Technology 148: 350–360. 33 Chunping, X., Arancon, R., Jalel, L. et al. (2014). Lignin depolymerisation strategies: towards valuable chemicals and fuels. Chemical Society Reviews 43: 7485–7500. 34 Marzieh, S., Keikhosro, K., and Taherzadeh, M. (2010). Pretreatment of spruce and oak by N-methylmorpholine-N-oxide (NMMO) for efficient conversion of their cellulose to ethanol. Bioresource Technology 101: 4914–4918. 35 Vaidya, A., Newman, R., Campion, S. et al. (2014). Strength of adsorption of polyethylene glycol on pretreated Pinus radiata wood and consequences for enzymatic saccharification. Biomass & Bioenergy 70: 339–346. 36 Wang, Z., Keshwani, D., Redding, A. et al. (2010). Sodium hydroxide pretreatment and enzymatic hydrolysis of coastal Bermuda grass. Bioresource Technology 101: 3583–3585. 37 Montella, S., Balan, V., Sousa, L. et al. (2016). Saccharification of newspaper waste after ammonia fiber expansion or extractive ammonia. AMB Express 6: 1–10. 38 Peng, L. and Chen, Y. (2011). Conversion of paper sludge to ethanol by separate hydrolysis and fermentation (SHF) using Saccharomyces cerevisiae. Biomass & Bioenergy 35: 1600–1606. 39 Kim, Y., Hendrickson, R., Mosier, N. et al. (2010). Effect of compositional variability of distillers’ grains on cellulosic ethanol production. Bioresource Technology 101: 5385–5393. 40 Poerschmann, J., Weiner, B., Wedwitschka, H. et al. (2014). Characterization of biocoals and dissolved organic matter phases obtained upon hydrothermal carbonization of brewer’s spent grain. Bioresource Technology 164: 162–169. 41 Appell, H., Fu, Y., Friedman, S. et al. (1972). Converting organic wastes to oil. Agricultural Engineering 53: 17–25. 42 Tekin, K. (2015). Hydrothermal conversion of Russian olive seeds into crude bio-oil using a CaO catalyst derived from waste mussel shells. Energy & Fuels 29: 4382–4392. 43 Racovalis, L., Hobday, M., and Hodges, S. (2002). Effect of processing conditions on organics in wastewater from hydrothermal dewatering of low-rank coal. Fuel 81: 1369–1378.

101

102

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

44 Chen, H., Dong, Z., Gang, L. et al. (2015). Macroalgae for biofuels production: progress and perspectives. Renewable & Sustainable Energy Reviews 47: 427–437. 45 Duan, P. and Savage, P. (2011). Hydrothermal liquefaction of a microalga with heterogeneous catalysts. Industrial & Engineering Chemistry Research 50: 52–56. 46 Duan, P. and Savage, P. (2011). Catalytic treatment of crude algal bio-oil in supercritical water: optimization studies. Energy & Environmental Science 4: 1447–1456. 47 Chen, H., Rao, Y., Cao, L. et al. (2019). Hydrothermal conversion of sewage sludge: focusing on the characterization of liquid products and their methane yields. Chemical Engineering Journal 357: 367–375. 48 Chen, K., Lyu, H., Hao, S. et al. (2015). Separation of phenolic compounds with modified adsorption resin from aqueous phase products of hydrothermal liquefaction of rice straw. Bioresource Technology 182: 160–168. 49 Leng, L., Li, J., Wen, Z. et al. (2018). Use of microalgae to recycle nutrients in aqueous phase derived from hydrothermal liquefaction process. Bioresource Technology 256: 529–542. 50 Si, B., Li, J.M., Zhu, Z. et al. (2016). Continuous production of biohythane from hydrothermal liquefied cornstalk biomass via two-stage high-rate anaerobic reactors. Biotechnology for Biofuels 9: 254. 51 Sasaki, M., Hayakawa, T., Arai, K. (2003). Measurement of the rate of retro-aldol condensation of D-xylose in subcritical and supercritical water. The Proceedings of the Seventh International Symposium on Hydrothermal Reaction. Jilin University, China, 169–176. 52 Savage, P., Levine, R., and Huelsman, C. (2010). Hydrothermal processing of biomass. RSC Energy & Environment 2010: 192–221. 53 Ando, H., Sakaki, T., Kokusho, T. et al. (2000). Decomposition behavior of plant biomass in hot-compressed water. Plos One 39: 3688–3693. 54 Minowa, T., Zhen, F., and Ogi, T. (1998). Cellulose decomposition in hot-compressed water with alkali or nickel catalyst. Journal of Supercritical Fluids 13: 253–259. 55 Yu, I., Tsang, D., Yip, A. et al. (2017). Catalytic valorization of starch-rich food waste into hydroxymethylfurfural (HMF): controlling relative kinetics for high productivity. Bioresource Technology 237: 222–230. 56 Gagic, T., Perva-Uzunalic, A., Knez, Z. et al. (2018). Hydrothermal degradation of cellulose at temperature from 200 to 300 degrees C. Industrial & Engineering Chemistry Research 57 (18): 6576–6584. 57 Fang, Z., Minowa, T., Ogi, T. et al. (2004). Liquefaction and gasification of cellulose with Na2 CO3 and Ni in subcritical. Industrial & Engineering Chemistry Research 10: 2454–2463. 58 Mok, W. and Antal, M. (1992). Uncatalyzed solvolysis of whole biomass hemicellulose by hot compressed liquid water. Indengchemres 31: 1157–1161. 59 Garrote, G., Domínguez, H., and Parajó, J. (1999). Hydrothermal processing of lignocellulosic materials. Holz als Roh- und Werkstoff 57: 191–202.

References

´ ´ 60 Pinkowska, H., Wolak, P., and Złocinska, A. (2011). Hydrothermal decomposition of xylan as a model substance for plant biomass waste – hydrothermolysis in subcritical water. Biomass & Bioenergy 35: 3902–3912. 61 Rahimi, A., Ulbrich, A., Coon, J.J. et al. (2014). Formic-acid-induced depolymerization of oxidized lignin to aromatics. Nature 515: 249–252. 62 Shuai, L., Amiri, M.T., Questell-Santiago, Y.M. et al. (2016). Formaldehyde stabilization facilitates lignin monomer production during biomass depolymerization. Science 354: 329–333. 63 Mcveigh, A., Bouxin, F., Jarvis, M. et al. (2016). Catalytic depolymerisation of isolated lignin to fine chemicals: part 2 – process optimisation. Catalysis Science & Technology 12: 4142–4150. 64 Ralph, J. (2010). Hydroxycinnamates in lignification. Phytochemistry Reviews 9 (1): 65–83. 65 Bobleter, O. (1994). Hydrothermal degradation of polymers derived from plants. Progress in Polymer Science 19: 797–841. 66 Sannigrahi, P., Pu, Y., and Ragauskas, A. (2010). Cellulosic biorefineries—unleashing lignin opportunities. Current Opinion in Environmental Sustainability 2: 383–393. 67 Xu, J., Chen, J., and Sun, R. (2015). Hydrothermal microwave valorization of eucalyptus using acidic ionic liquid as catalyst toward a green biorefinery scenario. Bioresource Technology 193: 119–127. 68 Cao, L., Luo, G., Zhang, S. et al. (2016). Bio-oil production from eight selected green landscaping wastes through hydrothermal liquefaction. RSC Advances 6: 15260–15270. 69 Jindal, M. and Jha, M. (2016). Catalytic hydrothermal liquefaction of waste furniture sawdust to bio-oil. Indian Chemical Engineer 58: 157–171. 70 Anouti, S., Haarlemmer, G., Déniel, M. et al. (2016). Analysis of physicochemical properties of bio-oil from hydrothermal liquefaction of blackcurrant Pomace. Energy & Fuels 30: 398–406. 71 Gao, Y., Chen, H., Wang, J. et al. (2011). Characterization of products from hydrothermal liquefaction and carbonation of biomass model compounds and real biomass. Journal of Fuel Chemistry & Technology 39: 893–900. 72 Zhu, X., Liu, Y., Zhou, C. et al. (2014). A novel porous carbon derived from hydrothermal carbon for efficient adsorption of tetracycline. Carbon 77: 627–636. 73 Funke, A. and Ziegler, F. (2010). Hydrothermal carbonization of biomass: a summary and discussion of chemical mechanisms for process engineering. Biofuels Bioproducts & Biorefining-Biofpr 4: 160–177. 74 Rillig, M., Wagner, M., Salem, M. et al. (2010). Material derived from hydrothermal carbonization: effects on plant growth and arbuscular mycorrhiza. Applied Soil Ecology 45: 238–242. 75 Zhu, X., Feng, Q., Liu, Y. et al. (2016). Controllable synthesis of magnetic carbon composites with high porosity and strong acid resistance from hydrochar for efficient removal of organic pollutants: an overlooked influence. Carbon 99: 338–347.

103

104

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

76 Maria-Magdalena, T. and Markus, A. (2009). Chemistry and materials options of sustainable carbon materials made by hydrothermal carbonization. Chemical Society Reviews 39: 103–116. 77 Bo, H., Kan, W., Liheng, W. et al. (2010). Engineering carbon materials from the hydrothermal carbonization process of biomass. Advanced Materials 22: 813–828. 78 Brown, T., Duan, P., and Savage, P.E. (2010). Hydrothermal liquefaction and gasification of Nannochloropsis sp. Energy & Fuels 24: 3639–3646. 79 Alba, L., Torri, C., Samorì, C. et al. (2012). Hydrothermal treatment (HTT) of microalgae: evaluation of the process as conversion method in an algae biorefinery concept. Energy & Fuels 26: 642–657. 80 Elliott, D. and Sealock, L. (1983). Aqueous catalyst systems for the water-gas shift reaction. 1. Comparative catalyst studies. Industrial & Engineering Chemistry Product Research & Development 22: 426–431. 81 Biller, P., Madsen, R., Klemmer, M. et al. (2016). Effect of hydrothermal liquefaction aqueous phase recycling on bio-crude yields and composition. Bioresource Technology 220: 190–199. 82 Xu, C. and Lad, N. (2008). Production of heavy oils with high caloric values by direct liquefaction of Woody biomass in sub/near-critical water. Energy & Fuels 22: 635–642. 83 Li, C., Yang, X., Zhang, Z. et al. (2013). Hydrothermal liquefaction of desert shrub Salix psammophila to high value-added chemicals and hydrochar with recycled processing water. Bioresources 8: 2981–2997. 84 Tommaso, G., Chen, W.-T., Li, P. et al. (2015). Chemical characterization and anaerobic biodegradability of hydrothermal liquefaction aqueous products from mixed-culture wastewater algae. Bioresource Technology 178: 139–146. 85 Panisko, E., Wietsma, T., Lemmon, T. et al. (2015). Characterization of the aqueous fractions from hydrotreatment and hydrothermal liquefaction of lignocellulosic feedstocks. Biomass & Bioenergy 74: 162–171. 86 Pedersen, T. and Rosendahl, L. (2015). Production of fuel range oxygenates by supercritical hydrothermal liquefaction of lignocellulosic model systems. Biomass and Bioenergy 83: 206–215. 87 Reza, M., Yan, W., Uddin, M. et al. (2013). Reaction kinetics of hydrothermal carbonization of loblolly pine. Bioresource Technology 139: 161–169. 88 Li, H., Wang, S., Yuan, X. et al. (2018). The effects of temperature and color value on hydrochars’ properties in hydrothermal carbonization. Bioresource technology 249: 574–581. 89 Serapiglia, M., Dien, B., Boateng, A. et al. (2017). Impact of harvest time and switchgrass cultivar on sugar release through enzymatic hydrolysis. BioEnergy Research 10: 377–387. 90 Wu, K., Gao, Y., Zhu, G. et al. (2017). Characterization of dairy manure hydrochar and aqueous phase products generated by hydrothermal carbonization at different temperatures. Journal of Analytical and Applied Pyrolysis 127: 335–342.

References

91 Chen, X., Li, H., Sun, S. et al. (2016). Effect of hydrothermal pretreatment on the structural changes of alkaline ethanol lignin from wheat straw. Scientific Reports 6: 39354. 92 Gai, C., Li, Y., Peng, N. et al. (2015). Co-liquefaction of microalgae and lignocellulosic biomass in subcritical water. Bioresource Technology 185: 240–245. 93 Zhu, Z., Rosendahl, L., Toor, S.S. et al. (2015). Hydrothermal liquefaction of barley straw to bio-crude oil: effects of reaction temperature and aqueous phase recirculation. Applied Energy 137: 183–192. 94 Erdogan, E., Atila, B., Mumme, J. et al. (2015). Characterization of products from hydrothermal carbonization of orange pomace including anaerobic digestibility of process liquor. Bioresource Technology 196: 35–42. 95 Robl, D., Santos Costa, P., Rabelo, S.C. et al. (2016). Use of ascomycete extracts in enzymatic cocktail formulations increases sugar cane bagasse hydrolysis. BioEnergy Research 9: 559–565. 96 Montella, S., Balan, V., Costa Sousa, L. et al. (2016). Saccharification of newspaper waste after ammonia fiber expansion or extractive ammonia. AMB Express 6: 18. 97 Yi, H., Yusup, S., Quitain, A. et al. (2014). Bio-oil production from oil palm biomass via subcritical and supercritical hydrothermal liquefaction. Journal of Supercritical Fluids 95: 407–412. 98 Grigoras, I., Stroe, R., Sintamarean, I. et al. (2017). Effect of biomass pretreatment on the product distribution and composition resulting from the hydrothermal liquefaction of short rotation coppice willow. Bioresource Technology 231: 116–123. 99 Feng, S., Yuan, Z., Leitch, M. et al. (2014). Hydrothermal liquefaction of barks into bio-crude – effects of species and ash content/composition. Fuel 116: 214–220. 100 Karagoz, S., Bhaskar, T., Muto, A. et al. (2005). Comparative studies of oil compositions produced from sawdust, rice husk, lignin and cellulose by hydrothermal treatment. Fuel 84: 875–884. 101 Sugano, M., Takagi, H., Hirano, K. et al. (2008). Hydrothermal liquefaction of plantation biomass with two kinds of wastewater from paper industry. Journal of Materials Science 43: 2476–2486. 102 Yang, L., Nazari, L., Yuan, Z. et al. (2016). Hydrothermal liquefaction of spent coffee grounds in water medium for bio-oil production. Biomass & Bioenergy 86: 191–198. 103 Akhtar, J. and Amin, N. (2011). A review on process conditions for optimum bio-oil yield in hydrothermal liquefaction of biomass. Renewable & Sustainable Energy Reviews 15: 1615–1624. 104 Zhang, B., Keitz, M., and Valentas, K. (2009). Thermochemical liquefaction of high-diversity grassland perennials. Journal of Analytical & Applied Pyrolysis 84: 18–24. 105 Wang, F., Chang, Z., Duan, P. et al. (2013). Hydrothermal liquefaction of Litsea cubeba seed to produce bio-oils. Bioresource Technology 149: 509–515.

105

106

4 Hydrothermal Liquefaction of Lignocellulosic Biomass for Bioenergy Production

106 Eboibi, B., Lewis, D., Ashman, P. et al. (2014). Effect of operating conditions on yield and quality of biocrude during hydrothermal liquefaction of halophytic microalga Tetraselmis sp. Bioresource Technology 170: 20–29. 107 Yu, G., Zhang, Y., Schideman, L. et al. (2011). Hydrothermal liquefaction of low lipid content microalgae into bio-crude oil. Transactions of the ASABE 54: 239–246. 108 Sangon, S., Ratanavaraha, S., Ngamprasertsith, S. et al. (2006). Coal liquefaction using supercritical toluene–tetralin mixture in a semi-continuous reactor. Fuel Processing Technology 87: 201–207. 109 Kabyemela, B., Adschiri, T., Malaluan, R. et al. (1997). Kinetics of glucose epimerization and decomposition in subcritical and supercritical water. Industrial & Engineering Chemistry Research 36: 1552–1558. 110 Jindal, M. and Jha, M. (2016). Effect of process parameters on hydrothermal liquefaction of waste furniture sawdust for bio-oil production. RSC Advances 6: 41772–41780. 111 Singh, R., Balagurumurthy, B., Prakash, A. et al. (2015). Catalytic hydrothermal liquefaction of water hyacinth. Bioresource Technology 178: 157–165. 112 Anastasakis, K. and Ross, A. (2011). Hydrothermal liquefaction of the brown macro-alga Laminaria saccharina: effect of reaction conditions on product distribution and composition. Bioresource Technology 102: 4876–4883. 113 Sudong, Y., Ryan, D., Matt, H. et al. (2010). Subcritical hydrothermal liquefaction of cattle manure to bio-oil: effects of conversion parameters on bio-oil yield and characterization of bio-oil. Bioresource Technology 101: 3657–3664. 114 Doren, L., Posmanik, R., Bicalho, F. et al. (2016). Prospects for energy recovery during hydrothermal and biological processing of waste biomass. Bioresource Technology 225: 67–74. 115 Tekin, K., Karagöz, S., and Bekta¸s, S. (2014). A review of hydrothermal biomass processing. Renewable & Sustainable Energy Reviews 40: 673–687. 116 Leng, L. and Zhou, W. (2018). Chemical compositions and wastewater properties of aqueous phase (wastewater) produced from the hydrothermal treatment of wet biomass: a review. Energy Sources Part A–Recovery Utilization and Environmental Effects 40: 2648–2659. 117 Posmanik, R., Labatut, R., Kim, A. et al. (2017). Coupling hydrothermal liquefaction and anaerobic digestion for energy valorization from model biomass feedstocks. Bioresource Technology 233: 134–143. 118 He, Y., Li, X., Xue, X. et al. (2017). Biological conversion of the aqueous wastes from hydrothermal liquefaction of algae and pine wood by Rhodococci. Bioresource Technology 224: 457–464. 119 Marcilla, A., Catala, L., Garcia-Quesada, J.C. et al. (2013). A review of thermochemical conversion of microalgae. Renewable & Sustainable Energy Reviews 27: 11–19. 120 Wirth, B., Mumme, J. and Erlach, B. (2012). Anaerobic treatment of waste water derived from hydrothermal carbonization. 20th European Biomass Conference and Exhibition, Milan (June 2012), pp. 683–692.

References

121 Hollinshead, W., Henson, W., Abernathy, M. et al. (2016). Rapid metabolic analysis of Rhodococcus opacus PD630 via parallel 13C-metabolite fingerprinting. Biotechnology and Bioengineering 113: 91–100. 122 Li, C., Li, Y., Cheng, X. et al. (2013). Immobilization of Rhodococcus rhodochrous BX2 (an acetonitrile-degrading bacterium) with biofilm-forming bacteria for wastewater treatment. Bioresource Technology 131: 390–396. 123 Zeng, J., Zheng, Y., Yu, X. et al. (2013). Lignocellulosic biomass as a carbohydrate source for lipid production by Mortierella isabellina. Bioresource Technology 128: 385–391. 124 Berge, N., Ro, K., Mao, J. et al. (2011). Hydrothermal carbonization of municipal waste streams. Environmental Science & Technology 45: 5696–5703. 125 Koon, M. (2011). Recovery of Carbon and Nutrients in Lignocellulosic Biomass during Hydrothermal Carbonization. Hamburg: Universität Hamburg, School of Integrated Climate System Science (SICSS). 126 Jena, U., Vaidyanathan, N., Chinnasamy, S. et al. (2011). Evaluation of microalgae cultivation using recovered aqueous co-product from thermochemical liquefaction of algal biomass. Bioresource Technology 102: 3380–3387. 127 Biller, P., Ross, A., Skill, S. et al. (2012). Nutrient recycling of aqueous phase for microalgae cultivation from the hydrothermal liquefaction process. Algal Research 1: 70–76. 128 Zhang, L., Champagne, P., and Xu, C. (2011). Supercritical water gasification of an aqueous by-product from biomass hydrothermal liquefaction with novel Ru modified Ni catalysts. Bioresource Technology 102: 8279–8287. 129 Zhou, Y., Schideman, L., Yu, G. et al. (2013). A synergistic combination of algal wastewater treatment and hydrothermal biofuel production maximized by nutrient and carbon recycling. Energy & Environmental Science 6: 3765–3779. 130 Du, Z., Hu, B., Shi, A. et al. (2012). Cultivation of a microalga Chlorella vulgaris using recycled aqueous phase nutrients from hydrothermal carbonization process. Bioresource Technology 126: 354–357.

107

109

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management Sudheer Salana 1 , Tuhin Banerji 1 , Aman Kumar 2 , Ekta Singh 2 , and Sunil Kumar 2 1 CSIR-National Environmental and Engineering Research Institute (CSIR-NEERI), Mumbai Zonal Centre, Research and Innovation Center, Mumbai, India 2 CSIR-National Environmental and Engineering Research Institute (CSIR-NEERI), Nagpur, India

5.1 Introduction Sanitation is an essential factor for a sound and productive life, and its provisions to cope up with the ever-increasing population is one of the world’s most urgent challenges. Presently, more than 700 million urban residents lack improved sanitation access globally, including 80 million who practice open defecation [1]. From the year 2000 to 2030, urban populations in the developing countries are expected to double and serving them with sanitation services would thereby require larger investments [2]. Meeting their needs and demands would require either the development of new systems that are safe, affordable, and sustainable over the long term or advancements in the existing technologies. This requires a better understanding of the dysfunctions in the current models of sanitation services and identifying the existing key gaps in the system that threaten our ability to expand sustainable sanitation coverage. We often miss out a special benefit and only consider sanitation to be limited to provide environmental and health-related benefits. Just like the agricultural and energy sectors, sustainable sanitation automatically offers opportunities for resource recovery and reusability. It would provide opportunities for setting up new businesses and job creation, which would make the urban systems more resource efficient. The composition of waste and its physicochemical properties determine the choice of the processes in the management system. A detailed understanding and examination of all these aspects is necessary to attain a design that is not only environmentally sound but also culturally acceptable. The waste generated in toilets chiefly comprise of feces and urine, which vary in terms of both nature of waste as well as its composition. A change in the kind of food being consumed along with the average water intake changes the nature of feces and urine. The amount and frequency of excretion are known to vary with age and gender [3–5]. Much of human feces and almost all of urine is water. The water content in feces is dependent on the amount of water intake, physical exertion, the health of Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

110

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

the person, and the time the fecal matter spends in the intestine before it is excreted. Moisture content of fresh human feces is around 75–80% [5–10]. The solids consist of proteins, undigested fats, bacterial biomass, ash, and undigested food remnants. Organic material contributes 84–93% of solids fraction [8, 9]. This includes bacterial biomass (25–54%), 2–25% of proteins, carbohydrate content of 25%, and up to 15% of undigested lipids [5]. Out of the total solids, volatile solids can be around 90–92% [5, 7]. The pH of fresh feces is around 6–7 [5]. Fecal coliforms are also a major component of feces and are very important as they are used as indicators of water quality. These are facultatively anaerobic non-sporulating bacterium, examples of which are Escherichia, Enterobacter, and Citrobacter [11]. In contrast to human feces, urine is relatively sterile though it may contain some pathogens. It is mostly water (greater than 96%) with the rest being composed of urea, inorganic salts, some organic compounds, and organic ammonium salts [5]. Urine contributes to nearly 67, 73, and 88% of excreted phosphorous, potassium, and nitrogen respectively [12]. Out of all the nitrogen found in urine, 75–90% is in the form of urea, which breaks down into ammonia in the presence of water [12]. The pathogen content in urine is very low compared to feces [12]. The rest of the nutrients are found in feces. Of the fecal nitrogen, about 17% is contained in the bacterial fraction and 10% as ammonia, which is formed from degradation of urea, peptides, and amino acids [13]. The remaining nitrogen is found in different organic compounds such as uric acid and enzymes, 50% of the nitrogen is water-soluble [13]. Todt et al. [14] reported that around 13–22% of nitrogen is particulate bound in blackwater and the rest is soluble. Although phosphorus is found in some organic compounds, it is mainly found in the form of calcium phosphates in feces [13]. Particle-bound phosphorous was mainly attached to the smallest supra-colloidal solids (100 μm) (31–45% of total COD), supra-colloidal (1–100 μm) (34–43% of the total COD), colloidal (10–3.1 μm), and soluble (1000 μm), and small-sized supra-colloidal solids (1–10 μm). Particulate organic matter of 67–76% (measured as COD) was found in blackwater [14]. Hocaoglu et al. [56] reported similar values for the particulate fraction of COD of blackwater as differentiated with the routine filtration size of 450 nm. Similar studies on the COD fractionation with respect to particle size found greater than 60% of COD to be above the size of 1.2 mm. Almost 50% of protein fraction was found to be greater than 2 mm in size and 68% of carbohydrates existed in greater than 0.45 mm-sized particles [57]. A total of 31.6% of COD of blackwater is soluble COD out of which 3.7% is inert [56]. Around 80% COD is slowly biodegradable and these values match with those of Zavala et al. [6]. The slowly biodegradable COD is further differentiated into rapidly hydrolysable (30%) and slowly hydrolysable (50%). Around 15% of COD is readily biodegradable, which closely resembles the easily biodegradable fraction of human feces. The particulate organic matter has been found to be around 68% with around 20% of the particulate matter being inert [56]. Particle size distribution and its relation with COD fractions of blackwater must be carefully studied, since longer transportation durations might lead to greater hydrolysis and hence may not represent the actual conditions at the site of generation [57]. Also, the use of macerator pumps might cause significant changes in the particle sizes, thus causing the size distribution vs. COD plot to be skewed toward the smaller sized particles [14]. This discussion shows that human excreta has sufficient energy potential to be considered for design of reactors focused on energy recovery from them. These reactors may utilize feces both in their dried and wet forms. Dry human excreta have high combustion potential due to high organic content and have been found to have a Higher Heating Value (HHV) comparable to that of wood [58]. Experiments on application of anaerobic digestion technology for treatment of domestic sewage have shown that 60–70% of influent COD can be converted to methane [54]. Methane is an important component of biogas. Human waste could produce 0.02–0.028 m3 biogas per kg of dry waste, which is roughly 50% the capacity of cattle and pig waste and 1/3 of chicken excreta [54]. Laboratory experiments conducted by Colón et al. [59] on undiluted simulated human feces showed that it was possible to achieve biogas production in the range of 24–44 normal liter (NL) per person per day. It is evident from this study that dry (waterless) sanitation systems may be designed to recover energy from human feces on similar lines. Regattieri et al. [60] reported a specific biogas production rate of 0.15 m3 /kg at an organic loading rate of 0.417 kg/m3 /day in a small laboratory-scale biodigester for human waste. Methane content of the total biogas produced was

5.4 Resource Recovery Sanitation in Urban Context

reported to be around 70%. Anaerobic digestion at lower temperatures such as 15–20 ∘ C has also been reported for human feces. Meher et al. [61] reported the production of 0.48 m3 /kg of total biogas with a methane content of 57% of the total biogas collected. Hanak et al. [62] designed a nano-membrane toilet that had energy recovery system integrated. This toilet generated enough power for the functioning of small electronic devices such as mobile phones and LED lights. However, improved energy recovery from anaerobic digestion of human waste is only possible through adequate pretreatment of the human waste and also addition of suitable additives. This is because anaerobic digestion requires a minimum C : N ratio of 20 : 1–30 : 1 [63]. The average C : N ratios for human feces, urine, and their combination are 8 : 1, 0.8 : 1, and 2.3 : 1, respectively [5]. Such low ratios could cause ammonia toxicity and inhibit microbial processes severely, thus affecting the overall gas production.

5.4.2

Johkasou Systems

On-site domestic wastewater treatment systems find an application in not only rural areas but also urban contexts. Such systems would become indispensable in major cities around the world where high-rise apartments are becoming a norm. In such cases, small modular units that have all the advantages of a conventional wastewater treatment plant are more advantageous financially and from the point of view of operation and maintenance. The units should be such that they could be installed and maintained without enforcing additional burden on the owners of the buildings. Decentralized wastewater treatment systems thus should become an integral part of buildings like any other service provided such as plumbing and electricity. In Japan, a water treatment system called Johkasou that finds application in both small-scale and large-scale situations is well known. Removal of BOD in Johkasou is accomplished by either fixed film or suspended growth processes, but is frequently a hybrid of the two [64]. Johkasou consists of a combination of anaerobic and aerobic processes in stages in a single container. The first stage consists of either filter media in an anaerobic filter tank or sedimentation tank (analogous to a septic tank), which separates solids, stores sludge, and degrades organic matter by anaerobic digestion [64].The second stage consists of contact aeration tank containing media for biofilm growth that is usually plastic with a variety of shapes being used depending on the tank purposes. Here the organic matter is further degraded and ammonia is oxidized. The aeration tank is followed by a sedimentation tank, which separates suspended solids from the supernatant and sends the supernatant to disinfection tank. The supernatant is disinfected with chlorine tablets [64]. The performance of Johkasou (Figure 5.5) is dependent on organic loading rate, temperature, and recycles rates, and is not considered sufficiently reliable and further improvements have been done to the system [64]. An improved version of Johkasou uses membrane technology and is considered to be a better treatment system than the activated sludge systems, but it needs maintenance every three months and sludge withdrawal and membrane cleaning with sodium hypochlorite

123

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

Transfer pump (circulation water/backwash water)

Figure 5.5

Treated water air lift pump

Outflow

Disinfection reservoir

8

6

Biofiltration reservoir 4

Treated water reservoir

Carrier flow reservoir 3

5

Anaerobic filter bed reservoir 2

1

Solid/liquid separation reservoir

7

Inflow

124

Schematic representation of functioning of Johkasou system.

every six months to prevent fouling of the membrane. Use of membranes also adds to the cost of the treatment unit [64]. For successful implementation of Johkasou-like systems, proper septage management systems should also be in place. Septage collection and treatment should be done at central facilities where co-digestion of the septage with other organic waste could yield favorable results.

5.4.3

Possibilities of Industrial-Scale Units

Industrial-scale units of energy recovery from human excreta and sewage sludge could be possible in the next few decades with better sewage transport and organic waste collection. The current state of research and development (R&D) toward such industrial facilities includes biofuel production, improved methane recovery and solid fuel production through carbonization. Biofuel production from organic waste is not a new phenomenon. Since many decades, it has been recognized that organic materials may be used to produce fuels that could potentially rival gas produced from fossil fuels. However, achieving similar results with human waste would require improvements in existing technologies. The United States environmental protection agency (US EPA) has been running a Renewable Fuel Standard program under which it has been encouraging sewage plants to produce biogas from waste. One such example is the Persigo wastewater treatment plant in Colorado city of Grand Junction, which claims to be the only plant in the United States that produces biofuel from sewage and uses it to power vehicles [65]. Similarly, in United Kingdom, a firm working in the wastewater sector, Northumbrian Water, has been converting sludge from sewage into biogas through improved anaerobic digestion [66]. The firm now meets 20% of its electricity bills on its own. Other firms have been following suit. For example, the much publicized “poo bus” run by GENeco in Bristol showed that human and other household waste could be used to produce enough energy to run public transport [66].

5.5 Life Cycle Assessment of Sanitation Systems

Policies regarding implementation of biofuels and bioenergy are necessary drivers in any country to induce a change in the sector of wastewater treatment. Traditional approaches must be modified to suit the need of the hour and more focus on the co-benefits of wastewater treatment is needed and this is possible only through certain incentives, subsidies, and tax measures for this industry. Other measures such as investment grants are also needed to attract attention from private sector for bioenergy production. This may be coupled with exemption from fuel taxes, which would steer the population away from fossil fuels. With targets, such as zero greenhouse gas emissions by 2050, Sweden has been on the forefront in this case [67]. It has policies in place that encourage and prioritize utilization of bioenergy for transportation, which has resulted in it making giant strides in renewable energy sector. Sweden produces more than 164 million cubic meters of biomethane [67]. It has almost completely banned landfilling of organic waste and this has forced producers of waste to look at alternatives for disposing it off. Some other noteworthy policies include [68]: 1) Energy Tax exemption of biogas in transport sector. 2) Law mandating provision of at least one renewable energy fuel by all fueling stations that cross a certain limit of annual sales. 3) Congestion charge exemption for cars that run on renewable energy. 4) Greater use of transport dependent on renewable fuels by public sector bodies: several municipalities across Sweden have adopted biofuel-powered vehicles and in some cases they alone account for a broad share of the environmental cars of their respective municipalities. Although such measures are not mandatory, unwritten policies like this are effective in setting an example for common public. Such policies are very much essential for a spurt in growth of vehicle sector that focuses on green fuel.

5.5 Life Cycle Assessment of Sanitation Systems Life Cycle Assessment (LCA) involves the evaluation of the environmental burden and impact of any process or product in its entire life. This exercise helps us in identifying the stages that have the most detrimental impact and then taking suitable measures to rectify them. It also helps us in comparing the performance of two similar products or processes and choosing the one with minimal environmental impact. LCA has been increasingly brought into use by industries to improve their performance and productivity. A similar evaluation can be applied to sanitation systems to help policy makers set long-term and short-term goals in making them more environmentally friendly. Conventional wastewater treatment systems either do not include or include marginal reuse or recovery. The linear approach of “treatment” ultimately discharges organic load and heavy metals into water bodies and soil and thus contribute

125

126

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

to increase in environmental impacts such as eutrophication and toxicity. It is evident from various studies that sanitation systems that adopt separate collection of various streams of waste have an advantage over the conventional single stream systems. This advantage becomes even more pronounced when the potential for reuse of human waste as fertilizer is taken into account [45]. LCA shows that greenhouse gas contributions of conventional systems are much higher in comparison to ecological sanitation systems. A comparison of Global Warming Potential of different types of toilets systems places the contribution of conventional latrines such as Ventilated Improved Pit (VIP) latrine and pour-flush latrine at nearly 25 times and 8 times as compared to composting latrine and biodigester latrine, respectively [55]. Ecological sanitation systems also contribute to much lesser toxic load to agricultural soils as they focus on reuse of waste as a fertilizer, thus avoiding the use of chemical application to improve the crop productivity as reported by Remy and Jekel [69]. This not only avoids influx of harmful heavy metals to crop fields but also improves the water-holding capacity of soil, resulting in much lesser water use for irrigation. Remy and Jekel [69] showed that nonconventional sanitation systems have higher energy demands than conventional systems, which lead to lower climate change effects but marginally higher ammonia (NH3 ) emissions. Similar conclusions were drawn in a LCA study conducted by Spangberg et al. [70] who showed that although utilization of human excreta and urine as fertilizer would have lesser global climate change impact compared to wastewater treatment units focusing on improved nutrient recovery, the overall emissions of ammonia in currently available options for blackwater and urine treatment would be much higher. In their assessment of ecological sanitation systems, Benetto et al. [71] showed that the terrestrial ecotoxicity is much lower in ecological sanitation systems when compared to conventional sewage treatment plants considering the toxicity potential of elements such as aluminum, zinc, and copper. They however also presented results that argued that considering the distances over which the end products of ecological sanitation would be required to be transported, ecological sanitation would have greater impact on climate change and human health. This was because of the high ammonia emissions that storage and processing of waste would cause. Both Spangberg et al. [70] and Benetto et al. [71] underlined the unsustainable nature of ecological sanitation systems with respect to eutrophication and hence, future research should consider this aspect. Anastasopoulou et al. [72] conducted the Life Cycle Assessment of three different toilet systems: the pour-flush toilet, nanomembrane toilet, and the urine diversion dehydrating toilet. The urine diversion dehydrating toilet was revealed to perform the best ecologically primarily due to the benefits associated with reuse of nutrients present in urine as feces in the form of fertilizers. Advanced toilets such as the nanomembrane toilet provided greater positive human health impacts as compared to the other two, as the safety from microbial contamination is greater.

5.6 Human Excreta and Sustainable Future

Table 5.2

Cost comparison of various toilet systems.

S. no. Toilet system

Cost (approx.) (USD)

Region

Year

References

1.

Improved septic tank with reed bed

700

India

2014

[73]

2.

Ecosan

500

India

2014

[73]

3.

Anaerobic pasteurization

2 500

Duke University, 2014 USA

[73]

4.

Plasma-driven gasification

340 000–450 000 Netherlands

2014

[73]

5.

Biofil toilet with microflush

1 000–2 000

Ghana

2014

[73]

6.

Pyrolysis toilet

60 000–100 000

Kenya

2014

[73]

7.

Zero dischargecomposting vermicomposting and flocculation

6 500

India

2014

[73]

8.

Sanergy

446.41

Kenya

2014

[74]

9.

Urban affordable, clean toilets ventilated improved pit (U-ACT VIP)

783.69

Uganda

2013

[74]

10.

Pour-flush with leach pit 280

Nepal

2015

[74]

5.6 Human Excreta and Sustainable Future 5.6.1

Economics of Resource Recovery Sanitation

Table 5.2 shows the comparison of approximate costs of various toilet systems. Only the capital costs of the treatment systems have been shown here excluding the costs of the superstructure and toilet pans. The cheapest toilets are the pour-flush toilets since they are dry toilets and have no elaborate piping systems to transport sewage. These toilets are specially designed for rural areas [73]. The improved version of these toilets is the Ecosan toilets that also have systems for recycling of nutrients present in the feces as well as special arrangements for collection of urine. On the other hand, composting toilet designs available in the United States and Europe targeting higher income groups cost comparatively more [74]. The major costs involved with treatment of waste streams are due to energy requirements for operation of machinery and transportation costs. In most treatment chains, energy requirements are high. Conventional treatment plants such as wastewater treatments based on activated sludge or membrane bioreactors require high energy inputs for aeration. Certain resource recovery-oriented treatment

127

128

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

technologies too require high energy inputs. For example, anaerobic systems require energy requirements for maintaining specified temperatures (in cold countries and during low temperature periods of the year in tropical countries), for motor requirements used in mixing. In the case of advanced treatment technologies operated at an industrial scale such as thermal hydrolysis (for improved biogas and biofuel recovery from organic wastes), energy requirements are even higher. Simpler methods for resource recovery such as urine diversion and solar dehydration require no energy inputs. When costs for the treatment of the solid and liquid waste streams are also included, the scenario would significantly change. The higher costs for the present resource recovery-oriented sanitation systems available are due to lesser demand. As demand for such systems would increase, the costs would ultimately come down and be much lesser than conventional systems. Favorable markets for fecal matter-based products are necessary to bring down the costs of such systems. Nevertheless, nutrient recovery from wastewater may prove to be a valuable source of income and could also solve the problems of water stress. For example, it is estimated that if the total wastewater generated by Class I and II cities in India is considered, a revenue of USD 0.4 million daily could be created [75]. One such option is struvite production. Its production is an attractive choice that could be adopted in low-income countries for the following reasons [18]: 1) Local population would be inclined to buy a fertilizer that is locally produced because it is cheaper than the chemical fertilizers. 2) Struvite precipitation leaves a solution that is rich in nitrogen. Hence, struvite production results in two different types of fertilizers: a solid form that is rich in phosphorus and a liquid fertilizer that is rich in nitrogen.

5.6.2

Sanitation Access and Resource Recovery

One of the targets of the Goal seven (Ensure Environmental Sustainability) of the Millennium Development Goals (MDGs) was to “Halve, by 2015, the proportion of the population without sustainable access to safe drinking water and basic sanitation” [76]. As per the 2015 report on the progress of Millennium Development Goals, 2.1 billion people have gained access to sanitation around the world. It is claimed that global population resorting to open defecation has been halved compared to the numbers of 1990 [77]. However, it has to be noted that more than half of the population in South Eastern Asia, Latin America, and Caribbean and nearly 70% of the population in Sub-Saharan Africa do not have access to improved sanitation facilities as reported by United Nations Children’s Emergency Fund (UNICEF) [1] and World Health Organization (WHO) [1]. It has also been reported by United Nations [77] that these regions are often deemed to be the poorest in the world and have the least access to improved sanitation. Sanitation access is often limited due to reasons such as technical constraints, lack of a willingness to pay for sanitation services among local population, lack of sufficient funds for supplying such services, and legal and ownership issues.

5.6 Human Excreta and Sustainable Future

Lack of willingness to pay is often cited in rural areas of developing countries where the inhabitants do not recognize the threat open defecation poses. Cultural issues also play a major role wherein establishment of such services inside the premises of one’s house is deemed an unclean practice and hence avoided. The unused toilets in rural India built with the subsidies provided by the government are a glaring example of this. Also, a lack of available technical expertise keeps people away from venturing into such projects. There is a need for adequate training to be provided for the owners and users of sanitation facilities in dealing with issues that arise during the operation and maintenance. The constraints of legal issues crop up among the urban poor more than anywhere else. This is due to the reason that most of these dwellings are on illegal properties and as such investment in water connections and sewerage systems here is difficult to rise. Hence, both development and sanitation access are interlinked. It is estimated that to meet the sanitation and clean drinking water target of the MDGs, an annual investment of USD 18 billion would be required by the developing countries. For maintaining the services to the population who already have access to these facilities, an additional USD 54 billion has to be spent [78]. It is also true that many countries cite the lack of funds as the cause for sluggish progress in reaching their targets. It is here that resource recovery-oriented sanitation would play a major role. From an environment point of view, improved sanitation access need not necessarily be always beneficial unless it involves a plan for safe treatment, reuse, or recycle of the wastewater generated. Indiscriminate construction of sanitation facilities without consideration of adequate treatment channels would result in damage to the environment. The objective should be to provide sustainable sanitation. Resource recovery-oriented sanitation provides exactly this along with generating income and jobs for the local population. Research has shown that ecological sanitation programs that involved awareness campaigns about toilet usage were more successful when they were associated with related benefits of such toilets to agriculture than those that focused solely on the hygiene and health aspects [35]. However, for it to survive, resource recovery sanitation must be a profitable venture. Five different models for businesses in this area could be proposed [79]: 1) Building toilets and recovery of energy at the toilets itself. 2) Businesses focusing on fecal sludge management (collection and transportation only). 3) Businesses focusing on 2 along with an additional step of treatment. 4) Businesses that focus on reuse of end products of the treatment of fecal sludge and wastewater emanating from the toilets. 5) Bigger business models that encompass all activities in a sanitation train. Pegged at USD 62 billion by 2021 [80], building toilets is a business opportunity that is waiting to be grabbed. There is a necessity to create a need for excreta-based products and to market them in a culturally acceptable way. A combination of sound technology with detailed market research is needed in this regard.

129

130

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

5.7 Conclusion and Recommendations There is no single solution for sanitation. The applicability of solutions depends on a number of factors. Water scarcity might lead to adoption of dry toilets. Similarly, a favorable market for excreta-based products might encourage the adoption of solid liquid separator and urine separation toilets. Space constraints discourage the choice of solid–liquid separator toilets as they need greater space for solids storage. This problem can be overcome by design of suitable disposable bags for storage of fecal solids. These bags can then be collected from each household daily and composted at a common facility. Problems of maintenance and cost can be overcome by better design and planning. Similarly, problems of cultural acceptability can be overcome through better education and awareness. Thus, in case the user has the option to choose between conventional and ecological sanitation solutions, ecological sanitation solutions are a better choice. Since different streams of wastewater have different characteristics, a single treatment method may not give adequate results, especially in case of domestic or municipal wastewater, which includes myriad components such as gray water, yellow water (stream containing chiefly urine), blackwater (stream containing chiefly excreta – yellow and brown water combine together to form blackwater), and green water (wastewater emerging from kitchens). Otterpohl [21] emphasizes the need to separate different qualities of domestic wastewater and provide appropriate treatment for reuse in the form of fertilizer from blackwater. This approach will also provide a good opportunity for reuse of treated gray water. Blackwater has a composition where most of the organic matter and particulate nutrients are in solid form [21, 57] and it has sufficient nutrients N, P, and K that must be conserved. The design of the system should be flexible and, like Johkasou [64], should be applicable for a single household or a community using appropriate solid and liquid treatment units. It is advantageous that the comfort of conventional toilets is also present. Behavioral changes are gradually achieved and hence, sanitation systems also should focus on bringing a gradual modification in designs. Changing the perception of toilets as resource recycling units and human excreta as something of value can bring about an improvement in sanitation practices in developing countries. Any system should have maintenance similar to that of conventional septic tank system. Simpler processes that are easy to handle attract greater attention and have higher chances of adoption. Emphasis should be placed on utilization of nutrients in human feces through the approach of solid–liquid separation. This approach has the greatest possibility of integration with conventional sewerage and piping systems in urban areas, although enhanced engineering designs to deal with separated solids would be needed. Vacuum-based toilets might be helpful in this regard. Vacuum systems may be attached to temporary solids collection units and transported to bigger units where further processing of the solids may take place. Research is also needed in improving the existing solid–liquid separation toilets. Greater the solid–liquid separation, lesser would be the organic load on the wastewater generated in toilets, and this, hence, would reduce the level of treatment required.

References

Biochemical studies may be conducted in future on the separated water to assess the applicability of various treatment technologies to treat the separated wastewater for the pollution load and pathogen count. Similarly, suitable systems to handle and treat the separated fecal solids may also be studied. Utilization of solids separated for energy recovery may also be explored. Favorable designs would be in line with existing systems and yet achieve the goals of nutrient conservation and safe effluent disposal. A human waste management system intended for developing countries should be based on principles of zero energy usage, zero chemical addition, minimum user maintenance, and simple design with locally available materials. It has been also seen that ecological sanitation solutions are much cheaper and have much lower environmental impacts although the acidification potential remains a cause of concern and requires further research. These solutions are also feasible in developing countries and require low energy and capital investments. Hence, regions that do not have adequate wastewater treatment facilities from the very beginning should concentrate on implementing technologies that are resource recovery oriented.

References 1 UNICEF and WHO (2015). WHO/UNICEF Joint monitoring programme for water supply and sanitation, progress on drinking water and sanitation. 2 Andersson, K., Dickin, S., and Rosemarin, A. (2016). Towards “sustainable” sanitation: challenges and opportunities in urban areas. Sustainability 8 (12): 1289. (1–14). 3 Heaton, K.W., Radvan, J., Cripps, H. et al. (1992). Defecation frequency and timing, and stool form in the general population: a prospective study. Gut 33 (6): 818–824. 4 Panigrahi, M.K., Kar, S.K., Singh, S.P., and Ghoshal, U.C. (2013). Defecation frequency and stool form in a coastal eastern Indian population. Journal of Neurogastroenterology and Motility 19 (3): 374–380. 5 Rose, C., Parker, A., Jefferson, B., and Cartmell, E. (2015). The characterization of feces and urine: a review of the literature to inform advanced treatment technology. Critical Reviews in Environmental Science and Technology 45 (17): 1827–1879. 6 Zavala, M.L., Funamizu, N., and Takakuwa, T. (2002). Characterization of feces for describing the aerobic biodegradation of feces. Journal of Environmental Systems and Engineering 720 (25): 99–105. 7 Tare, V. and Yadav, K.D. (2009). Fate of physico-chemical parameters during decomposition of human feces. Global Journal of Environmental Research 3 (1): 18–21. 8 Nwaneri, C.F., Foxon, K., Bakare, B.F. et al. (2008). Biological degradation processes within a pit latrine. WISA 2008 Conference, Sun City, South Africa, (18–22 May).

131

132

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

9 Bai, F. and Wang, X. (2011). Biodegradation of organic matter and holding of N, P during aerobic thermophilic composting of human feces. Procedia Environmental Sciences 10 (C): 2631–2637. 10 Franceys, R., Pickford, J., and Reed, R. (1992). A Guide to the Development of On-site Sanitation. Geneva: World Health Organization. http://www.who.int/ water_sanitation_health/hygiene/envsan/onsitesan.pdf (accessed 26 September, 2015). 11 Winblad, U., Simpson-Hebert, M., Calvert, P. et al. (2004). Ecological Sanitation: Revised and Enlarged Edition. Stockholm, Sweden: Stockholm Environment Institute. 12 Jonsson, H., Stintzing, A.R., Vinnerås, B., and Salomon, E. (2004). Guidelines on the Use of Urine and Faeces in Crop Production. EcoSanRes Programme Publication, Stockholm Environment Institute. 13 Vinneras, B. and Jonsson, H. (2002). Faecal separation for nutrient management – evaluation of different separation techniques. Urban Water 4 (4): 321–329. 14 Todt, D., Heistad, A., and Jenssen, P.D. (2015). Load and distribution of organic matter and nutrients in a separated household wastewater stream. Environmental Technology 36 (12): 1584–1593. 15 Katukiza, A.Y., Ronteltap, M., and Niwagaba, C.B. (2012). Sustainable sanitation technology options for urban slums. Biotechnology Advances 30 (5): 964–978. 16 Taylor, M., Clarke, W.P., and Greenfield, P.F. (2003). The treatment of domestic wastewater using small-scale vermicompost filter beds. Ecological Engineering 21 (2): 197–203. 17 Etter, B., Tilley, E., Khadka, R., and Udert, K.M. (2011). Low-cost struvite production using source-separated urine in Nepal. Water Research 45 (2): 852–862. 18 Tilley, E., Gantenbein, B., Khadka, R. et al. (2009). Social and economic feasibility of struvite recovery from urine at the community level in Nepal. International Conference on Nutrient Recovery from Wastewater Streams, Vancouver, Canada (10–13 May). 19 Lalander, C.H., Hill, G.B., and Vinnerås, B. (2013). Hygienic quality of faeces treated in urine diverting vermicomposting toilets. Waste Management 33 (11): 2204–2210. 20 Gajurel, D.R., Li, Z., and Otterpohl, R. (2003). Investigation of the effectiveness of source control sanitation concepts including pre-treatment with Rottebehaelter. Water Science and Technology 48 (1): 111–118. 21 Otterpohl, R. (2001). Design of Highly Efficient Source Control Sanitation and Practical Experiences (eds. P. Lens, G. Zeeman and G. Lettinga), 165–180. IWA Publishing. 22 Wahl, T.L. (2001). Hydraulic performance of Coanda-effect screens. Journal of Hydraulic Engineering 127 (6): 480–488. 23 Buelna, G., Dubé, R., and Turgeon, N. (2008). Pig manure treatment by organic bed biofiltration. Desalination 231 (1): 297–304.

References

24 Lens, P.N., Vochten, P.M., Speleers, L., and Verstraete, W.H. (1994). Direct treatment of domestic wastewater by percolation over peat, bark and woodchips. Water Research 28 (1): 17–26. 25 Thomas, G.V., Palaniswami, C., Prabhu, S.R. et al. (2013). Cocomposting of coconut coir pith with solid poultry manure. Current Science 104 (2): 245–250. 26 Low, W.P., Fadhil Md Din, M., Ponraj, M. et al. (2015). Application of low-cost fabricated column model for the adsorption analysis of pollutants from river water using coconut coir. Desalination and Water Treatment 53 (5): 1342–1351. 27 Amoah, P., Nartey, E.G., and Schrecongost, A. (2016). Performance evaluation of biofil toilet waste digester technologies in Ghana: the efficacy of effluent treatment options. Environmental Technology: 1–12. 28 Selvaraju, N. and Pushpavanam, S. (2009). Adsorption characteristics on sand and brick beds. Chemical Engineering Journal 147 (2–3): 130–138. 29 Hotta, S. and Funamizu, N. (2007). Biodegradability of fecal nitrogen in composting process. Bioresource Technology 98 (17): 3412–3414. 30 Mohan, P. and Kovilpillai, B. (2012). Addressing the challenges of ammonia loss from poultry droppings through indigenous carbon wastes. International Journal of Environmental Science and Development 3 (4): 400–406. 31 Yadav, K.D., Tare, V., and Ahammed, M.M. (2010). Vermicomposting of source-separated human faeces for nutrient recycling. Waste Management 30 (1): 50–56. 32 McConville, J. and Rosemarin, A. (2012). Urine diversion dry toilets and greywater system, Erdos City, Inner Mongolia Autonomous Region, China – case study of sustainable sanitation projects. Sustainable Sanitation Alliance (SuSanA). 33 Cheng, S., Li, Z., Uddin, S.M.N. et al. (2018). Toilet revolution in China. Journal of Environmental Management 216: 347–356. 34 Austin, A. (2003) Ecosan: an unsuccessful sanitation scheme at a rural school – lessons learned from the project failure. Proceedings of the 2nd International Symposium on Ecological. Lubeck, Germany (7–11 April 2003). 35 Dickin, S., Dagerskog, L., Jiménez, A. et al. (2018). Understanding sustained use of ecological sanitation in rural Burkina Faso. Science of the Total Environment 613: 140–148. 36 Musingafi, T., Musingafi, M.C., and Kaseke, K.E. (2015). Challenges for ecological sanitation systems in urban areas: the case of Victoria Ranch Residential Area, City of Masvingo. Public Policy and Administration Research 5 (6): 1–13. 37 World Bank (2010). Social Factors Impacting Use of EcoSan in Rural Indonesia. Water and Sanitation Programme: Learning Note. 38 Nawab, B., Nyborg, I.L., Esser, K.B., and Jenssen, P.D. (2006). Cultural preferences in designing ecological sanitation systems in North West Frontier Province, Pakistan. Journal of Environmental Psychology 26 (3): 236–246. 39 Tobias, R., O’Keefe, M., Künzle, R. et al. (2017). Early testing of new sanitation technology for urban slums: the case of the Blue Diversion Toilet. Science of the Total Environment 576: 264–272.

133

134

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

40 Foxon, K.M., Pillay, S., Lalbahadur, T. et al. (2004). The anaerobic baffled reactor (ABR): an appropriate technology for on-site sanitation. Water SA 30 (5): 44–50. 41 Tare, V. and Nema, A. (2010). UASB technology – expectations and reality. Indian Institute of Technology Kanpur and Foundation for Greentech Environmental Systems, Kanpur and New Delhi, India. 42 Starkl, M., Strenström, T.A., Roma, E. et al. (2013). Evaluation of sanitation and wastewater treatment technologies: case studies from India. Journal of Water, Sanitation and Hygiene for Development 3 (1): 11–21. 43 Mihelcic, J.R., Fry, L.M., and Shaw, R. (2011). Global potential of phosphorus recovery from human urine and feces. Chemosphere 84 (6): 832–839. 44 Heinonen-Tanski, H. and Wijk-Sijbesma, C. (2005). Human excreta for plant production. Bioresource Technology 96 (4): 403–411. 45 Lundin, M., Bengtsson, M., and Molander, S. (2000). Life cycle assessment of wastewater systems: influence of system boundaries and scale on calculated environmental loads. Environmental Science & Technology 34 (1): 180–186. 46 Santoro, C., Ieropoulos, I., Greenman, J. et al. (2013). Power generation and contaminant removal in single chamber microbial fuel cells (SCMFCs) treating human urine. International Journal of Hydrogen Energy 38 (26): 11543–11551. 47 Kuntke, P., Smiech, K.M., Bruning, H. et al. (2012). Ammonium recovery and energy production from urine by a microbial fuel cell. Water Research 46 (8): 2627–2636. 48 Qasim, R.S. (2010). Wastewater Treatment Plants, Planning, Design and Operation, 2e. Boca Raton, FL: CRC Press. 49 Grady Jr,, C. L., Daigger, G. T., Love, N. G., Filipe, C. D. (2011) Biological Wastewater Treatment, 3rd Ed., CRC Press, Boca Raton, FL. 50 Middle, G. (1996). Environmental requirements for the disposal of effluent from wastewater disposal systems. Desalination 106 (1): 323–329. 51 Moussavi, G., Kazembeigi, F., and Farzadkia, M. (2010). Performance of a pilot scale upflow septic tank for on-site decentralized treatment of residential wastewater. Process Safety and Environmental Protection 88 (1): 47–52. 52 Al-Shayah, M. and Mahmoud, N. (2008). Start-up of an UASB-septic tank for community on-site treatment of strong domestic sewage. Bioresource Technology 99 (16): 7758–7766. 53 Aiyuk, S., Odonkor, P., Theko, N. et al. (2010). Technical problems ensuing from UASB reactor application in domestic wastewater treatment without pre-treatment. International Journal of Environmental Science and Development 1 (5): 392–398. 54 Elmitwalli, T.A., Oahn, K.L., Zeeman, G., and Lettinga, G. (2002). Treatment of domestic sewage in a two-step anaerobic filter/anaerobic hybrid system at low temperature. Water Research 36 (9): 2225–2232. 55 Galvin, C. M. (2013) Embodied energy and carbon footprint of household latrines in rural Peru: the impact of integrating resource recovery. Master thesis, University of South Florida, Tampa, Florida. http://scholarcommons.usf.edu/etd/ 4489. (accessed June 27, 2013)

References

56 Hocaoglu, S.M., Insel, G., Ubay Cokgor, E. et al. (2010). COD fractionation and biodegradation kinetics of segregated domestic wastewater: black and grey water fractions. Journal of Chemical Technology and Biotechnology 85 (9): 1241–1249. 57 Hocaoglu, S.M. and Orhon, D. (2013). Particle size distribution analysis of chemical oxygen demand fractions with different biodegradation characteristics in black water and gray water. CLEAN – Soil, Air, Water 41 (11): 1044–1051. 58 Onabanjo, T., Patchigolla, K., Wagland, S.T. et al. (2016). Energy recovery from human faeces via gasification: a thermodynamic equilibrium modelling approach. Energy Conversion and Management 118: 364–376. 59 Colón, J., Forbis-Stokes, A.A., and Deshusses, M.A. (2015). Anaerobic digestion of undiluted simulant human excreta for sanitation and energy recovery in less-developed countries. Energy for Sustainable Development 29: 57–64. 60 Regattieri, A., Bortolini, M., Ferrari, E. et al. (2017). Biogas micro-production from human organic waste - a research proposal. Sustainability 10: 330. (1-14). 61 Meher, K.K., Murthy, M.V.S., and Gollakota, K.G. (1994). Psychrophilic anaerobic digestion of human waste. Bioresource Technology 50 (2): 103–106. 62 Hanak, D.P., Kolios, A.J., Onabanjo, T. et al. (2016). Conceptual energy and water recovery system for self-sustained nano membrane toilet. Energy Conversion and Management 126: 352–361. 63 Parkin, G.F. and Owen, W.F. (1986). Fundamentals of anaerobic digestion of wastewater sludges. Journal of Environmental Engineering 112 (5): 867–920. 64 Gaulke, L.S. (2006). On-site wastewater treatment and reuses in Japan. Water Management 159 (2): 103–109. 65 Sevcenko, M. (2016). Power to the poop: one Colorado city is using human waste to run its vehicles. https://www.theguardian.com/environment/2016/jan/ 16/colorado-grand-junction-persigo-wastewater-treatment-plant-human-wasterenewable-energy (accessed January 16, 2016). 66 Hope, K. (2016). The firms turning poo into profit. https://www.bbc.com/news/ business-37981485 (accessed November 16, 2016). 67 DNA India (2014). Policy watch: lessons we can learn from Sweden in waste management. https://www.dnaindia.com/business/report-policy-watch-lessonswe-can-learn-from-sweden-in-waste-management-2007453 (accessed August 3, 2014). 68 Lonnqvist, T. (2017) Biogas in Swedish transport – a policy-driven systemic transition. PhD thesis, KTH – Royal Institute of Technology, Stockholm, Sweden. 69 Remy, C. and Jekel, M. (2008). Sustainable wastewater management: life cycle assessment of conventional and source-separating urban sanitation systems. Water Science and Technology 58 (8): 1555–1562. 70 Spangberg, J., Tidåker, P., and Jönsson, H. (2014). Environmental impact of recycling nutrients in human excreta to agriculture compared with enhanced wastewater treatment. Science of the Total Environment 493: 209–219. 71 Benetto, E., Nguyen, D., and Lohmann, T. (2009). Life cycle assessment of ecological sanitation system for small-scale wastewater treatment. Science of the Total Environment 407 (5): 1506–1516.

135

136

5 Resource Recovery-Oriented Sanitation and Sustainable Human Excreta Management

72 Anastasopoulou, A., Kolios, A., Somorin, T. et al. (2018). Conceptual environmental impact assessment of a novel self-sustained sanitation system incorporating a quantitative microbial risk assessment approach. Science of the Total Environment 639: 657–672. 73 BMGF (2014). Reinvent the Toilet Challenge, Delhi, India – Program and Technical Guides. Seattle, WA: Bill & Melinda Gates Foundation. 74 Ulrich, L., Salian, P., Saul, C. et al. (2016). Assessing the Costs of on-Site Sanitation Facilities, Study Report. Zurich: EAWAG. 75 World Bank. Water and Sanitation Program (WSP) (2016). Recycling and Reuse of Treated Wastewater in Urban India: A Proposed Advisory and Guidance Document. Colombo, Sri Lanka: International Water Management Institute (IWMI), CGIAR Research Program on Water, Land and Ecosystems (WLE). 76 United Nations (2000). Resolution Adopted by the General Assembly, United Nations Millennium Declaration. http://www.un.org/millennium/declaration/ ares552e.htm (accessed September 18, 2000). 77 United Nations (2015). Millennium Development Goals Report. Department of Public Information, United Nations Publications. 78 Prüss-Üstün, A., Bos, R., Gore, F., and Bartram, J. (2008). Safer Water, Better Health: Costs, Benefits and Sustainability of Interventions to Protect and Promote Health. Geneva: World Health Organization. 79 Rao, K.C., Kvarnström, E., Di Mario, L., and Drechsel, P. (2016). Business Models for Fecal Sludge Management. Resource Recovery and Reuse Series No. 6. Colombo, Sri Lanka: International Water Management Institute (IWMI), CGIAR Research Program on Water, Land and Ecosystems (WLE). 80 Toilet Board (2017). Introducing the Sanitation Economy. http://www.toiletboard .org/media/30-Sanitation_Economy_Final.pdf (accessed November 2017).

137

6 Resource Recovery and Recycling from Livestock Manure: Current Statue, Challenges, and Future Prospects for Sustainable Management Tao Liu * , Hongyu Chen * , Junchao Zhao, Parimala Gnana Soundari, Xiuna Ren, Sanjeev Kumar Awasthi, Yumin Duan, Mukesh Kumar Awasthi, and Zengqiang Zhang College of Natural Resources and Environment, Northwest A&F University, Yangling, P R China

6.1 Introduction The rapid development of China’s economy has greatly improved people’s living standards, and the demand for meat and milk has also increased sharply, leading to the rapid expansion of livestock farms. Thus, livestock manure has become one of the major sources of pollution in China [1]. Strengthening the control, reduction, and recycling of manure contaminants is not only the key to ensuring a sustainable development of agriculture but also of great significance for energy conservation in China. Traditionally, livestock manure is directly applied to farmland, causing serious environment pollution [2]. Manure can be used as fertilizer because it contains rich nutrients and organic matter (OM) for the growth of plants [3]. Therefore, it is necessary to find a high-efficiency method to convert manure into organic fertilizer. In this regard, composting has been considered a high-efficiency method for manure treatment and the reduction of OM and toxic metals mobility [1] and for converting the manure into the high-quality organic fertilizer [4]. However, conventional composting technologies take a long period to mature, and unstable composting could also spread diseases, release odors, and emit large amounts of greenhouse gases (GHGs) emission (Figure 6.1), which could hinder the development of composting technologies [5, 6]. Some research studies have reported that mineral and biochar amendment composting could reduce the toxicity of heavy metals, the loss of nitrogen, and GHGs emission; promote the composting process; and improve the final products’ quality [1]. The role of additives, such as wood ash [7], lime [8], coal fly ash [9], bottom ash [10], kaoline [11], and biochar [12], in the composting process has been widely investigated. The investigation revealed that the additives amendment mixed with manure can influence the composting

* Equally contributing authors. Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

138

6 Resource Recovery and Recycling from Livestock Manure

25%

Europe

44%

Oceania

6%

3%

Africa Americas

22%

Figure 6.1

Asia

Emission of CO2 by continent in manure management, 2006–2016.

processes differently, depending on the proportion added, moisture, pH, C/N ratio, and so on [1]. The lime and phosphogypsum added as an amendment into composting could reduce the total organic carbon losses [9]. Velthof [13] reported that zeolite could enhance the immobilization of Ni, Cr, Pb, and Hg greatly, and lead to the reduction of 58% carbon. Earthworm composting is an advanced and cost-effective process that converts organic waste into high-quality organic fertilizer through the native microorganisms and earthworm working together [14]. The final product vermi-compost is then utilized as manure to improve the physical, chemical, nutritional, and biological character of soil [15, 16]. The species of earthworm used for composting include red worms (Eisenia foetida), African night crawlers (Eudrilus eugeniae), red wrigglers (Lumbricus rubellus), and India blue worm (Perionyx excavatus). However, the degradation of cellulose, hemicellulose, and lignin is difficult for earthworm. A huge amount of livestock and poultry manure are produced by China, the resources of manure are abundant which contains a great potential for comprehensive utilization. It is not only beneficial for the sustainable development of agriculture, but also an important measure to prevent the pollution of livestock and improve the quality of the surrounding environment. In China, among the existing technologies for resource utilization of livestock manure, the practices of bio-aerobic composting to produce organic fertilizer and the anaerobic fermentation to produce biogas are relatively mature, and are the main technologies for livestock manure pollution control and utilization in farms in the future. But the present technology is not perfect, and we should continue to strengthen the utilization of livestock manure resources technology research. With the continuous development of technology, the use of advanced livestock manure treatment technology to deal with manure can really help to reduce toxic metal concentration, and utilize biomass as a resource.

6.2 Present Scenario and Global Perspective of Manure Generation and Recycling

Table 6.1

Number of animals by animal type in selected countries in 2017.

Country

Cattle Head

Chickens 1000 Head

Goats Head

Pigs Head

China

83 355 177

France

19 233 244

166 000

1 223 816

367 162

12 301 293

12 281 195

160 000

Germany India

185 103 532

4 973 912 139 916 096

Horses Head

Sheep Head

5 509 787 440 639 481 161 351 017 6 935 185

140 000

448 146

27 577 568

1 579 793

783 269 133 347 926

625 453

8 800 350

63 068 632

Italy

5 949 393

148 349

992 177

367 561

8 570 807

7 215 433

Japan

3 822 000

313 823

15 959

14 722

9 346 000

14 798

93 704 600

1 971 216

2 640 000 10 510 748

73 414 900

5 250 000

214 899 796

1 425 700

9 592 079

5 501 872

41 099 460

17 976 367

17 147 467

312 554

8 132 928

99 854

17 999 257

1 653 994

United States of America Brazil Myanmar Pakistan

44 400 000

494 924

72 200 000

400 000

30 100 000

0

Russian

18 752 531

497 657

2 099 357

1 381 331

22 027 698

22 744 376

6 465 747

137 500

3 059 731

260 961

29 971 357

15 963 107

Spain

6.2 Present Scenario and Global Perspective of Manure Generation and Recycling 6.2.1

Sanitization and Hygiene in Manure Management

With the continuous improvement of people’s living standards, residents’ demand for milk and meat products are increasing, which has led to the rapid development of the global livestock and poultry breeding industry. Besides, the farming methods have also been transformed from traditional decentralized to large-scale, intensive, and specialized breeding, resulting in the generation of large quantities of manure. Table 6.1 shows details of numbers of animals and poultry breeding in different countries. Improper management of manure could lead to the atmospheric, aquatic, and solid pollution due to manure’s characteristics like high content of OM, high moisture, and nutrients. Therefore, it is urgent to find a practical way to treat livestock manure in a hygienic and harmless way. Currently, biological methods including aerobic composting and anaerobic digestion have been used widely in the field of manure management. 6.2.1.1 Aerobic Composting

Aerobic composting is the process of bio-stable bio-derived organic solid waste under controlled conditions, and the resultant end product is harmless to the environment. The composting process relies on many types of natural bacteria, actinomycetes, fungi, and other microorganisms to control the processes of

139

140

6 Resource Recovery and Recycling from Livestock Manure

biodegradation of OM to form stable humus under certain artificial conditions [17]. The composting processes and the quality of final products are also influenced by many factors such as C/N, pH, moisture content, oxygen aeration, and particles [4]. Moisture Content

Moisture content is one of the key factors affecting compost maturity. The metabolism of microorganisms requires the presence of water. When the moisture content in the heap is less than 40%, the OM is not easily degraded, and the growth of microorganisms is inhibited. When the moisture content is greater than 70%, the pore blockage in the heap is not conducive to ventilation and oxygen supply, causing anaerobic conditions. Studies have demonstrated that moisture content of aerobic microorganisms between 50 and 60% is the most suitable. C/N Ratio

Carbon and nitrogen are the main energy sources for microorganisms. The appropriate C/N ratio is conducive for better growth and reproduction of microorganisms. When the C/N ratio in the heap is too high, the nitrogen content cannot be adequately supplied, and the microorganisms will be degraded. The growth rate will cause a large loss of OM and the quality of compost will decrease. When the C/N ratio in the heap is too low, the carbon content will decrease; the microorganisms will not have enough energy supply; and the temperature of the heap will rise slowly, resulting in more nitrogen loss. Many studies have found that reasonable adjustment of the C/N ratio in the heap is conducive to improving the quality of compost and accelerating the composting process. The suitable C/N ratio of compost is between 25 : 1 and 35 : 1 [18]. Oxygen Aeration

It is confirmed that a good oxygen supply is a necessary condition for aerobic composting [19]. In recent years, the oxygen supply methods for compost use natural ventilation, turning and forced aeration. The oxygen supply can provide the necessary oxygen for the growth and reproduction of microorganisms, stabilizing the moisture and temperature of the heap, and reducing the release of malodors in the heap. The oxygen supply should not be too high or too low, and the suitable oxygen content is 14–17%. pH

pH is also one of the main factors regulating the metabolism of microorganisms. Studies have shown that most microorganisms are suitable for survival under neutral or weak alkaline conditions [20]. The organic acid produced in the early stage of composting will lower the pH. As the composting process progresses, the mineralization and ammonization of the OM of the compost material will result in the increase of pH. Therefore, the pH can reflect the activity of microorganisms in the compost. The optimum pH in the compost is generally between 7.5 and 8.5.

6.2 Present Scenario and Global Perspective of Manure Generation and Recycling

Particle Size

The particle size of the material in the manure compost can affect porosity in the heap, and thus influence the reproduction of microorganisms. With the increased size of raw materials, the specific surface area will decrease, and the contact of microorganisms was insufficient. While the smaller the particle size, the smaller the porosity, which is not conducive to the ventilation and oxygen supply of the compost. Generally, the porosity in the compost is between 35 and 50%.

6.2.2

Importance and Significance of Resource Recovery

As an important part of agriculture, livestock and poultry farming, the issue of how to deal with and utilize a large amount of manure scientifically and effectively has not only become the focus of universal concern, but also one of the important factors restricting the development of China’s ecological agriculture. 6.2.2.1 Nitrogen and Phosphorus Recovery from Livestock Manure

The pollution generated in livestock and poultry breeding is the main source of agricultural nonpoint pollution with a contribution as high as 58.21%. Among these pollutions, the total nitrogen (TN), total phosphorus (TP), and chemical oxygen demand (COD) in livestock and poultry breeding respectively account for 38, 96, and 56% of agricultural sources. In addition to abundant OM, nitrogen, phosphorus, and potassium, manure also contains metallic elements, pathogenic microorganisms, various colloids, and plant residues that are not completely digested and have odor problems. With the increase in livestock and poultry manure, random stacking has led to some environmental problems, thus restricting the development of livestock and poultry industry. The nitrogen and phosphorus in the manure of livestock and poultry have aggravated the eutrophication of water bodies, causing the algae to grow wild, and reducing the dissolved oxygen in the water body, and leading to the of fish and aquatic organism in large areas. In addition, if manure is not treated and applied directly to the farmland, it will cause the enrichment of nitrogen and phosphorus in the soil. Excess nitrogen, phosphorus, and other elements in the soil will infiltrate into the groundwater, resulting in an increase in the content of ammonium nitrogen, nitrate nitrogen, and nitrous oxide in the groundwater. Excessive levels of nitrate nitrogen in drinking water could cause blue baby syndrome or induce cancer. The nitrogen and phosphorus contents in livestock manure in China’s intensive farms are as high as 3.3–14% and 0.4–5.8%, respectively. Livestock manure is an available resource in agricultural production that is rich in nutrients required for OM and plant growth (nitrogen, phosphorus, potassium, copper, zinc, etc.). The scientific treatment and resource utilization of poultry manure and agricultural and forestry waste can not only reduce the environmental pollution caused by livestock manure, turn waste into treasure, reduce the crisis of resources and energy, but also bring considerable economic and environmental benefits. The recycling industry and ecological agriculture could develop promptly. At present, the main methods for the utilization of livestock manure resources in China are feed, energy and fertilizer.

141

142

6 Resource Recovery and Recycling from Livestock Manure

6.2.2.2 Heavy Metal Recovery from Livestock Manure

Animal husbandry is considered an important cause of current environmental problems, and livestock waste is considered a major contributor. Growing demand for meat and eggs has led to an increase in production in China’s livestock industry, which produces large quantities of poultry waste every day. Uncontrolled discharge of manure can lead to serious environmental pollution and the spread of pathogens and infectious agents. Heavy metals are used as feed additives, which aim to promote the growth of animals. Long-term application of heavy metal-rich livestock manure will inevitably lead to heavy metal pollution in the soil, increasing the handling difficulty and actual cost of livestock manure [21]. Due to the unreasonable application of livestock and poultry manure, the content of Cu and Zn in farmland soil has increased significantly, exceeding the second and third standard limits in China’s Soil Environmental Quality Standard [22]. In addition, adsorption rates of Zn, Cu, and Mn in broiler chicks were only 50–60%, 30–40%, and 5–15%, respectively, causing high levels of heavy metals in animal manure and long-term contamination of soils. Heavy metal pollution caused by excessive application of livestock manure not only inhibits the growth of crop roots, reduces the yield and quality of agricultural products, but also harms human health through the food chain. Therefore, it is essential to remove/recover heavy metals form livestock manure. In recent years, there were several practical methods to recover heavy metals. For example, the bioleaching process inoculated with sulfur-oxidizing bacteria for metals leaching is regarded as an economical and efficient way due to its lower plant capacities and higher solid concentrations, compared to chemical leaching, iron-mediated bioleaching process, and bioleaching process [23]. Additionally, a vast array of innovative biosorbents have been found and used in the removal of heavy metals, such as the clay, natural zeolites, and chitosan [24, 25]. While considering their efficiencies and the economic benefits, combining various treatment methods are the tendency of heavy metal pollution treatment in the future.

6.3 Resource Recovery Technologies and Logistics for Handling, Transport, and Distribution of Manures 6.3.1

Nutrient Recovery from Manure

The high concentrations of nutrients and OM in manure are valuable resources that are produced at low cost and in large amounts. Nutrient recovery technologies are necessary for resource recovery. However, concerns about manure management include not only the reuse technologies, but also their potential environmental pollution such as organic acid, odor emission, pathogens, weed seeds, residual antibiotics, transformation of antibiotics resistance genes, heavy metals, and so on [26]. Thus these technologies for nutrient recovery of manure should effectively minimize their environmental risks. The digestive residue of chicken manure is of relatively less percentage, thereby lead to a low utilization ratio of nutrients in

6.3 Recovery Technologies and Logistics for Handling, Transport, and Distribution of Manures

Table 6.2 Nutrients, organic matter, pH, and electrical conductivity (EC) in organic wastes (dry matter). Organic waste

Chicken manure Pig manure Cattle manure

Mean

N (g/kg)

P2 O5 (g/kg)

K2 O (g/kg)

25 ± 13.9

35.9 ± 17.7

21.7 ± 8.1

Range

7.4–73.1

8.4–104.9

4.3–49.2

Mean

21.6 ± 6.6

47.4 ± 21.1

15.4 ± 5.9

OM (%)

pH

42.1 ± 13.5 7.4 ± 0.7 9.0–69.8

EC (mS/cm)

9.4 ± 4.2

6.1–9.4

1.5–24.6

54.4 ± 12.4 7.3 ± 0.6

6.5 ± 3.2

Range

7.1–38.5

11.0–97.8

3.1–34.3

18.4–71.6

5.7–8.6

2.4–17.4

Mean

14.6 ± 4.2

16.1 ± 12.2

13.9 ± 9.6

57.4 ± 12.6 7.5 ± 0.5

5.9 ± 3.7

0.8–40.2

25.3–73.5

Range

4.7–22.5

3.5–62.2

Mean

17.2 ± 5.2

13.1 ± 7.3

Sheep manure

Range

6.7–25.5

3.8–34

Other

Mean

19 ± 28.2

20.2 ± 20.6

Range 0.8–141.1

2.7–88.7

6.3–9.1

1.5–19.3

20.6 ± 13.6 54.5 ± 13.3 7.9 ± 0.6

7.9 ± 5.1

6.7–9.1

3.0–25.4

20.6 ± 15.6 36.5 ± 20.9 7.3 ± 1.5

4.1–59

9.7 ± 16.2

2.7–71.4

19.5–72.6 5.8–74.6

5.2–12.5

1.6–80.7

Other stands for organic waste including pigeon droppings, horse manure, deer feces, rabbit manure, biogas fertilizer, mushroom residue, etc.; OM, organic matter; EC, electrical conductivity.

the feed; however, relatively high nutrient content such as nitrogen, phosphorus, potassium, and OM content is present in livestock manure. Similarly, there were also very high nutrient contents in other livestock manure. Table 6.2 lists the nutrient concentrations in China according to a survey made by Chinese researchers. There was a big difference in content of N, P, K, and OM among different kinds of manures. The TN, TP, and total nutrient content in chicken and pig manure are relatively high. Because cattle and sheep mainly feed on grasses, their nitrogen content in the feces is low and the OM content is high [3, 23, 27, 28]. The nutrient concentration in animal feces is mainly dependent on the composition of initial feedstock. Additionally, the EC value of various kinds of manure is at a higher level, which is one of the reasons why animal manure is unsuitable for direct use in soil (Figure 6.2). Resource recovery based on nutrient refinement and advanced technology system are attractive approaches. Many techniques for organic manure recycling streams are available such as ammonia stripping and electrochemical from anaerobic digestion and digestate, struvite crystallization, and additive amendment. Notably, ammonia stripping is a more common approach compared with membrane processes and ion-exchange which are physical chemical technologies in nutrient surplus condition. However, treated manure addition is necessary for nutritionally deficient areas of soil, its redistribution is related to and limited by transportation costs and loss. In view of fertilizer attributes, liquid digestate has great potential to improve the plants’ tolerance under environmental stress owing to high contents of available nutrients. Liquid digestate as fertilizer is widely applied in seed soaking, foliar spray, and drip irrigation to promote soil amendment and further facilitate plant growth. Ammonia stripping and vacuum evaporation are

143

144

6 Resource Recovery and Recycling from Livestock Manure

suitable techniques for disposal of liquid digestate. Regarding nutrient recovery, implemented technology such as mesophilic anaerobic digestion combined with struvite crystallization is an efficient choice for phosphorus recovery. Combining composting and ammonia stripping for solid and liquid waste can remarkably contribute to nitrogen recovery. Integrated technologies are essential to achieve high-value products, improve nutrient and energy recovery, and simultaneously

4% 12%

Europe

37%

Oceania

21%

Africa Americas Asia

26%

(a)

29%

Europe

42%

Oceania

1%

Africa

4% 24%

Americas Asia

(b)

Figure 6.2 (a) Amount excreted in manure (N content) by continent. (b) Manure treated (N content), all animals by continent. (c) Amount excreted in manure (N content), world. (d) Manure applied to soils (N content), all animals by continent. (e) Manure treated (N content), world. (f) Manure applied to soils (N content), manure treated by all animals, world average in 2006–2016.

6.3 Recovery Technologies and Logistics for Handling, Transport, and Distribution of Manures

Cattle, nondairy

15%

Sheep 6%

37% Goats

7%

Swine, breeding Swine, market

5%

Chickens

6%

Buffaloes 12%

12%

Cattle, dairy

(c)

Europe

27%

Oceania

43%

Africa

1% 24%

(d)

Figure 6.2

(Continued)

Americas

5%

Asia

145

146

6 Resource Recovery and Recycling from Livestock Manure

Cattle, nondairy 18%

18% 2% 5%

Sheep Swine

3%

Swine, market Chickens, broilers 16%

Chickens, layers

12%

Chickens Buffaloes

5%

Turkeys

14%

7%

Cattle, dairy

(e)

Cattle, nondairy

16%

16%

Sheep

3%

5%

Swine Swine, market 16%

14%

Swine, breeding Chickens, broilers Chickens, layers

6%

Chickens 14% 8%

Buffaloes

2% Cattle, dairy

(f)

Figure 6.2

(Continued)

6.3 Recovery Technologies and Logistics for Handling, Transport, and Distribution of Manures

facilitate the economic viability based on local conditions and degree of nutrient recovery.

6.3.2

Bioenergy Production by Anaerobic Digestion/Co-digestion

Anaerobic digestion and co-digestion are value-added technologies for manure, which could decompose OM and generate methane [29]. The main biological reactions/processes during anaerobic digestion/co-digestion, that is hydrolysis, acidification, acetic acidation, and methanation [30], are listed in Figure 6.3. Carbohydrates, proteins, and fats in manure are hydrolyzed into smaller molecular weight substances, and further converted into methane, which stores most of the energy of manures. Biogas produced from anaerobic digestion/co-digestion can be applied in heat and electricity generation. Furthermore, the residues after anaerobic digestion/ co-digestion, like digestate, are good liquid fertilizers for agriculture production. The volume and weight of organic waste like manure can be effectively reduced by anaerobic digestion/co-digestion. However, a limitation of traditional anaerobic digestion is the long time required and low digestion efficiency. Thus co-digestion is recommended for their advantages since they can accelerate the anaerobic digestion process.

6.3.3

Composting/Co-composting

Composting is widely recognized and used all over the world for nutrient recovery and hygienization of organic waste, which contributes to their bioconversion from unstable organic waste to humus-like end product [31, 32]. Compared with anaerobic digestion and co-digestion, aerobic micro-organisms consumed oxygen, generating huge quantities of CO2 and heat and synchronously producing stable humus [4]. Composting is an efficient and environment-friendly approach for

Suspended particulate organic matter Hydrolysis phase Soluble organic matter Acidification phase

Volatile fatty acids Acetic acid

H2/CO2

Acetic acid phase Methanation phase

Methanation phase CH4 + CO2

Figure 6.3

Subsequent steps in the anaerobic digestion process.

147

148

6 Resource Recovery and Recycling from Livestock Manure

Window composting

Figure 6.4

Aerated static pile

Reactor composting

Examples of three types of composting.

resource recovery and recycling from livestock manure. Additionally, due to the over use of food additives with a considerable content of heavy metals and the utilization of antibiotics in the breeding of livestock, many pollutants coupled with nutrients are detected in manures, which has increased the requirement for nutrient recovery technologies [10]. Hopefully, composting can not only fix the valuable essential elements from manure, but also achieve the hygienization of these materials [21, 33, 34]. After composting, unstable and even harmful materials like manure can be biologically transformed into stable and humus-like organic fertilizers for safe land utilization [35]. According to the complexity of technology and usage, composting methods are divided into windrow, static pile, and reactor system, and examples of these methods are given in Figure 6.4. Different composting types have distinctive advantages, and the selection of composting methods should consider many factors like site, handling capacity, budget, and so on.

6.3.4

Centralized and De-centralized Models?

Anaerobic digestion and aerobic composting are both excellent and alternative approaches for nutrient recovery from organic solid waste such as poultry manure, sewage sludge (SS), and food waste. They each have their own advantages and disadvantages. However, as reported by Righi et al. [36], anaerobic co-digestion of dewatered SS and organic fraction of municipal solid waste in small plants combined with composting post-treatment may be considered an environment-friendly and sustainable method for solid waste management. Thus, the centralized models combine the advantages of both and achieve higher benefits and efficiency. Hospido et al. [37] also illustrated that anaerobic digestion can be effective in stabilization of sewage sludge, but it proved to be weak for the total removal of pollutants. Therefore, anaerobic digestion coupled with aerobic composting can be effective and safe in that the final product, mature compost, can be used as fertilizer for agriculture production. However, there were also clear advantages in de-centralized models such as: (i shorter distances of transportation for waste and lesser time spent on transportation; (ii) lower space required for waste;

6.4 Energy Matters and Economic Feasibility

(iii) maximization of the benefits for local companies; and (iv) increased public acceptability [36].

6.4 Energy Matters and Economic Feasibility 6.4.1

Energy Production

As we all know, the consumption of energy is one of the key influencing factors for economic development, which affects the quantity and type of energy resource, as well as the conversion efficiency of potential energy. The long-term development of society depends on the demands and limitations of the energy sources since there is constant depletion of the fossil fuels [38]. The dependency on energy resources and others has brought down the industrialization and development of countries. According to EU council (2007), the bioenergy is considered to be the utmost solution of replacing fossil fuels for energy demands; achieving this needs one should need to improve the bioenergy production systems without harming the existing natural ecosystem. More specifically, the EU has already aimed that renewable sources should account for at least 20% consumption of the total energy by 2020 [39]. Bio-wastes are promising energy sources, the energy stored in them can be efficiently utilized in converting their chemical components into useful fuels. The choice of bioenergy production from bio-waste mainly depends on the choices of power engines; some of the useful types of energy that can be produced are hydrogen, CH4 (biogas), bioethanol, biodiesel (bio-oil), and biochar. From the perspectives of saving energy and nutrient recovery, sanitation steps have to be followed after the digestion process . The nitrogen that is mobilized after anaerobic digestion can be recovered through ammonia stripping [40]. Among different technologies for renewable energy production, anaerobic digestion (AD) of animal manure and slurries has been widely used all over the world, and offers many advantages by improving fertilizer qualities, reducing odors and pathogens, and producing biogas. The manure contains large amount of easily degradable OM, which could be effectively used for the production of bioethanol. The dried matter needs to be milled to appropriate sizes based on the particle requirement by the furnace or boiler where it is burnt. The Danish power plants efficiently use the straw residue from manure for energy productions; more than 700 tons is burnt daily generating net calorific value of 117.5 MW. The manures with relatively high moisture are not suitable for gasification and can be efficiently employed in anaerobic digestion for biogas production. The biogas plant subjects the manure to fermentation to produce methane and hydrogen. In addition to the production of biogas, this treatment may also be helpful in several ways like odor removal, sanitation of manure, reducing indoor air pollution of kitchen when biogas is used instead of other fuels for cooking (such as fire wood and coal), reduction of COD, reduction of greenhouse gas emission, and also as remnant slurry as fertilizer for plants [41]. A study by Hamelin et al. [42] suggested that besides production of biogas from livestock manure, the treatment efficiently reduces GHG emission.

149

150

6 Resource Recovery and Recycling from Livestock Manure

6.4.2

Mineral Reutilization

The high content of carbon, nitrogen, and phosphorus in manure makes it highly suitable for targeted mineral resource recovery; ammonia stripping, struvite crystallization, and use of mineral concentrates are few promising processes. Nutrient recovery from manure reduces the nutrient load to the environment, leading to generation of additional revenue and is also compliant with the environmental regulations. 6.4.2.1 Ammonia Stripping

Through evaporation, the water and the volatiles such as fatty acids and NH3 can be removed from the manure slurry by subjecting them to temperatures little above the boiling point. Acidification of the slurry may be helpful in enhancing the total ammonical nitrogen (TAN) from the slurry. The energy costs can be reduced by using single or series of evaporator. Foust et al. [43] suggested the use of three to six evaporators for an industrial application for efficient recovery though it requires more investment. The other technique employed in ammonium recovery is stripping (air, stream). Air stripping is a process in which the liquid portion after separation from solids is brought into contact with the air, upon which the NH3 evaporates and is carried away by the gas; instead of air, steam is used in the case of stream stripping. The stripped ammonia could be absorbed by sulfuric acid. Vacuum thermal stripping on manure digestate as a post-treatment to recover nutrients has been successfully studied by Ukwuani and Tao [44]. 6.4.2.2 Struvite Crystallization

Phosphorous in the animal slurry exists in the form of crystalline or amorphous struvite (MgNH4 PO4 ⋅6H2 O) and apatite with formula Ca5 (OH)(PO4 )3 (hydroxyapatite); their formation is due to the changes in physical and chemical properties. The struvite formation helps to remove the minerals like phosphorous and ammonia from the animal slurry but sometimes interferes with the biogas plant due to their deposition on the tube surfaces, thus obstructing the path of slurry flow [31]. The struvite formation could be beneficial to reduced the mobility of ionic molecules like dissolved Mg by their complexes with dissolved organic matter [45] and decreased the pH with lowed the ability [25]. The optimal pH that is required for the struvite formation is 9. CaCO3 (calcite) may interference in hydroxyapatite formation and hence it is not preferred for removing phosphorous from the animal manure. One of studies by Suzuki et al. [31] found that struvite could be removed along with the slurry as a precipitant. The phosphorous removal of the animal manure is also enhanced by the aeration of the slurry, addition of some nucleation agents as well as addition of Mg to the slurry; Fe addition can also improve this process. 6.4.2.3 Mineral Concentrates

The nutrient-rich mineral concentrates are obtained through ultrafiltration or evaporation for their separated from the digestate; these may be applied directly to the agricultural lands. Velthof [13] reported that the industrial plants with large-scale

6.5 Resource Recovery Sanitation in Developed and Developing Countries

production and wide utilization of mineral concentrates in Europe adopted various chemical and mechanical separations of raw slurry and liquid fractions through coagulation, floatation, filtration, and concentration through reverse osmosis. The mineral concentrates are characterized as NK fertilizers with pH 7.9, implying less risk of N losses by volatilization.

6.5 Resource Recovery Sanitation in Developed and Developing Countries Sanitization of the manure is important to prevent serious human and environmental issues; special attention is needed for farms that are close to the cities in order to prevent the contamination of soil and water bodies [46]. Ignorance and negligence of the people about the risks related to the environmental pollution often put their lives at stake; and this will exist until and unless the necessary actions are carried out in terms of mitigating it. Culture, market power and expertise, public policy measures, and civic society pressure are the four major drivers in governing and organizing societies. The government policies aim at modifying human behavior to achieve societal objectives in terms of well-being. The policies used are different from country to country; a few of the general policies widely adopted with respect to livestock manure are discussed in this section. The conventions and protocols of many of the intergovernmental organizations such as United Nations (UN) and multilateral environmental agreements (MEA) have already made great contribution in solving lots of environmental issues in the past and are still working to tackle the concerns about the environmental and sustainable development of environment regulations all over the world. Few of the regulations regarding the greenhouse gases (one main emission contributor is animal manure) are as follows: UNFCCC (United nation framework convention on climate change) protocols outlined the regulations for decreasing greenhouse gas emission; United nation economic commission for Europe (UNECE) has hosted five conventions such as Convention on Long-Range Trans-boundary Air Pollution (CLRTAP) which usually has eight protocols that identify specific measures to be taken by its unions (51 countries) to cut down emission of pollutants (SO2 , O3 , NOx, NH3, and volatile organic compounds). The specific measures for NH3 emission reduction from animal manure and fertilizers were specified in annex IX of Gothenburg protocol [47]. The intergovernmental panel on climate changes (IPCC) devised the protocol for estimating CH4 and N2 O emission from livestock manure. EU-related environmental policies are mostly established by directives, which are imposed to achieve environmental purposes by member states. The cause–effect relationship should be understood in order to deal with the policies amended by EU on manure management perspectives; where emissions reflect a simple source–receptor/effect model (source is the one emitting pollutant [manure] and receptor is the one getting affected [atmosphere, surface and ground water]). In 1958, EU Common Alerting Protocol (CAP) was established for addressing the modernization and productivity of agriculture and food security; indirectly

151

152

6 Resource Recovery and Recycling from Livestock Manure

contributing increased use of natural resources, animals as well as environmental effect by agriculture. Cross compliance becomes an element coupling existing environmental policies to agriculture income support; Single Farm Payments with two major aspects such as Statutory Management Requirements (SMRs) and Good Agricultural and Environmental Condition (GAEC), and a few of them manage manure management [48]. The EU environmental directory deals with manure management with respect to the environmental concerns. The 1991 nitrate directive deals with reduction and prevention of water pollution caused by nitrates of agricultural sources, suggesting the member states to take steps in water monitoring in respect to nitrate concentration, identifying contaminated water, designation of vulnerable zones, establishment of codes and necessary actions, and finally reviewing the reported sites for every four years of management in order to prevent eutrophication and further monitoring of the application of manure and fertilizer to the land, thus also covering the manure management in terms of generation and storage. The 2008 IPPC (Integrated Pollution Prevention and Control) directive suggests industrial pollution managements by Best Available Techniques (BAT). The implementation of these laws in the member state differs and is a comparatively slow process in many cases due to many reasons such as large differences in farming systems with regard to complexity in manure management and nutrient cycling, varying interpretations of directives, perceived high cost to farmers, reluctance of introducing mechanisms, legislative delays, different constraints and potential antagonisms between measures. The treatments for the reduction of pathogens such as Enterococcus faecalis, Salmonella, and other associated pathogens to achieve limits of five decimals (5 log 10) in animal manure were prescribed under category 2 and category 3 of EU regulations on animal by-products [48]; some thermotolerant viruses like parvoviruses must be reduced to three decimals (3 log 10) through one hour at above 70 ∘ C or related treatments. Through the chemical treatments must be accepted when there was the reduction of 3 log 10 for Ascaris spp., since this strain used as model for chemical treatment system [49]. The Danish regulations on livestock production and management of manure aimed at achieving sustainable recycling of livestock manure and preventing the emission of plant nutrients and environmental pollutants. The “Harmony regulations” with respect to EU nitrate directive were implemented during early 1990s specifying a maximum density corresponding to 140–170 kg animal manure per hectare for available land application; it also focuses on reducing leaching, odor, and NH3 emissions. In compliance with EU directives, Danish regulations [50] also regulate animal manure production and management as well as environmental protection with some of the following guidelines: direct discharge of animal manure to water/field is strictly forbidden encouraging sanitation standards, zoning of various production units, construction of production units preventing leaching, restriction in duration for applying manure, BAT for applying liquid manure preventing emission, and also solid manure application restrictions in the lands in close contact with water by preventing runoff.

6.5 Resource Recovery Sanitation in Developed and Developing Countries

The CWA (clean water act) and CAA (clean air act) of the United States mainly regulates the livestock production and manure management in addition to many additional state-specific regulations. The livestock in barns and feedlots are reported to be the point source of water pollution; NPDES (National Pollutant Discharge Elimination System) requires permission under CWA for their discharge in waters of the United States [51], with laws enforcing treatment of manure before spraying them to the fields or water units and also confining the emission of SO2 , NO2 , O3 , NH3 , CO, and Pb and other particulate matter. It also encourages the farmer to get pre-construction permits proposing air abatement technologies [52]. The National Resource Conservation Service has set standards for manure nutrient management and sets criteria for fertilizer and manure application minimizing the nutrient entry into ground, water, and atmosphere as well as improving soil qualities [53]. The EST (environmentally superior waste management technology) enacted by state of North Carolina in 2007 enforces five environmental performance standards as a basic requirement for construction or expansion of swine farms with objectives of elimination of direct discharges, ammonia emission, odor beyond boundaries of farm location, release of disease-transmitting vectors or pathogens, and also nutrient or heavy metal contamination [54]. Large animal feeding operations identified as Concentrated Animal Feeding Operations (CAFO) are required to have a comprehensive nutrient management plan (CNMP); these are complex and expensive management systems and are followed by Maryland, Delaware, and Virginia. The National Environmental Strategy 2011–2015 of Cuba prescribes the application of sustainable production and consumption toward principles of cleaner concepts. They obligate all the companies toward environmental protections for achieving the goals of sustainable development by the Law 81 [55]. In Japan, because of the incessant complaints of pollution caused by livestock manure, various laws were enforced by the government, such as regulations and laws of treatment and utilization on livestock manure (1999), law of water pollution control (1970), law of offensive odor control (1971), and the law of air pollution control aiming to encourage appropriate sanitation treatments and promote agricultural usages, setting limits on the composition of waste from livestock farms to the discharge targets like water/field, limiting concentration of 22 offensive odors caused by manure management, preventing air pollutants respectively, which are actively in practice [56].

6.5.1

Operational Guidelines for Septage Treatment and Disposal

The treatment of the manure slurry may be aerobic, anaerobic, composting or treatment with chemical agents, with the time as additional factor for evaluating pathogen content in animal excreta. The solid part of the manure is rich in phosphorous, most of the nitrogen stays in the liquid part. The feedstock is usually sanitized at the beginning before entering into the digester in the case of continuous flow-stirred tank reactor as a means of reducing the possibilities of short circuit flows. The high-temperature treatment might be used when quick sanitation of manure is needed.

153

154

6 Resource Recovery and Recycling from Livestock Manure

6.5.1.1 Storage

Storage is the main treatment option for the reduction of pathogen load; as long as the fresh manure is not added to the manure store tank, the substrate required for the growth and multiplication of the pathogens will be reduced. The inactivation of bacterial pathogens generally occurs faster when compared to the other parasites and viruses. 6.5.1.2 Pasteurization

Consequently, treatment is necessary to minimize the risk of infection. For urine, the present recommended minimum treatment is storage for six months without introducing fresh liquid [57]. Pasteurization (70 ∘ C for one hour) or an equivalent treatment is necessary for fecal matter and sludge [57]. Energy savings could be achieved by heating up less material after digestion as well as separating a lower mass flow in the centrifuges due to water losses during the sanitation. 6.5.1.3 Chemical Treatments

The common chemical used for sanitation is ammonia and urea; when urea reacts with animal wastes, it releases ammonia. The ammonia treatment can inactive the microbial population due its high pH. The ammonium ion (NH4 + -N) is harmless to the microorganisms; inactivation is by NH3 . There are many ways to propose, which can avoid destroying membrane potentials, and achieve easy solubility and transport over membranes, denaturation of proteins, cleavage of nucleic acids, rapid alkalization of cytoplasm and also microbial compensation by losing essential elements [58]. Ottoson et al. [58] reported the decimal reduction for the different microorganisms for different NH3 concentrations, temperatures, and pH in sanitizing animal manure. However, when combining these data it is possible to estimate the time required for inactivation according to the ABP treatment regulations [49] at different NH3 concentrations. 6.5.1.4 Anaerobic Treatments

The energy components of the degradation of OM in an anaerobic treatment ends with the methane, and there will not be generation of any surplus heat; thus this process requires external heating to reach the ambient temperature. The anaerobic process may occur in a wide range of conditions from psychrophillic to hyperthermophillic temperatures and are always associated with the production of biogas. Pathogenic Enterobacteriaceae are involved in the acidification process and are considered to be the initial degradation step that reaches a threshold in a psychrophilic or mesophilic reactor. Its inactivation can be achieved by three factors: heat inactivation, ammonia accumulation and retention time. Heat inactivation can be achieved via pre-pasteurization by heating the animal manure at 70 ∘ C for an hour during their fed time into the reactor or from the thermophilic process at not less than 50 ∘ C [59]. The anaerobic treatment of manure at 70 ∘ C for minimum two hours ensures the eradication of some of the pathogenic bacteria such as E. coli, Staphylococcus, and Salmonella, which obstruct the digestive system of humans. Once after the adaptation of pathogens in the initial stages of digestion, a high concentration of ammonia

6.5 Resource Recovery Sanitation in Developed and Developing Countries

in the digester can be achieved by introduction of high protein load. The sudden exposure of NH3 can be lethal to the microorganisms, many of the pathogenic population will be decreased [60]. The retention time plays an important role in deactivation of pathogens. Many of the common reactors are fully continuous anaerobic types, and the minimum retention time is very shorter, so combating pathogens is a bottleneck. Maximizing the hydraulic retention time will increase the concentration of pathogens, but also due to the increase in reactor time. The sanitation standards of Europe strongly recommend a hydraulic retention time of one hour at 70 ∘ C for sanitation; the bottleneck is the high energy demands [61]. 6.5.1.5 Composting

Composting is one of the common techniques for manure management; it helps in conversion as well as sanitation of manures. Several processes such as competition, nutrient deficiency, and temperature claim to inactivate the pathogens. But the main responsible factor is heat generation during the oxidative degradation of the OM, wherein the initial days of composting the energy are often spent on the growth of the bacterial population instead of heat release. Some pathogens are easily inactivated in mesophilic temperature such as bird influenza, but many of the pathogens need at least 50 ∘ C for inactivation; a few microbes even start their growth around this temperature range as in the case of Salmonella (47 ∘ C optimal growth), so they require longer treatment of say a minimum of one week under 55 ∘ C [62]. The challenging part is that the treatment must be performed in a way that all of the material reaches this high temperature, and many techniques like aeration may be helpful for achieving this. To ensure safe sanitation according to the protocols of US Environmental Protection Agency [63], one should ensure that the slurry treatment is withheld for at least four complete hours at 55 ∘ C inside the reactor. Tyndallization, a technique with repeated heating peaks, enhances the inactivation of pathogens by exposing organisms to more heat stress [64]. To overcome spore formation, the irregular temperature pattern of composting may have a positive hygienic effect compared to maintaining a constant high temperature. The heat lost in solid composting is mainly by water evaporation, and either the temperature may be adjusted by insulation or by adding organic substances. There could be fair chance of pathogenic microbial regrowth at the colder regions of the pile, so it needs proper mixing, and especially during high temperatures, which may result in the proper exposure of the materials. As per USEPA, as a thumb rule of composting, a windrow or pile must be mixed at least five times and should have a temperature above 55 ∘ C for ensuring sanitation. According to Vanotti et al. [64], maturation of the compositing material and proper mixes many times decrease the risk of pathogen regrowth.

6.5.2

Testing the Possibilities of Commercial-Scale Resource Recovery

Advanced manure processing helps in reducing undesirable features in manure, such as water content, pathogen load, and odor, and helps to generate stable and transport-feasible manure with high economic value. The nutrients that are necessary for plant growth such as nitrogen (N) and phosphorous (P) are rich in manure.

155

156

6 Resource Recovery and Recycling from Livestock Manure

The bioenergies, nutrients, and fertilizers are the economically valuable by-products of manure processing. Many of the developing and developed countries like China and Europe started commercial-scale resource recoveries from the livestock manure. There are many such plants established on the basis of bioenergy generation from animal manures. One of the beneficial techniques for the production of bioenergy is the anaerobic digestion method. Liuminying plant is one of longest-running biogas projects in China, started with a single digester in 1192, with the extension of two more digesters by 2009 generating approximately 1680 m3 biogas, which is locally distributed to the nearby seven villages. Minhe, the largest broiler breeder in China also produces 11 million m3 of biogas along with 13 260 and 237 000 tons of solid and liquid fertilizers annually. The Deqingyuan (DQY) chicken farm is Asia’s largest egg farm, it is the first large-scale biogas project in China that could effectively utilize chicken manure. Roeblingen, the first large-scale bioenergy process with integrated N removal, was established in Germany; it used batch stripping process for ammonia recovery. Whereas in Haren, Germany, ammonia was removed by two-step processes of acidification and ammonia recovery via stripping and this was the first plant of this kind. Greendal Vergisting is a biogas plant located near Tully Quarry of North Ireland [21]. Nutrient recovery systems can be broadly classified as chemical or thermochemical systems [65]; some of the processes employed in commercial or pilot plants in different countries are described below: The chemical processes (i.e. Airprex, PASH, PRISA, Seaborne and SEPHOS [Germany]; Aquatron [Sweden]; BioCon, P-RoC [Denmark]; Crystalactor [Netherlands]; Kurita, Phosnix [Japan]; Ostara Pearl [Canada]; PEARS, Quick Wash [US]; REM-NUT [Italy]) are primarily based on recovery of phosphorus (and nitrogen in struvite-producing systems) through pH control. The thermochemical processes (i.e. AQUA RECI [Sweden]; ASH DEC and Mephrec [Germany]) employ oxidation processes at high temperatures to produce phosphorus pentaoxide or silico-phosphates. These processes help in efficient recovery of nutrients at commercial scale. Manure granulation is a widespread processing technology that can be adopted in manure processing for facilitating the export of manure nutrients to the market products; these nutrients can be altered through addition of macro- and micronutrients to match market demands. The FEECO International, Inc. (Green Bay, USA) is a pilot plant that uses digested dairy manure sludge for marketable manure granules. The digested sludge following dewatering, is followed by natural drying, mechanical pressing, fractionation, and nutrient enrichment before packaging for targeting the market needs.

6.6 Life Cycle Assessment of Sustainable Manure Management Systems LCA is considered to be an effective method that could be widely utilized for the comparison of environmental impacts of products or services providing the same function. During LCA, the entire life cycle of the products needs to be considered right from the raw materials extraction through the manufacturing,

6.7 Innovation in Sustainable Manure Management Systems and Recycling

and transportation up to final disposal; the secondary processing associated with the production systems also needs to be included. According to the ISO standards, LCA usually includes four stages, which are goal and scope definition, inventory analysis, impact assessment, and interpretation, respectively. To define the scope of the study, a functional unit must be chosen first and the functional unit represents the function that the product provides. The dynamic model development in order to clarify the time-dependent variables needs to be addressed in the life cycle structure [66]. The LCA of different sanitary treatment schemes used in manure treatment will be studied in terms of energy loss, emission standards, nutrient and mineral recovery, digestive characteristics of fertilization, and the impact of disease outbreaks on public and environmental health. So far so many studies been conducted by various researchers on optimizing sanitization standards along with the beneficial by product generations which are of advantageous to the society using animal manure. The manure runoff needs to be considered and is measured to avoid ponding, runoff, leakages to the subsurface, and harmful effect on plant nutrients. North Carolina passed a regulation EST that manages the manure from a large pig and poultry production of the state to meet the environmental standards to ensure sanitation in discharges, reduction in air and water pollution, and control of emission and pathogenic outbreaks and was found to be economically feasible [67]. The conversion technologies to treat digestate products and their economic feasibility in Flanders were studied by Ivan et al. [50], and they suggested the treatment of digestate before direct application in the fields decreases the impacts of sanitation issues, global warming, ammonia emission, and improves energy and mineral recovery efficiencies. The use of mixing triangle approach in decision-making clearly conveyed the messages to the stack holders about the most appropriate technologies to be opted for achieving efficiencies; the raw materials incoming and processing-related emissions, nutrient recovery, and fertilizer aspects of the digestate have been clearly studied and explored. Another study by Amaro de et al. [65] on bioethanol production from cattle manure and LCA stated some of the impacts on human and environment (in climate change [CC], human toxicity [HT], particulate matter formation [PMF], and fossil resource depletion [FD]) could be overcome by analyzing and adopting changes in the processing of inputs (dried raw material by process heating techniques). The process sources that cause greatest impacts from this study are found to be energy consumption, drying emission, pretreatment by sulfuric acids, enzymatic hydrolysis, and sodium phosphates during fermentation, which need to be looked into and managed for efficient production [68].

6.7 Innovation in Sustainable Manure Management Systems and Recycling 6.7.1

Economics of Resource Recovery from Manure and Sanitation

Before the circular economy and waste-to-energy (WTF) technologies emerged, vast majority of manure was disposed as stinky waste, merely part was utilized as a

157

158

6 Resource Recovery and Recycling from Livestock Manure

Reduction

Resources

Products

Recycle

Reuse

Figure 6.5

Wastes

The Revolving Model of circular economy.

natural fertilizer in agroecosystems. In 1990, two British environmental economists, David Pierce and Kelly Turner, clearly proposed the concept of circular economy. Circular economy is a new type of production and management activity that could achieve the aim of sustainable development by decreasing resource consumption and waste discharge as well as improving the efficiency of waste utilization [69]. It organizes economic activities into a “resources-products-renewable resources” cycle by using the laws of ecology, as shown in Figure 6.5. The 3R principles followed by the circular economy are reduce, reuse, and recycle, which involve reducing the input of material and energy (especially hazardous substances), extending the life cycle of the materials used during production process, and converting waste into renewable resources at the waste output and returning it to the economic system to reduce environmental pollution, respectively [70]. The 3R principles have been widely adopted in the technical model of livestock waste management. Based on the widespread application of fertilizer technology, combining livestock and poultry industry with planting in the form of industrial symbiosis or circular economy can not only achieve the virtuous cycle of ecological substances, but also has a positive significance on the formation of an economical, efficient, and stable circular economy industry.

6.7.2

Business Models for a Circular Economy

Since the 1990s, the strategy of sustainable development has gradually became a world tendency, and numerous developed countries have regarded the development of circular economy as a pivotal indicator to intensify their sustainable development strategy. It not only has been vigorously promoted by the national government, but also has been actively responded and practiced by the business community, and has made great achievements, forming a circular economy development model with national characteristics [71]. In general, there are three typical business models of

6.7 Innovation in Sustainable Manure Management Systems and Recycling

circular economy: social level, industrial level, and enterprise level. The systems of Germany and Japan are two typical circular economy models at the social level. The established dual dual system in German the recycling of substance and energy during and after consumption process. This model of dual-track recycling system is a non-governmental organization which specializes in recycling packaging waste. The organization usually undertakes the commission from the company; then, it employs shippers to classify and recycle their packaging wastes, and separate them to the corresponding recycling manufacturer, and meanwhile returns the packaging wastes that could be directly reused. The circular social model in Japan is a multi-level legal system promoted by government. The introduction of this model in Japan indicates that the development of environmental protection policies and industrial economy have entered a new period; the social structure begins to transform from the traditional economic society of “mass production, mass consumption and mass abandonment” to the circular economic society, which could positively reduce environmental load and realize sustainable development of the economy and society [72]. The Kalundborg Industrial Park in Denmark is one of the most representative industrial ecosystems all over the world. It assembles different types of factories together according to the principles of industrial ecology to constitute a symbiotic combination that can share resources and exchange by-products. The interoperability of these factories motivates waste gas, water, heat , and residue generated from one plant to become the raw materials and energy for another plant, and through this kind of energy transfer and materials exchange it establishes an environment-friendly circular industrial ecological park [72]. The main enterprises in Kalundborg Industrial Park are power plant, oil refinery, pharmaceutical factory, and gypsum board factory. Regarding these four companies as the core, by-products and wastes generated in each production process could be traded with each other and reused as raw materials. This kind of industrial model can not only significantly decrease the total amount of generated wastes and the cost of recycling, but also provide economic benefits (i.e. the cost of products is greatly reduced) and more importantly, it generates a virtuous cycle of economic development and environmental protection [28]. The DuPont Chemical Company in the United States is another typical example of circular economy at enterprise level. It reduces the demand for raw materials and energy from external sources during production process as far as possible, thus reducing the discharge of waste and toxic substances to outside environment. Through organizing material circulation between different processes in different enterprises and extending production chain, these circular industries are exploring to maximize the comprehensive utilization of renewable resources and energy and improve the durability of product to achieve clean production [73].

6.7.3 Enabling Environment Sanitation and Financing for Resource Recovery There will be no doubt that this revolution will significantly decrease the negative impact of industrial production on environment and thus can enable environment

159

160

6 Resource Recovery and Recycling from Livestock Manure

sanitation. It can significantly reduce the waste output to environment. In this revolution, policy, technology, and business model are three important factors in the circular economy. Only when these three elements are combined reasonably can we promote the establishment of a resource recycling system. Almost all companies are required to participate if all wastes are recycled [74]. So, we should enrich the regulatory functions of the government and adopt diversified policy to clarify the responsibility of relevant stakeholders to standardize its techniques and behaviors, promote the construction of a close vertical relationship as well as actively practice a centralized and harmless treatment mode. So, governments should be responsible for finance issues concerning resource recovery. They can transform the existing financial blueprint to allocate a portion of the tax to the development of resource recovery through formulating policies. Meanwhile, the revolution mainly aims on industry, governments can also appeal and mandatory related enterprises.

6.8 Challenges and Limitation The pollution of livestock and poultry manure is caused by the separation of planting and breeding, and its treatment effect is significantly affected by the behavior of relevant stakeholders. According to the essence of economic, the transformation of livestock and poultry waste as a by-product of breeding industry between important fertilizer resources and pollution sources is the result of externalities in market failure. However, from the point of view of the appearance, it is due to the disintegration of the breeding industry in the process of specialization, regularization and regionalization, and the waste cannot be smoothly consumed. The vertical relationship between stages in farm waste management is loose and the transaction chain is unstable; the planting waste treatment technology is mainly for manure production and the influencing factors are different. The scope of government’s governance and supervision has been broadened and strengthened, but the policy measures are not perfect and need to be constantly improved.

6.9 Conclusion and Future Prospects In the further development of the city, as the economy continues to develop and the population surges, the potential pressure on waste disposal will increase. Developing circular economy can not only accelerate the strategy of sustainable development, but also be an important approach to reduce pollution emissions and promote efficient utilization of resources. This paper proposes the following countermeasures and suggestions for its future development: technological innovation and actively promoting ecological planting mode. Technology is one of the most critical factors in substance recycling. Only making breakthroughs in technology can make the production cost of end product less than the production cost of raw materials. Based on the principle of maximizing profit, companies will actively apply the recycling technology to all aspects of product production.

References

Expanding the scale of the breeding industry and improve the microbial treatment of livestock manure: Expanding the planting scale of the farming industry can not only make full use of crop straws and by-products, but also increase the amount of livestock manure and organic fertilizer produced by microbial treatment. Optimizing the industrial structure: With the continuous improvement of the circular economy in the process of ecological development, industrial structure, product structure, and energy structure also need to be continuously improved and optimized accordingly. Establishing a policy mechanism to promote the development of circular economy, and further improve the incentive mechanism: The government should accurately grasp the development situation, give full play to the advantages of local characteristics, actively lead the construction of agricultural economy, and promote the exploration and realization of the leap-forward development mode of modern agriculture. Increasing publicity and education to promote public participation: The development of circular economy is not only the duty of government departments, but also relates to the living environment and quality of every citizen. The participation of the public is an important force in the process of conserving resources, protecting the ecological environment, and is of great significance to promoting the construction of agricultural circular economy.

Acknowledgements The authors are grateful to the National Natural Science Foundation of China (Grant No. 31750110469), China; the Shaanxi Introduced Talent Research Funding (A279021901), China; and The Introduction of Talent Research Start-up fund (No. Z101021904), College of Natural Resources and Environment, Northwest A&F University, Yangling, Shaanxi Province 712100, China for the financial support from Research Fund for International Young Scientists. The support from all the colleagues and research staff members is greatly acknowledged for their constructive advice and help.

References 1 Li, R., Wang, J.J., Zhang, Z. et al. (2012). Nutrient transformations during composting of pig manure with bentonite. Bioresource Technology 121: 362–368. 2 Gao, M., Liang, F., Yu, A. et al. (2010). Evaluation of stability and maturity during forced-aeration composting of chicken manure and sawdust at different C/N ratios. Chemosphere 78: 614–619. 3 Ren, X., Awasthi, M.K., Wang, Q. et al. (2018). New insight of tertiary-amine modified bentonite amendment on the nitrogen transformation and volatile fatty acids during the chicken manure composting. Bioresource Technology 266: 524–531.

161

162

6 Resource Recovery and Recycling from Livestock Manure

4 Bernal, M.P., Alburquerque, J.A., and Moral, R. (2009). Composting of animal manures and chemical criteria for compost maturity assessment. A review. Bioresource Technology 100: 5444–5453. 5 Aguirre-Villegas, H.A. and Larson, R.A. (2016). Evaluating greenhouse gas emissions from dairy manure management practices using survey data and lifecycle tools. Journal of Cleaner Production 143: 169–179. 6 Su, L.L., Lee, L.H., and Wu, T.Y. (2016). Sustainability of using composting and vermicomposting technologies for organic solid waste biotransformation: recent overview, greenhouse gases emissions and economic analysis. Journal of Cleaner Production 111: 262–278. 7 Fernández-Delgado Juárez, M., Gómez-Brandón, M., and Insam, H. (2015). Merging two waste streams, wood ash and biowaste, results in improved composting process and end products. Science of the Total Environment 511: 91–100. 8 Werner, F., Wang, X., Wolfgang, G. et al. (2018). Tackling ammonia inhibition for efficient biogas production from chicken manure: status and technical trends in Europe and China. Renewable and Sustainable Energy Reviews 97: 186–199. 9 Gabhane, J., Prince William, S.P.M., Bidyadhar, R. et al. (2012). Additives aided composting of green waste: effects on OM degradation, composting maturity, and quality of the finished compost. Bioresource Technology 114: 382–388. 10 Koivula, N., Räikkönen, T., Urpilainen, S. et al. (2004). Ash in composting of source-separated catering waste. Bioresource Technology 93: 291–299. 11 Himanen, M. and Hänninen, K. (2009). Effect of commercial mineral-based additives on composting and compost quality. Waste Management 29: 2265–2273. 12 Awasthi, M.K., Wang, Q., Chen, H. et al. (2017). Evaluation of biochar amended biosolids co-composting to improve the nutrient transformation and its correlation as a function for the production of nutrient-rich compost. Bioresource Technology 237: 156–166. 13 Velthof, G.L. (2011). Synthesis of the Research within the Framework of the Mineral Concentrates Pilot, Alterra Report 2224. Alterra: Wageningen. 14 Dominguez, J. (2011). Microbiology of Vermicomposting. Vermiculture Technology: Earthworms, Organic Wastes, and Environmental Management, 53–66. Boca Raton: Taylor and Francis LLC. 15 Chaoui, H.I., Zibilske, L.M., and Ohno, T. (2003). Effects of earthworm casts and compost on soil microbial activity and plant nutrient availability. Soil Biology and Biochemistry 35: 295–302. 16 Arancon, N.Q., Edwards, C.A., Lee, S., and Byrne, R. (2006). Effects of humic acids from vermicomposts on plant growth. European Journal of Soil Biology 42: S65–S69. 17 Das, M., Uppal, H.S., Singh, R. et al. (2011). Co-composting of physic nut (Jatropha curcas) de oiled cake with rice straw and different animal dung. Bioresource Technology 102: 6541–6546. 18 Onwosi, C.O., Igbokwe, V.C., Odimba, J.N. et al. (2017). Composting technology in waste stabilization: on the methods, challenges and future prospects. Journal of Environmental Management 190: 140–157.

References

19 Chen, Z., Zhang, S., Wen, Q., and Zheng, J. (2015). Effect of aeration rate on composting of penicillin mycelial dreg. Journal of Environmental Sciences 37: 172–178. 20 Chan, M., Selvam, A., and Wong, J.W.C. (2016). Reducing nitrogen loss and salinity during “struvite” food waste composting by zeolite amendment. Bioresource Technology 200: 838–844. 21 Wang, Q., Wang, Z., Awasthi, M.K. et al. (2016). Evaluation of medical stone amendment for the reduction of nitrogen loss and bioavailability of heavy metals during pig manure composting. Bioresource Technology 220: 297–304. 22 Chen, Y., Huang, X., Han, Z. et al. (2010). Effects of bamboo charcoal and bamboo vinegar on nitrogen conservation and heavy metals immobility during pig manure composting. Chemosphere 78: 1177–1181. 23 Lu, C., Gu, J., Wang, X. et al. (2018). Effects of coal gasification slag on antibiotic resistance genes and the bacterial community during swine manure composting. Bioresource Technology 268: 20–27. 24 Elving, J., Ottoson, J., Vinneras, B., and Albihn, A. (2010). Growth potential of faecal bacteria in simulated psycrophilic/mesophilic zones during composting of organic waste. Journal of Applied Microbiology 108: 1974–1981. 25 NBSC (National Bureau of Statistics of China) (2016). China Statistical Yearbook. Beijing: China Statistics Press. 26 USDA (2017). World Agricultural Supply and Demand Estimates. United States Department of Agriculture. 27 Chowdhury, M., De, N., and Jensen, L. (2014). Potential of aeration flow rate and biochar addition to reduce greenhouse gas and ammonia emissions during manure composting. Chemosphere 97: 16–25. 28 Wang, Q., Awasthi, M.K., Zhao, J. et al. (2018). Utilization of medical stone to improve the composition and quality of dissolved OM in composted pig manure. Journal of Cleaner Production 197: 472–478. 29 Appels, L., Baeyens, J., Degrève, J., and Dewil, R. (2008). Principles and potential of the anaerobic digestion of waste-activated sludge. Progress in Energy and Combustion Science 34: 755–781. 30 Fagbohungbe, M.O., Herbert, B.M., Hurst, L. et al. (2016). The challenges of anaerobic digestion and the role of biochar in optimizing anaerobic digestion. Waste Management 61: 236–249. 31 Suzuki, K., Tanaka, Y., Kuroda, K. et al. (2007). Removal and recovery of phosphorous from swine wastewater by demonstration crystallization reactor and struvite accumulation device. Bioresource Technology 98: 1573–1578. 32 Lashermes, G., Barriuso, E., and Houot, S. (2012). Dissipation pathways of organic pollutants during the composting of organic wastes. Chemosphere 87: 137–143. 33 Selvam, A., Xu, D., Zhao, Z., and Wong, J.W. (2012). Fate of tetracycline, sulfonamide and fluoroquinolone resistance genes and the changes in bacterial diversity during composting of swine manure. Bioresource Technology 126: 383–390.

163

164

6 Resource Recovery and Recycling from Livestock Manure

34 Awasthi, M.K., Wang, Q., Awasthi, S.K. et al. (2018). Influence of medical stone amendment on gaseous emissions, microbial biomass and abundance of ammonia oxidizing bacteria genes during biosolids composting. Bioresource Technology 247: 970–979. 35 Colon, J., Martinez-Blanco, J., Gabarrell, X. et al. (2010). Environmental assessment of home composting. Resources, Conservation and Recycling 54: 893–904. 36 Righi, S., Oliviero, L., Pedrini, M. et al. (2013). Life cycle assessment of management systems for sewage sludge and food waste: centralized and decentralized approaches. Journal of Cleaner Production 44: 8–17. 37 Hospido, A., Moreira, M.T., Martín, M. et al. (2005). Environmental evaluation of different treatment processes for sludge from urban wastewater treatments: anaerobic digestion versus thermal processes. International Journal of Life Cycle Assessment 10: 336–345. 38 Wong, J.W., Fung, S.O., and Selvam, A. (2009). Coal fly ash and lime addition enhances the rate and efficiency of decomposition of food waste during composting. Bioresource Technology 100: 3324–3331. 39 European Parliament and Council. (2009) Directive 2009/125/EC of the European Parliament and of the Council of 21 October 2009 establishing a framework for the setting of ecodesign requirements for energy-related products. 40 De la Rubia, M.A., Walker, M., Heaven, S. et al. (2010). Preliminary trials of in situ ammonia stripping from source segregated domestic food waste digestate using biogas: effect of temperature and flow rate. Bioresource Technology 101: 9486–9492. 41 Masse, D.I., Droste, R.L., Kennedy, K.J. et al. (1997). Potential for the psychrophilic anaerobic treatment of swine manure using a sequencing batch reactor. Canadian Agricultural Engineering 39: 25–33. 42 Hamelin, L., Wesnæs, M., Wenzel, H., and Petersen, B.M. (2011). Environmental consequences of future biogas technologies based on separated slurry. Environmental Science & Technology 45: 5869–5877. 43 Foust, A.S., Wenzel, L.A., Clump, C.W. et al. (1980). Principle of Unit Operations, 2e, 768. New York: John Wiley & Sons. 44 Ukwuani, A.T. and Tao, W. (2016). Developing a vacuum thermal stripping – acid absorption process for ammonia recovery from anaerobic digester effluent. Water Research 106: 108–115. 45 Ribaudo, M.O. and Johansson, R.C. (2007). Nutrient management use at the rural-urban fringe: does demand for environmental quality play a role? Review of Agricultural Economics 29: 689–699. 46 Parshina, S.N., Nekrasova, V.K., Tsarikov, N. et al. (1997). Manure digestion under thermophilic and psycbrophilic conditions. Proceedings of 5111 FAO/SREN-workshop Anaerobic Conversion for Environmental Protection. Sanitation and Reuse of Residues, Gent, Belgium, pp. 59–67. 47 UNECE (2011). Guidance Document for Preventing and Abating Ammonia Emissions from Agricultural Sources. United Nations Economic Commission for Europe, Convention on Long-range Transboundary Air Pollution. Geneva: UNECE.

References

48 EC. (2009). Regulation (EC) no. 1069/2009 of the European parliament and of the council of 21 October 2009. Laying down health rules as regards animal by-products and derived products not intended for human consumption and repealing Regulation (EC) No 1774/2002 (Animal By-products Regulation). Official Journal of the European Union, European Community, Brussels. 49 EU. (2011). EU 142/2011. Regulation (EU) No. 142/2011 of 25 February 2011 implementing Regulation (EC) No 1069/2009 of the European Parliament and of the Council laying down health rules as regards animal by-products and derived products not intended for human consumption and implementing Council Directive 97/78/EC as regards certain samples and items exempt from veterinary checks at the border under that Directive. Official Journal of the European Union. European Community, Brussels. 50 Ivan, V.-R., Katarzyna, G., Viooltje, L. et al. (2015). Environmental assessment of digestate treatment technologies using LCA methodology. Waste Management 43: 442–459. 51 EPA (2009). EPA targets clean water act violations at livestock feeding operations. Enforcement Alert 10: 1–4. 52 Grossman, M.R. (2003). Legal Issues of Livestock Odor and Waste Management. Washington, DC: National Research Council, National Academy of Sciences. 53 Natural Resource Conservation Service. (2012). Nutrient Management (ACRE) Code 590. https://efotg.sc.egov.usda.gov/references/public/MN/590mn.pdf (retrieved 5 July 2012). 54 Fang, M., Wong, J.W.C., Ma, K.K., and Wong, M.H. (1999). Co-composting of sewage sludge and coal fly ash: nutrient transformations. Bioresource Technology 67: 19–24. 55 MAFF. (2012). State of Animal Waste Management (in Japanese). MAFF, Tokyo; http://www.maff.go.jp/j/chikusan/kankyo/taisaku/t_mondai/02_kanri/index .html (accessed August 2012). 56 Elving, J., Emmoth, E., Albihn, A. et al. (2012). Composting for avian influenza virus elimination. Applied and Environmental Microbiology 78: 3280–3285. 57 Bujozek, G. (2001). Influence of ammonia and other abiotic factors on microbial activity and pathogen inactivation during processing of high-solid residues. Dissertation, University of Manitoba, Manitoba. 58 Ottoson, J., Schnurer, A., and Vinneras, B. (2008). In situ ammonia production as a sanitation agent during anaerobic digestion at mesophilic temperature. Letters in Applied Microbiology 46: 325–330. 59 Colleran, E. (2000). Hygienic and sanitation requirements in biogas plants treating animal manures or mixtures of manures and other organic wastes. In: Anaerobic Digestion: Making Energy and Solving Modern Waste Problem, 77–86. Ad-nett. 60 Mitscherlich, E. and Marth, E.H. (1987). Microbial survival in the environment – bacteria and rickettsiae important in human and animal health. International Journal of Food Microbiology 4: 82.

165

166

6 Resource Recovery and Recycling from Livestock Manure

61 United States Environmental Protection Agency (USEPA) (1994). Water Quality Standards Handbook, chapter 3. Washington DC: US Environmental Protection Agency, Office of Water. 62 Nieminen, J. (2010). Phosphorus recovery and recycling from municipal wastewater sludge. Vesitalous 3: 7–15. 63 Levasseur, A., Lesage, P., Margni, M., and Samson, R. (2012). Biogenic carbon and temporary storage addressed with dynamic life cycle assessment. Journal of Industrial Ecology 17: 117–128. 64 Vanotti, M.B., Szogi, A.A., Millner, P.D., and Loughrin, J.H. (2009). Development of a second-generation environmentally superior technology for treatment of swine manure in the USA. Bioresource Technology 100: 5406–5416. 65 Amaro de, A., Franccesca, F., Mateus da Silva, S. et al. (2017). Life cycle assessment of bioethanol production from cattle manure. Journal of Cleaner Production 162: 1021–1030. 66 Barth, J. (2009). Survey on Biological Waste Treatment in Europe. Annually Survey by the European Compost Network Updated by Reports from the Member States. ECN/ORBIT e.V. - European Compost Network. 67 Wang, L.K., Wang, M.-H.S., and Hung, Y.T. (2016). Natural Resources and Control Process, Handbook of Environmental Engineering, vol. 17. Springer. 68 Wallis-Lage C. (2013) Resource recovery: the next wave. Turning wastewater treatment plants into recovery centers for valuable resources. Water and Wastes Digest. http://www.wwdmag.com/resource-recovery-next-wave (accessed 22 June 2014). 69 Roland, W.S., Amit, R., Fridolin, S.B. et al. (2014). Sustainable phosphorus management – a global trans-disciplinary roadmap. Ecological Economics 114: 245–246. 70 Tiquia, S.M. (2010). Reduction of compost phytotoxicity during the process of decomposition. Chemosphere 79: 506–512. 71 Fukumoto, Y., Suzuki, K., Osada, T. et al. (2006). Reduction of nitrous oxide emission from pig manure composting by addition of nitrite-oxidizing bacteria. Science of Total Environment 40: 6787–6791. 72 Kader, N.A., Robin, P., Paillat, J.M., and Leterme, P. (2007). Turning, compacting and the addition of water as factors affecting gaseous emissions in farm manure composting. Bioresource Technology 98: 2619–2628. 73 Pagans, E., Barrena, R., and Font, X. (2006). Ammonia emissions from the composting of different organic wastes. Dependency on process temperature. Chemosphere 62: 1534–1542. 74 Liao, Q., Wei, G., Jiang, Z. et al. (2013). Research progress on resource utilization of livestock and poultry manure. Journal of Southern Agriculture 44: 338–343.

167

7 Utilization of Microalgae and Thraustochytrids for the Production of Biofuel and Nutraceutical Products Ying Liu and Jay J. Cheng Biological and Agricultural Engineering Department, North Carolina State University, Raleigh, NC, USA

7.1

Introduction

7.1.1

Microalgae

Microalgae are a group of diverse unicellular microorganisms that are commonly found in aquatic and marine systems. Generally, microalgae have a high growth rate with a doubling time of as short as 3.5 hours during the exponential growth phase. The photosynthetic efficiency of microalgae is typically around 3% (as high as 8.3% in some cases), which is much higher than that of a terrestrial plant (around 1%) [1]. Microalgae use the captured solar energy to synthesize a variety of chemical compounds including lipids and valuable bioactive products, which can be used as a renewable feedstock for biofuels (e.g. biodiesel or jet fuel) and nutraceutical products [2]. Most microalgae can grow autotrophically through assimilation of CO2 from natural and industrial sources and consequently reduce the negative effects of CO2 on the environment. Some microalgae have a high capacity of absorbing nutrients (nitrogen, phosphorus, and minerals) and are used in the treatment of domestic wastewater as well as water bodies with eutrophication.

7.1.2

Thraustochytrids

Thraustochytrids, osmo-heterotrophic unicellular marine protists, are distributed widely in a variety of marine habitats [3]. They were misplaced under the label “algae” until recently. Ecological investigations have revealed that thraustochytrids exist in high abundance and are closely correlated with particulate organic carbon and chlorophyll a concentration in marine habitats. They play a significant and unique role in marine biogeochemical cycles. Thraustochytrids have also been identified as an important source of various valuable products including enzymes, extracellular polysaccharides, biodiesel, and polyunsaturated fatty acids (PUFAs) [4].

Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

168

7 Utilization of Microalgae and Thraustochytrids

7.1.3

Biodiesel and Biobased Jet Fuel

Because of increasing concerns on the depletion of fossil fuel resources and the adverse environmental impacts caused by the extensive use of fossil fuels, the development of biofuels from renewable feedstock has attracted great attention in the past decades. The U.S. Energy Information Administration reported that the biodiesel consumption in the United States increased from 10 million gallons in 2001 to 2060 million gallons in 2016. Biodiesel is widely used as a transportation fuel. The commercial use of biofuel for aircraft engines as jet fuel was officially approved by ASTM (American Society for Testing and Materials) International in 2011. Compared to terrestrial energy crops, microalgae and thraustochytrids have superior advantages in regard to yield, which make the latter promising candidates for biofuel production. Both microalgae and thraustochytrids have much higher biomass yields than terrestrial crops and the cultivation of the microorganisms does not require arable land. They also exhibit promising potentials of accumulating much higher lipids, triacylglycerides (TAGs) in particular in biomass than most other microorganisms. Many microalgae and thraustochytrid species are isolated from marine systems and can be cultivated in saline or brackish water, which alleviates the consumption of freshwater. All these advantages suggest that microalgae and thraustochytrids could be promising feedstock alternatives for biofuel production. There have been extensive research activities on the production of biodiesel and biobased jet fuel from microalgae and thraustochytrids in the last two decades. A number of industrial companies around the world are working on the commercial production of microalgae/thraustochytrids-based biofuel.

7.1.4

Docosahexaenoic Acid (DHA) and Eicosapentaenoic Acid (EPA)

Docosahexaenoic acid (DHA) and eicosapentaenoic acid (EPA) are valuable ω-3 PUFAs and basic nutrients for human health. DHA is a primary component of human brain, cerebral cortex, and retina. A good level of DHA can reduce the risk of heart attack, cardiovascular disease, and Alzheimer‘s disease. EPA is also believed to have clinical significance against valproate-induced hepatic malfunction, necrosis, and steatosis. ω-3 PUFAs such as DHA and EPA have a huge market as nutraceutical supplements. Currently, the main industrial source for DHA/EPA is selected marine fish oil, which is not ideal due to its undesirable smell, possible metal contamination, and unstable production. The production of DHA/EPA from microalgae or thraustochytrids can overcome these shortcomings. Microalgae are the original producers of DHA/EPA in marine ecosystems, and thraustochytrid cells are also rich in ω-3 PUFAs. The cultivation of microalgae or thraustochytrids under controlled conditions makes it possible to mass-produce high-quality DHA/EPA. Overall, microalgae and thraustochytrids have demonstrated unique features as promising alternative sources for biofuel and DHA/EPA production. Previous review articles on microalgae or thraustochytrids have focused on certain aspects of their biotechnological applications, e.g. lipid accumulation [2, 5], harvesting and lipid extraction techniques [6, 7], the conversion of crude lipid into biofuels [8], and

7.2 Microalgae for Biodiesel and Jet Fuel Production

the integration of cultivation with carbon sequestration or wastewater treatment [9, 10]. The aim of this chapter is to provide a comprehensive review of microalgae and thraustochytrids, focusing on recent advances in the production of biofuel and nutraceutical products (DHA/EPA) from the microbes with an emphasis on the current challenges and future perspectives.

7.2

Microalgae for Biodiesel and Jet Fuel Production

7.2.1

Selection of Microalgae

For the production of biofuel (biodiesel or jet fuel) from microalgae, the selection of adequate microalgal strains has always been the first and primary step. Lipid yield, lipid profile, and growth rate of the microalgal strains are the key factors affecting the biofuel output and quality. All these factors should be considered simultaneously to identify the most adequate strains for efficient production of high-quantity biodiesel/jet fuel [11, 12]. First and foremost, the lipid content in microalgae varies from 2 to 75% of the dry weight biomass as reported by Mata et al. [13]. Lipid productivity, which is directly affected by biomass productivity and lipid content in the microalgae, has also shown a significant difference among various microalgal strains. Some strains, mostly from Botryococcus, Chlorella, Dunaliella, Nannochloropsis, and Scenedesmus species with both high lipid content and high biomass productivity have been extensively studied for the production of biodiesel or jet fuel [14, 15]. For an economic production of biofuel, algal oil productivity has been identified as the key parameter to affect the overall biofuel production cost, thus microalgal strains with high lipid productivity would be preferred for biofuel production [16]. Secondly, lipid profile of the microalgae contributes substantially to the quality of biodiesel production. Saturated fatty acids (SFAs) and monounsaturated fatty acids (MUFAs) with carbon chain from 12 to 22 (C12–C22) are preferred fatty acids for biodiesel production. The presence of PUFAs in microalgal oil will likely reduce the oxidative and thermal stability of the biodiesel. Jet fuel has a different composition from biodiesel and is generally a mixture of alkanes, alkenes, cycloalkanes, and aromatics with shorter carbon chain from 8 to 16 (C8–C16). Thus, the microalgae with desired fatty acids of the microalgal oil will be beneficial for either biodiesel or jet fuel production [2]. Lastly, the robustness of microalgal strains determines whether they can be cultivated under different conditions. Recently, most studies on microalgal biofuel have tried to integrate algal lipid production with wastewater treatment and carbon mitigation [17, 18]. This strategy addresses the issue of not only sustainable production of biofuels but also alleviation of environmental concerns. Thus, the desirable microalgae strains are those that can grow in a wide range of cultivation conditions, especially in complex wastewaters or high-CO2 concentrations. Some recent studies have reported that the microalgae that are tolerant to oxidative stress are more efficient for biofuel production [19]. These microalgae are

169

170

7 Utilization of Microalgae and Thraustochytrids

better candidates to grow in wastewaters and subsequently meet the demand of economical production of biofuels while minimizing adverse environmental impact.

7.2.2 7.2.2.1

Processes of Microalgae to Biofuel Microalgae Cultivation

Achieving high cell density and high lipid content of microalgae in cultivation is crucial for microalgal biofuel production. Several factors in the cultivation process have been proved to have great impact on microalgae’s growth and lipid production, including light, temperature, and nutrients. The optimal cultivation conditions for each individual microalgal strain are different and their lipid content is in a range of 20–70% of dry cell weight. Environmental stress has frequently been applied during their cultivation to enhance lipid accumulation in microalgae [5, 20, 21]. Oxidative stress induced by temperature, salinity, pH, or light intensity is likely to increase lipid content in most microalgae as lipid is the preferred energy storage product under stressed conditions [15, 22–24]. For example, the chemical composition of Scenedesmus sp. changed evidently when NaCl concentration in the culture medium increased, and the lipid content of the microalgae increased 1.8-fold [15]. Either increase or decrease from the optimal temperature for Nannochloropsis oculata cell growth resulted in increased lipid content [25]. Similarly, the carbon flux in microalgal cell would be diverted from protein and carbohydrate syntheses to lipid synthesis under nutritional stress conditions, resulting in the accumulation of large amounts of TAGs that can be utilized in biodiesel production [26–28]. Particularly, nitrogen deprivation or limitation has been proved by numerous studies to stimulate lipid accumulation in microalgae. The total lipid content of Chlorella pyrenoidosa could achieve over 50% of dry cell weight under nitrogen deficiency conditions [29]. Different approaches with environmental stress for enhanced lipid accumulation in microalgae are summarized in Table 7.1. However, these treatments cannot always result in lipid productivity enhancement as biomass growth might be impaired by the unfavorable cultivation conditions. Hence, to resolve the conflict of biomass growth and lipid accumulation, recent studies have been looking into the possibility of using alternative cultivation mode for more efficient overall lipid production [37]. Multiple-stage cultivation with stress operation is suggested as a promising approach to obtain high lipid productivity. Specifically, two-stage cultivation with the first stage focusing on biomass growth followed by a second stage of lipid induction has been widely investigated recently [30]. The lipid productivity of Chlorella vulgaris increased over twofold when the operation was switched from the nutrient repletion stage to the second stage with nitrogen starvation [36]. In addition, the synergistic effects of two or more environmental factors would also affect the total lipid productivity (Table 7.1). It was reported that under nitrogen starvation stress, the excessive supply of phosphorus would positively affect microalgal cell formation and consequently enhance lipid production [27, 38].

Table 7.1

Effect of environmental stress on lipid accumulation in microalgae.

Stress factor

Salinity

Temperature

Stress conditions

Microalgal species

Lipid content (% of dry cell weight)

Lipid productivity (mg (L d))

Reference

400 mM NaCl

Scenedesmus sp.

33.13 (1.75-fold increase)

Not available (N.A.)

[15]

150 mM NaCl

Hindakia

68.39 (1.80-fold increase)

26.00 (twofold increase)

[22]

17 mM NaCl

Mixed microalgae

23.40 (1.98-fold increase)

N.A.

[30]

From 30 to 25 ∘ C From 20 to 25 ∘ C

Chlorella vulgaris

14.71 (1.49-fold increase)

20.22 (1.48-fold increase)

[25]

Nannochloropsis oculata

13.89 (0.76-fold increase)

10.10 (24% increase)

Chlorella sp.

33.03 (1.44-fold increase)

71.85 (1.63-fold increase)

Light irradiation intensity

From 400 to 40 μmol photon m2 /s)

Light spectrum

200 μmol (m2 /s) blue vs. white light

Ultrasonic stress

5 min, power = 200 W

Anabaena variabilis

46.90 (1.46-fold increase)

54.20 (1.86-fold increase)

[33]

N deficiency

N free vs. N rich

Chlorella pyrenoidosa

50.32

47.05 (1.59-fold increase)

[29]

N free vs. N rich

Scenedesmus sp.

27.93 (1.48-fold increase)

N.A.

[34]

N from 1.0 to 0.2 g/L

Scenedesmus sp.

52.60 (3.84-fold increase)

N.A.

[35]

N starvation

Chlorella vulgaris

43.00 (1.7-fold increase)

77.80 (2.4-fold increase)

[36]

N and P limitation

Chlorella minutissima

47.13

49.10 (1.47-fold increase)

[26]

Synergistic stress

Monoraphidium dybowskii

43.47 (1.42-fold increase)

81.81 (1.93-fold increase)

Chlorella vulgaris

23.50 (1.12-fold increase)

N.A.

[31]

[32]

N limitation and P excess

Chlorella regularis

42.30 (2-fold increase)

310.0 (1.30-fold increase)

[27]

N depletion and osmotic stress

Chlorella protothecoides

39.2 (1.92-fold increase)

177.3 (1.60-fold increase)

[37]

172

7 Utilization of Microalgae and Thraustochytrids

Table 7.2 Advantages and disadvantages of conventional microalgal biomass harvesting techniques. Harvesting technique

Centrifugation

Advantages ● ●

Sedimentation

●

●

Flocculation

● ●

Effective and reliable Good capability of harvesting most types of microalgal cells The simplest way to harvest microalgal biomass Cheap in operation High microalgae recovery rate Relatively inexpensive

Disadvantages ● ●

● ●

● ● ● ● ●

Flotation

● ● ●

Filtration

●

High efficiency (70–99%) Short time requirement Relatively inexpensive High efficiency of microalgae harvesting

●

●

●

7.2.2.2

High energy consumption High capital and operationalcosts, not cost-effective Time-consuming Low biomass recovery rate Slow and unreliable process Strain-specific recovery efficiency High flocculant requirement Possible flocculant contamination Difficult recovery of flocculants after harvesting Requirement of surfactant or coagulant Possible clogging or fouling by settled microalgal cells Short life span of membranes

Microalgae Harvesting

Harvesting microalgal biomass from cultivation systems is an essential step for the subsequent biofuel production. However, the harvesting of microalgal biomass requires great energy or time input, due to the small size of microalgal cells and the low cell density in the cultivation systems. The harvesting process generally accounts for about 20–30% of the total biofuel production costs, making the harvesting step one of the bottlenecks for the commercialization of microalgal biofuel production [6, 39]. Conventional harvesting techniques including centrifugation, sedimentation, filtration, flotation, and flocculation are quite mature but costly. The advantages and disadvantages of the conventional techniques for microalgae harvesting are listed in Table 7.2. Some of these harvesting methods require huge energy or time input or high investment cost. Negative “side effects” on end product may be caused with some techniques during the harvesting process. For instance, chemical flocculants used during flocculation would increase the risk of contamination of microalgal biomass. Thus, great efforts have been made in the investigation of alternative efficient and cost-effective harvesting methods. Bio-flocculation

Of the recently developed harvesting methods, bio-flocculation using microorganisms (bacteria and/or fungi) or bio-flocculants for microalgae entrapment has been extensively explored in the last decade. It is considered as an effective and environment-friendly harvesting approach. The presence of bacteria in the medium contributed profoundly to the increase of microalgae floc size, resulting

7.2 Microalgae for Biodiesel and Jet Fuel Production

in quick sedimentation and high flocculating efficiency. The role of bacteria/fungi in microalgae bio-flocculation is still speculative. Two possible mechanisms were proposed for the bacteria-associated bio-flocculation. Lee et al. [40] suggested that the extracellular polymeric substances (EPSs) of bacteria or bacteria themselves would directly associate with microalgae cells and promote the aggregation. They reported that bacteria and their EPS collectively played major roles in the flocculation of Chlorella regularis. Charge neutralization was considered as the other mechanism [41]. Fungi were also proved to induce flocculation of various microalgal strains. Filamentous fungi could pelletize together with microalgae when they are cultivated in the same system [42, 43]. The microalgal cells were either attached or entrapped in the pellets, forming a symbiotic lichen structure. Fungal strains from Aspergillus species were frequently tested for their capability of forming large pellets with microalgae [43, 44]. Almost complete harvesting of autotrophically grown Chlorella vulgaris was achieved with Aspergillus oryzae palletization under optimal conditions [43]. Moreover, the total lipid production was even improved when oleaginous fungus and microalgae were co-cultured in the same system [44]. However, the co-cultivation of bacteria or fungi with microalgae has also raised some concerns in commercial-scale production. For instance, additional nutrients will be required for the cultivation of bacteria or fungi. The presence of exogenous bacteria or fungi may cause contamination in microalgal products. Further studies on in situ aggregation of microalgae by bacteria or fungi are necessary for large-scale application of the technology. Lately, natural bio-flocculants originated from renewable organisms have been reported as alternative agents for microalgal biomass harvesting [45]. Specifically, some bacteria were able to aggregate microalgal cells through EPS to produce flocculants, which showed great potential in large-scale application [46]. It was reported that γ-PGA produced by Bacillus licheniformis achieved high flocculation efficiency of over 97% with 200 L of Desmodesmus brasiliensis. However, high dosage of the bio-flocculants was still required for efficient microalgae harvesting and most bio-flocculants presented species-specific aggregation property. In addition to bacteria and fungi, some microalgae also possess the ability to produce EPS and form bio-flocs to stimulate bio-flocculation [47]. These microalgae can aggregate together by themselves, leading to auto-flocculation [48]. Besides, the increase of pH in the culture medium also induced microalgae auto-flocculation [49]. These self-flocculating microalgae could also be used for the bio-flocculation of other non-flocculating microalgae [50, 51], which has been suggested as a superior harvesting method because no contamination of bacteria or fungi is introduced in the microalgae bulk. Magnetic Separation

Magnetic separation was revealed as another promising option for collecting negatively charged microalgal cells from culture broth. High harvesting efficiency of over 95% could be achieved in a very short time with magnetic particles [52]. More importantly, magnetic particles could be easily recovered from microalgae flocs and regenerated for their reuse in harvesting, in contrast to traditional

173

174

7 Utilization of Microalgae and Thraustochytrids

chemical flocculants [53–55]. Therefore, the cost for microalgae harvesting could be significantly reduced. The surfactant-decorated Fe3 O4 magnetic nanoparticles developed by Seo et al. [54] indeed facilitated lipid extraction by disrupting cell membranes with the cationic surfactant. High lipid and TAG yields were obtained from the harvested biomass with the magnetic nanoparticles without any additional cell disruption step, suggesting that this approach could be a green process for efficient integration of microalgae harvesting and lipid extraction [54]. 7.2.2.3

Extraction of Oil from Microalgae

A commonly used method for microalgal oil extraction is organic solvent extraction. A classic Bligh and Dyer’s method employing a combination of chloroform and methanol was extensively used in lipid extraction from various microalgae. Hexane with low affinity to non-lipid compounds has also been widely used in Soxhlet extraction [56, 57]. The nonpolar molecules of chloroform and hexane have shown high selectivity to the extraction of neutral lipid fractions, which are the favorable precursors for biofuel production. Organic solvents-based extraction can generate high recovery efficiency for the majority of microalgal oils, but the toxic nature of these organic solvents is a big concern for their applications. In recent years, supercritical fluid (e.g. CO2 ) technology has attracted increasing interest [58–60]. It was indicated as a more efficient method because supercritical fluid can diffuse deeper in the cell matrix. Taher et al. [60] demonstrated that a higher lipid extraction yield from Scenedesmus sp. was obtained with supercritical CO2 , compared to conventional solvent extraction. A concern associated with supercritical fluid extraction is the high operating pressure, which probably cause an increase in cost. Pretreatment of microalgal biomass has been studied for efficient release of lipid droplets prior to oil extraction. A drying process is usually conducted as the water content in the harvested microalgae can be as high as over 90%, leading to the formation of a polar barrier preventing the interaction between the solvent and the lipid [56, 61]. On the other hand, the cell wall of many microalgae is a barrier for efficient oil extraction. Thus, cell wall disruption is generally applied to facilitate the liberation of oil, and the methods include ultrasonication, microwave treatment, and autoclaving [57, 62]. However, most pretreatment processes are energy intensive, creating another economic challenge for microalgal biofuel production. Extensive attempts have been made to eliminate the drying step and to search for alternative cell disruption methods for microalgal oil extraction. Several recent studies have explored new technologies to extract oil directly from wet microalgae, which eliminated the drying process and led to substantial reduction in total energy consumption [63, 64]. Some of the technologies achieved comparable oil recovery efficiency from wet biomass to that from dried biomass [56]. Yet, most of these methods had relatively low extraction efficiency with wet biomass. Therefore, extensive studies have been focused on the investigation of efficient cell disruption of microalgae in media to facilitate oil extraction from wet biomass [65, 66]. For instance, enzymatic treatment of fresh microalgae resulted in efficient hydrolysis of cell walls and remarkably increased lipid extraction yield [65]. Surfactants also showed good potential in disrupting microalgal cell wall for a high yield of oil

7.2 Microalgae for Biodiesel and Jet Fuel Production

Microalgae culture solution

1

2

3

4

5

Microalgae harverting

Biomass dewatering

Cell disruption Simultaneous cell disruption and oil extraction Oil extraction

Transesterification

Figure 7.1 A scheme of alternative microalgal oil extraction and biodiesel production 1 conventional procedure including biomass harvesting, dewatering, and cell processes:  2 cell disruption in microalgae solution followed by oil disruption, and oil extraction;  3 direct oil extraction from wet microalgae biomass;  4 simultaneous cell extraction;  5 direct transesterification disruption and oil extraction from wet microalgae biomass;  from wet microalgae biomass.

extraction from wet microalgae [66, 67]. In addition, novel methods were developed for simultaneous cell disruption and lipid extraction from wet microalgae and showed significant enhancement in oil extraction [57, 68]. Figure 7.1 shows a scheme of alternative microalgal oil extraction and biodiesel production processes. 7.2.2.4

Biodiesel Production from Microalgal Oil

Biodiesel is a mixture of fatty acid esters. Transesterification of triglyceride and alcohol to fatty acid esters and glycerol is the most widely used method in which a

175

176

7 Utilization of Microalgae and Thraustochytrids

O

O

C R1 O O C R2 O

CH2 O CH

CH2 O

R1 Catalyst +

3 CH3OH

R2

C R3

Triglyceride

Figure 7.2 glycerol.

R3

Methanol

C O CH3 O C O CH3 O C O CH3

Fatty acid methyl esters (Biodiesel)

CH2 OH +

CH

OH

CH2 OH

Glycerol

Transesterification reaction to convert triglyceride and alcohol to biodiesel and

catalyst is used for the chemical reaction as shown in Figure 7.2. Free fatty acids can also be utilized for biodiesel production through esterification reactions. Homogenous chemical catalysts such as alkalines and acids are usually used for transesterification and esterification, respectively, because of their high efficiency and low cost. However, most homogeneous catalyses will need an extensive subsequent product purification process and generate wastewater, especially when the raw material contains a high content of free fatty acids. Among the catalysts addressing the drawbacks of homogeneously catalyzed transesterification/esterification, enzymatic catalyst (lipase) and heterogeneous catalysts have attracted growing attention. Both of the novel catalysts can reduce the downstream processes and wastewater generation. Lipase-catalyzed transesterification can also perform in mild environmental conditions and make the biodiesel purification process much easier. A recent report has shown a high conversion of lipids from Scenedesmus obliquus to biodiesel with immobilized Aspergillus niger cells that generate lipase [69]. Ion-exchange resins have successfully been used as heterogeneous catalyst for efficient transesterification/esterification of microalgal oil with a high concentration of free fatty acids, which overcomes the soap formation problem during homogenous alkali transesterification [70, 71]. Tungstated zirconia catalyst is also reported to convert lipids from Scenedesmus obliquus to good-quality biodiesel [72]. Besides the so-called two-step microalgal biodiesel production, i.e. oil extraction from microalgae and transesterification of the microalgal oil to biodiesel, one-step in situ or direct transesterification from microalgal biomass has also been studied recently [73, 74]. It was reported that less time and energy were consumed with in situ transesterification because of the reduction of the units in the process. Moreover, direct transesterification with wet microalgae was also studied without the drying process [63]. A green technology of simultaneous cooling and microwave heating, which accelerated the cell wall disruption, was used in a direct transesterification of wet Nannochloropsis sp. and led to high efficiency of biodiesel production as well as excellent biodiesel quality [73]. 7.2.2.5

Jet Fuel Production from Microalgal Oil

In addition to biodiesel production, microalgae have also been considered as an alternative feedstock for jet fuel production. The possibility of converting microalgal oil

7.3 Thraustochytrids for Biodiesel Production

for green jet fuel production was recently reviewed by Bwapwa et al. [75]. However, stringent regulations on jet fuel quality such as freeze point, flash point, energy density, aromatic content, and specific gravity make the commercialization of microalgal jet fuel production very difficult [76]. Several approaches were tested for the conversion of microalgae to green jet fuel, including pyrolysis, gasification, and liquefaction [75, 77]. To meet the jet fuel standard, an upgrading step is generally necessary for the conversion processes. Specifically, deoxygenation and decarboxylation to produce alkanes from fatty acids are usually required in microalgal jet fuel production [77]. Environmental concerns and benefits should also be considered for viable microalgal jet fuel production. A life cycle analysis of microalgae-derived jet fuel showed that 76% reduction in greenhouse gas emission was realized for the optimized hydrothermal liquefaction of microalgae grown in wastewater, compared to conventional jet fuel [78].

7.3

Thraustochytrids for Biodiesel Production

Some species of thraustochytrids have also been identified as great candidates for biodiesel production. The lipid content in the thraustochytrids cells can be over 50% of dry biomass. Moreover, some thraustochytrids strains contain high amounts of SFAs and MUFAs, including palmitic acid (C16 : 0) and oleic acid (C18 : 1), which are useful for biodiesel production [79, 80]. Marine ecosystems are believed to harbor abundant and diverse thraustochytrids [3], but only limited thraustochytrids species have been isolated and investigated for their biotechnological applications [4]. Recently, thraustochytrids species originated from Australian aquatic habitats, Indian marine environments, and the coastal habitats of South China have been evaluated for their biodiesel-producing potential [4, 80, 81]. In general Aurantiochytrium sp. and Schizochytrium sp. with their remarkable features such as fast growth, high lipid content, and simple fatty acid profiles, demonstrated excellent biodiesel production potential [4, 80, 82, 83]. They have a much lower doubling time than many other thraustochytrid species. Additionally, SFAs are the predominant component in their fatty acid profiles and can constitute over 30% of the total fatty acids. According to Liu et al. [4], Schizochytrium sp. and Aurantiochytrium sp. had a high percentage of C16 : 0 (around 50% of the total fatty acids), making them great potential sources for biodiesel production. Thraustochytrids are heterotrophic microorganisms and glucose is usually used as the carbon source in their cultivation. However, the utilization of glucose raises questions on the production cost. To reduce the cost for thraustochytrids-based biodiesel production, recent studies have focused on the utilization of alternative carbon sources from industrial and agricultural wastes for thraustochytrids cultivation [82, 84, 85]. Some alternative sources have shown good potential for thraustochytrids-based biodiesel production with similar or even improved biomass and fatty acid yields, as compared to glucose [82]. Specifically, glycerol, the major byproduct of the transesterification process for biodiesel production has been

177

178

7 Utilization of Microalgae and Thraustochytrids

extensively studied and proven to be a promising carbon source for thraustochytrids cultivation [84, 86]. According to a recent report, glycerol outperformed glucose as a carbon source for an improved production of thraustochytrids biomass and fatty acids that could be used for biodiesel production [87]. One of the advantages of thraustochytrids is their capability of using both organic and inorganic nitrogen for their growth. Some inexpensive nitrates, as a nitrogen source, generated similar yields of thraustochytrids biomass and palmitic acid when compared with commonly used yeast extract [82]. In addition, the feasibility of recycling the hydrolysate effluent after thraustochytrids lipid extraction for the next cultivation of the same microbial culture was investigated and the results indicated that the recycled hydrolysate provided essential nutrients for the growth of thraustochytrid cells as well as oil production [88–90]. Hence, nutrient consumption in the thraustochytrids cultivation can be substantially reduced, resulting in a significant reduction of the overall cost.

7.4 Challenges of Microalgae and Thraustochytrids to Biofuel Although the production of biofuel from microalgae and thraustochytrids has a great potential, commercial-scale operation is still immature. There are still key obstacles in terms of environmental, technical, and economic viabilities to the commercialization of the technology. First of all, there are environmental concerns associated with the microbial cultivation and downstream processing in biofuel production, including greenhouse gas emission, water consumption, and wastewater discharge. It is a commonly accepted belief that biofuel production could offset the environmental concerns generated by fossil fuel production and consumption. However, it is still a question if microalgae biofuel production is eventually a carbon-neutral or net negative process. Although most of the cultivation process of microalgae indeed are process reducing carbon emissions, other operations involved in the microalgae-based biofuel production including construction and maintenance of the cultivation systems, downstream processes of biomass harvesting and biofuel conversion will certainly contribute to greenhouse gas emission. Due to different assumptions, the reported life cycle analyses of microalgae to biofuel processes have shown varied results with regard to greenhouse gas emission. Some reports even presented similar or higher emission in comparison with fossil fuels [91, 92]. Similarly, a recent life cycle assessment of thraustochytrids-based biodiesel production also indicated a comparable greenhouse gas emission to fossil diesel [86]. Other environmental concerns are the huge amount of water and nutrient consumptions and wastewater discharge. Although using wastewater for microalgae/thraustochytrids cultivation has been considered as an environmentally friendly win-win strategy, the nutrient profile of some wastewaters is not necessarily ideal for the microbial cell growth and lipid accumulation, which limits the commercial scale implementation of the technology. Besides, wastewaters generated during

7.5 Microalgae and Thraustochytrids for DHA and EPA Productions

oil extraction and transesterification/esterification processes need appropriate treatment. Secondly, several technical barriers must be overcome for the realization of industrial production of biofuel from microalgae/thraustochytrids. The productivity of microalgae- and thraustochytrids-based biofuel production with the current technology is still very low. Literature data in the last five years have shown that biomass productivity of microalgae in the laboratories is basically less than 0.3 g/(L day) [12, 14, 17, 22, 24], which still needs improvement. Enhanced biomass accumulation and lipid production could be achieved through technical renovations such as smart reactor design as well as strain and cultivation optimization. As revealed in a mechanistic model, the areal microalgal TAG productivity can be significantly improved by a factor of 5, reaching 10.9 g/(m2 day) with optimized technologies [93]. The biomass and lipid productivity of thraustochytrids can be one magnitude higher than that of microalgae. So far, the vast majority of both microalgae and thraustochytrids biofuel studies have focused on lab-scale investigation, leaving huge knowledge gaps related to commercial-scale processes [94]. It is imperative that every component of the biofuel production be tested for its operability and practicality stepwise from lab scale to pilot scale, and eventually to commercial scale. Thirdly and most importantly, current microalgae- or thraustochytrids-based biofuel production cannot compete economically against conventional fossil fuels. Techno-economic assessment of economic feasibility of microalgae biofuels indicated that the costs of establishment, operation, and maintenance were prohibitively too high [39]. According to the literature data, the total cost of microalgal oil production was estimated to be in the range $6.72–$31.61 per gallon, depending on the process assumptions [39, 91, 95–97]. Regardless of the great variation, the estimated cost is much higher than the current diesel price. Similarly, a comprehensive economic assessment is needed for thraustochytrids-based biofuel production. The economic feasibility of microalgae- or thraustochytrids-based large-scale biofuel production is believed to be dependent on the improvement in biomass harvesting and lipid extraction as well as compensations from co-products sales [98]. Finally, the environmental, technical, and economic challenges of biofuel production from microalgae and thraustochytrids should be considered simultaneously because improvement in one aspect can likely have effects on other aspects. For example, a closed photobioreactor (PBR) system has technical superiority with much higher biomass and lipid production as well as improved culture stability over an open raceway pond (ORP) system. Besides, PBR is environmentally advantageous with much less water usage than ORP. However, many economic assessments concluded that PBR requires higher capital and operational costs than ORP [91, 98].

7.5 Microalgae and Thraustochytrids for DHA and EPA Productions Both microalgae and thraustochytrids produce a variety of bioactive compounds, including astaxanthin, β-carotene, squalene, and ω-3 PUFAs, which have great

179

180

7 Utilization of Microalgae and Thraustochytrids

Table 7.3

Representatives of microalgae/thraustochytrids as EPA and DHA producers.

Microalgae/thraustochytrids species

EPAa) content (% TFAsb))

DHAc) content (% TFAs)

Reference

[104]

EPA producers Nannochloropsis oculata

35.51

—

Pavlova viridis

15.71

7.17

Phaeodactylum tricornutum

36.57

—

[105]

Phaeodactylum tricornutum

30.72

0.86

[106]

Phaeodactylum tricornutum

22.90

1.60

[103]

Trachydiscus minutus

51.10

—

[101]

DHA producers —

41.13

[107]

Thraustochytrium sp.

7.72

37.02

[108]

Thraustochytriidae sp.

—

36.93

[109]

Aurantiochytrium sp.

—

35.00

Thraustochytrium sp.

7.50

69.00

Aurantiochytrium sp.

0.50

46.70

Aurantiochytrium sp.

0.40

48.40

[84]

—

23.48

[110]

0.73

49.19

[111]

Crypthecodinium cohnii

Schizochytrium limacinum SR21 Schizochytrium sp.

[83]

a) EPA: eicosapentaenoic acid. b) TFAs: total fatty acids. c) DHA: docosahexaenoic acid.

antioxidative ability and potential applications as nutraceuticals [99, 100]. This section presents a review of recent studies on ω-3 PUFAs, especially DHA and EPA productions from microalgae and thraustochytrids. As the primary producers of DHA and EPA in the aquatic food chain, EPA and DHA fractions of the total fatty acids in microalgae cells generally range from 20 to over 45% [99]. Nannochloropsis, Dunaliella, Phaeodactylum tricornutum, Isochrysis galbana, Pavlova, and Crypthecodinium cohnii species are considered as good candidates for commercial production of EPA and DHA. While Crypthecodinium cohnii typically have a higher percentage of DHA than EPA, most of the other above-mentioned microalgae species contain less DHA [101–105]. Based on their distinct PUFA profiles, these microalgae species can be divided as EPA or DHA producers as shown in Table 7.3 [112]. Most thraustochytrid species contain relatively high DHA content, e.g. over 60% of the total fatty acids detected in some of the Thraustochytrium, Aurantiochytrium and Schizochytrium species [83, 113]. Specifically, Schizochytrium limacinum SR21 (also known Aurantiochytrium limacinum SR21) has been investigated extensively due to its great industrial application potentials for DHA production [110, 114–116].

7.5 Microalgae and Thraustochytrids for DHA and EPA Productions

Under optimized conditions, the DHA yield and productivity of Schizochytrium limacinum SR21 reached 32.36 g/L and 337.1 mg (L h), respectively [114]. Commercial DHA production from some Thraustochytrium and Schizochytrium species has been reported by Borowitzka [112] and Patil and Gogate [110]. Cultivation conditions have a significant impact on ω-3 PUFA accumulation in microalgae and thraustochytrids biomass. Physical and chemical parameters such as temperature, light intensity, pH, salinity, dissolved oxygen, and nutrients have been studied for maximizing EPA and DHA yields [101, 102, 109, 113]. Microalgae and thraustochytrids strains originated from cold regions tend to have higher EPA and DHA contents than those isolated from temperate and tropical areas, as the synthesis of unsaturated fatty acids is likely related to the maintenance of membrane fluidity [80, 117]. According to the reports, low temperature induced the increment of fatty acid unsaturation during the cultivation process in most cases [103, 109, 117]. For instance, the EPA content (% dry weight) of Nannochloropsis gaditana increased remarkably in the winter time as compared to that in the summer time in a tubular PBR [118]. In the study of ω-3 PUFA production from Phaeodactylum tricornutum, the decrease of temperature from 20 to 15 ∘ C caused significant increase of the DHA and EPA percentages of the total fatty acids [106]. However, due to the different impacts of temperature on biomass generation and total lipid accumulation, the decrease of cultivation temperature led to variable changes of EPA and/or DHA productivity. The EPA fraction of total fatty acids in microalga Trachydiscus minutus reached the highest level at 20 ∘ C, but the highest EPA productivity was observed at 28 ∘ C (the optimal temperature for the microalgal cell growth) due to a significant inhibition of T. minutus growth at 20 ∘ C [101]. The availability of carbon and nitrogen sources also has significant impact on lipid production and profile. In general, cellular lipid content will increase under nitrogen-limiting conditions or high C/N ratio, as the carbon source is diverted into lipid synthesis pathway. Meanwhile, the proportion of SFAs and MUFAs in TAG typically increases, while the percentage of PUFAs decreases with elevated C/N ratio [106]. As a result, EPA and DHA yield and productivity turn unpredictable with high C/N ratio in the media [87, 104]. Sufficient nitrogen supply (relatively low C/N ratio) is not only beneficial for cell formation, but also favorable for high fractions of EPA/DHA in total fatty acids. Consequently, the best C/N ratio for maximal EPA/DHA production must be determined for each individual strain [119]. Applicable nitrogen sources for PUFA production include inorganic chemicals such as ammonium and nitrate compounds, organic sources such as yeast extract and peptone, and complex substances such as steep liquors, which are cheap waste materials containing both carbon and nitrogen from industrial processes [110]. Efforts have also been made to explore alternative nitrogen sources to achieve economically feasible and efficient EPA/DHA production. It was found that equivalent EPA production was attained when urea was used as the nitrogen source as compared to that with potassium nitrate as the nitrogen source [101]. A variety of carbon sources including mono-, di- and poly-saccharides were also tested in microbial cultivation for DHA and EPA production, of which glycerol (especially crude glycerol from biodiesel production) was considered as a

181

182

7 Utilization of Microalgae and Thraustochytrids

Table 7.4 Heterotrophic microalgae/thraustochytrids DHAa) production with glycerol as the carbon source. Microalgae/ thraustochytrids strain

Carbon source

Cultivation mode

DHA content (% TFAsb))

DHA DHA yield productivity (g/L) (mg (L/h)) Reference

Glycerol: Initial Fed-batch concentration, 85 g/L; residual concentration, 5–20 g/L

41.75

28.93

301

[120]

Aurantiochytrium Glycerol: 40 g/L Fed-batch sp. TC 20

48.40

14.30

—

[84]

—

520

[121]

—

—

[108]

32.36

337

[114]

—

—

[115]

—

14.57

2910

[87]

—

10.12

—

23.48

5.73

—

Schizochytrium sp. S31

Schizochytrium limacinum

Crude glycerol: Continuous 31.09 90 g/L

Thraustochytrium Glycerol: sp. AMCQS5-5 0.5–10% (w/v)

Batch

21.65–37.02

Aurantiochytrium Glucose and limacinum SR21 glycerol mix: total carbon source concentration, 40–60 g/L

Fed-batch

43.71

Schizochytrium limacinum SR21

Crude glycerol: Batch 30 g/L

Aurantiochytrium Glycerol: sp. DBTIOC-18 120 g/L

Batch

Schizochytrium sp. DBTIOC-1 Schizochytrium limacinum SR21

Glucose: glycerol = 6 : 3

Batch

233.7 mg/g dry biomass

[110]

a) DHA: docosahexaenoic acid. b) TFAs: total fatty acids.

cost-effective substrate for the heterotrophic microalgae or thraustochytrids DHA production. Many recent studies have shown that glycerol is a suitable substitute for glucose [87, 108, 110, 120]. Glycerol can also be used as a supplement in the glucose medium to facilitate PUFA accumulation in microalgae cultivation [110]. Recent advances in microalgae/thraustochytrids DHA production using glycerol as the carbon source are summarized in Table 7.4. In addition to the cultivation conditions, cultivation modes have also been studied to improve DHA and EPA productions in microalgae and thraustochytrids cultures [120, 121]. Fed-batch process allows a constant increase of cell biomass and has been widely used in EPA/DHA production. In most cases, the carbon source is supplemented at certain points of the cultivation during a fed-batch process to improve cell growth rate and EPA/DHA accumulation [84, 111]. Besides,

7.6 Future Perspectives

the supplement of carbon source and nutrients may be able to meet the needs of microalgae and thraustochytrids for varied carbon/nutrient demands and specific EPA/DHA assimilation properties at different growth stages. Kim et al. [119] developed a novel fed-batch process for the cultivation of Aurantiochytrium sp. KRS101 in which yeast extract was fed to the microbes for one day and glucose fed for the remaining period based on the nutritional uptake patterns of the microbes. The novel process resulted in a maximum yield of 31 mg/g DHA, which was 82 and 41% higher than those obtained in only glucose- and yeast extract-fed processes, respectively. On the other hand, continuous cultivation also has its own advantage. Desired steady-state conditions can be maintained for EPA and DHA production in continuous mode with the adjustment of dilution rate of the system operation. Ethier et al. [121] revealed that in the continuous cultivation of Schizochytrium limacinum, stable fatty acid profile could be maintained for DHA production at the dilution rate of 0.3 d−1 , regardless of the different substrate concentrations in the feed. When used as nutraceutical supplements, DHA and EPA from neutral lipids, especially TAGs, are more readily absorbed by human bodies. Many studies have shown that TAGs are major components of total lipids in most microalgae and thraustochytrids species [84, 109, 122]. This indicates that microalgae and thraustochytrids have a great potential in the production of DHA/EPA as natural nutraceutical supplements.

7.6

Future Perspectives

7.6.1 Integrated Microalgae/Thraustochytrids Cultivation and Harvesting System Biomass harvesting from suspended cultures is one of the major bottlenecks for microalgae/thraustochytrids applications. In the two most commonly used microalgae cultivation systems, i.e. open raceway pond and closed photobioreactor, the microalgal cell concentration can be as low as 0.5 g/L and 2–6 g/L, respectively [98]. The thraustochytrids cell density in fermenters usually can be higher than that of microalgae, reaching over 50 g/L [100]. Still, huge amount of energy will be consumed to concentrate the biomass using conventional harvesting methods. The strategy of growing microalgae on the surface of specific supporting materials opens a new door to tackle the harvesting challenge associated with suspended cultivation. Through biomass scraping, microalgae harvesting can be easily integrated with cell cultivation in attached biofilm systems [123–125]. Moreover, the microalgal biofilm systems can likely achieve higher biomass productivity, compared to the suspended systems under the same conditions. Energy and water required for microalgae cultivation can also be significantly reduced in the biofilm systems [124, 126]. As revealed by Gross et al. [124], the specific water consumption rate in their microalgal biofilm system was only 7% of that in the control raceway pond. The initial application of the microalgal biofilm system was mainly for wastewater treatment so that

183

184

7 Utilization of Microalgae and Thraustochytrids

the algal biofilm system can be an integrated one for simultaneous microalgae production and nutrient removal from the wastewater [127, 128]. Although there has been increasing interest in microalgal biofilm systems in the last few years, the technology is still in its infant stage and very limited studies have been reported in the area. In terms of attached-growth cultivation of thraustochytrids, there are no relevant reports at present. Nevertheless, a number of thraustochytrids strains were found to produce extracellular polysaccharides, which may result in the formation of microbial aggregation, develop adhesive ability of microbial cultures, and consequently set a foundation for the establishment of thraustochytrids biofilms [4, 83]. For successful development of attached-growth microalgae/thraustochytrids systems, the following factors should be considered: 1) Selection of proper strains and supporting materials is of primary importance. Different strains may have diverse attachment abilities to a specific supporting material. On the other hand, the physicochemical properties of a supporting material, including roughness, surface free energy, liquid contact angle, and surface texture, will have great impacts on its interaction with the microorganisms. In addition, the durability of the supporting material will be important for the long-term performance of the biofilm system. 2) The structure of a biofilm system should be optimized for optimal nutrient transfer, water usage, and biomass accumulation for microalgae/thraustochytrids cultures as well as CO2 and light delivery for autotrophic microalgae. 3) Techno-economic assessment on the sustainability of a biofilm system is also important for its industrial applications.

7.6.2 Genetically Modified Microalgae/Thraustochytrids for High Oil and Easy Extraction of Lipids With the development of gene sequencing and regulation technologies, many efforts have been made to uncover the mystery of lipid accumulation in microalgae/thraustochytrids. Metabolic engineering of genes involved in the lipid biosynthesis pathways has emerged as an important alternative for enhanced production of oil from microalgae/thraustochytrids. Recently, the genomes of some microalgae and thraustochytrids strains have been sequenced and the release of the genome data provides great information for lipid regulations [129–133]. Indeed, the genomic analysis of Thraustochytrium sp. 26185 has revealed that 451 genes are involved in lipid transport and metabolism, and indicated the coexistence of two pathways for fatty acid biosynthesis [133]. These findings have not only shown new insight on the mechanism of fatty acid synthesis in the Thraustochytrium, but also provided guidance on genetic engineering strategies for fatty acid production. Genome-scale reconstruction and modeling were conducted to elucidate the lipid biosynthesis mechanism [134]. Furthermore, comparative transcriptomic and proteomic studies have contributed to the identification of key genes/enzymes and regulators involved in the production of lipids [135, 136]. The studies predicted that

7.6 Future Perspectives

abundant supply of acyl-CoA and NADPH would be essential for lipid synthesis [134, 137]. Yet, the genome data of most oleaginous microalgae/thraustochytrids strains still need to be unveiled for better understanding of lipid synthesis. In addition to gene annotation and statistical analysis at the omics level, further studies should be carried out to characterize the gene function for the understanding of lipid synthesis mechanisms and/or performing transgenic engineering for enhanced lipid synthesis [138, 139]. Through targeted gene modification, transgenic strains with increased lipid accumulation could be readily achieved [140, 141]. Acetyl-CoA synthases and carboxylases were identified as the key enzymes catalyzing the formation and conversion, respectively, of acetyl-CoA, which was the starting point of lipid biosynthesis. Hence, the overexpression of either or both of the enzymes might facilitate lipid accumulation. As observed by Yan et al. [142], increased biomass and fatty acid proportion was obtained in Schizochytrium transformants when the exogenous acetyl CoA synthase gene was incorporated and expressed in the microbe. The malic acid enzyme was revealed to promote the production of NADPH and pyruvate, which are vital for fatty acid accumulation [137, 143]. The overexpression of malic acid enzyme gene in transgenic Phaeodactylum tricornutum significantly improved the lipid content (2.5-fold), compared to the wild type [143]. In addition to the upregulation of lipid synthesis-related genes, the downregulation of the genes involved in competing pathways provides an alternative route for enhanced lipid accumulation. Particularly, the knockdown of a multifunctional lipase (lipid catabolism) resulted in enhanced lipid accumulation without compromising cell growth [141]. Two distinct pathways were identified for specific fatty acid accumulation in microalgae/thraustochytrids, i.e. fatty acid synthesis (FAS) pathway (also known as aerobic pathway) and polyketide synthase (PKS) pathway (also known as anaerobic pathway) [133, 144]. In FAS pathway, a series of desaturases and elongases are responsible for the introduction of double bond and carbon chain elongation. On the other hand, polyketide synthases are the key enzymes in the PKS pathway. As reported by Singh et al. [144], the regulation of specific genes involved in FAS or PKS pathway could result in enhanced production of target PUFAs in thraustochytrids. Likewise, the expression of Δ5-elongase in Phaeodactylum tricornutum also significantly increased the DHA content in the microorganism [103]. Although great achievements have been made in the genetic engineering of microalgae/thraustochytrids for enhanced lipid production, further research is still necessary to improve the production efficiency through a fundamental understanding of the genetic principles for cell growth and lipid synthesis. Strategies to control lipid synthesis in microalgae/thraustochytrids through metabolic engineering need to be developed and optimized for cost-effective lipid production. Another challenge in microalgae/thraustochytrids oil production is the high cost for oil extraction from the cells. A recent study revealed that the long-chain acyl-CoA synthetases were involved in fatty acid importation, and the knockdown of corresponding encoding genes resulted in extracellular fatty acids secretion [145]. Metabolic engineering targeting on easy extraction of oil from microalgae/ thraustochytrids cells would also be an important direction for future research.

185

186

7 Utilization of Microalgae and Thraustochytrids

7.6.3 Integrated Microalgae/Thraustochytrids System for Biofuel and DHA/EPA Production Although numerous efforts have been made in improving the efficiency and reducing the costs for biofuel production from microalgae/thraustochytrids in the last two decades, commercialization of the technology has not been realized at this point because of the prohibitively high cost of the process. Co-production of high-value byproducts, e.g. DHA or EPA, to offset the high cost of biofuel production may offer a solution for the sustainable commercial development of microalgae/thraustochytrids industrial process. Microalgae and thraustochytrids contain diverse biochemical components and are capable of producing various bioproducts, such as PUFAs, pigments, amino acids, starch, and polysaccharides. Biorefinery approach in which microalgae/thraustochytrids biomass is refined into various chemical products has been studied intensively in the last few years [146–151]. For example, a green processing platform was developed with sequential extractions of carotenoid, fucoxanthin, polar lipids, carbohydrates, and proteins from the microalga Isochrysis galbana [58]. An important area of microalgae/thraustochytrids biorefinery research has been the fractionation of different fatty acids. Most microalgae/thraustochytrids strains contain a complex fatty acid profile with a carbon chain length from 4 to 26 carbons and the degree of unsaturation in the range of 0–6 double bonds. In particular, considerable previous research data have indicated that a number of microalgae and thraustochytrids strains contain SFAs and MUFAs, which are suitable for biodiesel production, and simultaneously PUFAs, which are promising nutraceutical products. Dong et al. [152] found that zeolite HZSM-5 selectively catalyzed the esterification of short-chain fatty acids in microalgal biomass and enriched the remaining PUFAs in the original form. Hence, an integrated process was developed for efficient separation of short-chain fatty acids for biodiesel production and PUFAs for nutraceutical products. The co-production of PUFAs from microalgae/thraustochytrids, which have a much higher value than biodiesel, has a great potential to offset the cost for biodiesel production. Moreover, the quality of the biodiesel produced from the short-chain fatty acids was remarkably improved in terms of oxidative and thermal stability in the separation of PUFAs from the total lipids. Currently, most studies on the integrated production of biodiesel and PUFAs are in the conceptual level and focused on the development of individual processes in the biorefinery [153]. Further research and development in the integration of the processes are necessary for the optimization and eventually commercialization of the overall co-production technology.

References 1 Liu, T.Z., Wang, J.F., Hu, Q. et al. (2013). Attached cultivation technology of microalgae for efficient biomass feedstock production. Bioresource Technology 127: 216–222.

References

2 Katiyar, R., Gurjar, B., Biswas, S. et al. (2017). Microalgae: an emerging source of energy based bio-products and a solution for environmental issues. Renewable and Sustainable Energy Reviews 72 (C): 1083–1093. 3 Liu, Y., Singh, P., Liang, Y. et al. (2017). Abundance and molecular diversity of Thraustochytrids in coastal waters of Southern China. FEMS Microbiology Ecology 93 (6): fix070. 4 Liu, Y., Singh, P., Sun, Y. et al. (2014). Culturable diversity and biochemical features of thraustochytrids from coastal waters of Southern China. Applied Microbiology and Biotechnology 98 (7): 3241–3255. 5 Chen, B.L., Wan, C., Mehmood, M.A. et al. (2017). Manipulating environmental stresses and stress tolerance of microalgae for enhanced production of lipids and value-added products – a review. Bioresource Technology 244: 1198–1206. 6 Barros, A.I., Gonçalves, A.L., Simões, M., and Pires, J.C. (2015). Harvesting techniques applied to microalgae: a review. Renewable and Sustainable Energy Reviews 41: 1489–1500. 7 Kumar, S.J., Kumar, G.V., Dash, A. et al. (2017). Sustainable green solvents and techniques for lipid extraction from microalgae: a review. Algal Research 21: 138–147. 8 Yang, C., Li, R., Cui, C. et al. (2016). Catalytic hydroprocessing of microalgae-derived biofuels: a review. Green Chemistry 18 (13): 3684–3699. 9 Luo, Y., Le-Clech, P., and Henderson, R.K. (2017). Simultaneous microalgae cultivation and wastewater treatment in submerged membrane photobioreactors: a review. Algal Research 24: 425–437. 10 Mondal, M., Goswami, S., Ghosh, A. et al. (2017). Production of biodiesel from microalgae through biological carbon capture: a review. 3 Biotech 7 (2): 99. 11 El Arroussi, H., Benhima, R., El Mernissi, N. et al. (2017). Screening of marine microalgae strains from Moroccan coasts for biodiesel production. Renewable Energy 113: 1515–1522. 12 Nascimento, I.A., Marques, S.S.I., Cabanelas, I.T.D. et al. (2013). Screening microalgae strains for biodiesel production: lipid productivity and estimation of fuel quality based on fatty acids profiles as selective criteria. Bioenergy Research 6 (1): 1–13. 13 Mata, T.M., Martins, A.A., and Caetano, N.S. (2010). Microalgae for biodiesel production and other applications: a review. Renewable and Sustainable Energy Reviews 14 (1): 217–232. 14 Lam, M.K., Yusoff, M.I., Uemura, Y. et al. (2017). Cultivation of Chlorella vulgaris using nutrients source from domestic wastewater for biodiesel production: growth condition and kinetic studies. Renewable Energy 103: 197–207. 15 Pancha, I., Chokshi, K., Maurya, R. et al. (2015). Salinity induced oxidative stress enhanced biofuel production potential of microalgae Scenedesmus sp. CCNM 1077. Bioresource Technology 189: 341–348. 16 Brownbridge, G., Azadi, P., Smallbone, A. et al. (2014). The future viability of algae-derived biodiesel under economic and technical uncertainties. Bioresource Technology 151: 166–173.

187

188

7 Utilization of Microalgae and Thraustochytrids

17 Álvarez-Díaz, P., Ruiz, J., Arbib, Z. et al. (2017). Freshwater microalgae selection for simultaneous wastewater nutrient removal and lipid production. Algal Research 24 (SI): 477–485. 18 Thawechai, T., Cheirsilp, B., Louhasakul, Y. et al. (2016). Mitigation of carbon dioxide by oleaginous microalgae for lipids and pigments production: effect of light illumination and carbon dioxide feeding strategies. Bioresource Technology 219: 139–149. 19 Osundeko, O., Davies, H., and Pittman, J.K. (2013). Oxidative stress-tolerant microalgae strains are highly efficient for biofuel feedstock production on wastewater. Biomass and Bioenergy 56: 284–294. 20 Minhas, A.K., Hodgson, P., Barrow, C.J., and Adholeya, A. (2016). A review on the assessment of stress conditions for simultaneous production of microalgal lipids and carotenoids. Frontiers in Microbiology 7: 1–19. 21 Paliwal, C., Mitra, M., Bhayani, K. et al. (2017). Abiotic stresses as tools for metabolites in microalgae. Bioresource Technology 244 (Part 2): 1216–1226. 22 Daroch, M., Shao, C.C., Liu, Y. et al. (2013). Induction of lipids and resultant FAME profiles of microalgae from coastal waters of Pearl River Delta. Bioresource Technology 146: 192–199. 23 Chokshi, K., Pancha, I., Ghosh, A., and Mishra, S. (2017). Salinity induced oxidative stress alters the physiological responses and improves the biofuel potential of green microalgae Acutodesmus dimorphus. Bioresource Technology 244 (SI): 1376–1383. 24 Chokshi, K., Pancha, I., Trivedi, K. et al. (2015). Biofuel potential of the newly isolated microalgae Acutodesmus dimorphus under temperature induced oxidative stress conditions. Bioresource Technology 180: 162–171. 25 Converti, A., Casazza, A.A., Ortiz, E.Y. et al. (2009). Effect of temperature and nitrogen concentration on the growth and lipid content of Nannochloropsis oculata and Chlorella vulgaris for biodiesel production. Chemical Engineering and Processing: Process Intensification 48 (6): 1146–1151. 26 Arora, N., Patel, A., Pruthi, P.A., and Pruthi, V. (2016). Synergistic dynamics of nitrogen and phosphorous influences lipid productivity in Chlorella minutissima for biodiesel production. Bioresource Technology 213: 79–87. 27 Fu, L., Cui, X., Li, Y. et al. (2017). Excessive phosphorus enhances Chlorella regularis lipid production under nitrogen starvation stress during glucose heterotrophic cultivation. Chemical Engineering Journal 330: 566–572. 28 Wong, Y., Ho, Y., Ho, K. et al. (2017). Maximization of cell growth and lipid production of freshwater microalga Chlorella vulgaris by enrichment technique for biodiesel production. Environmental Science and Pollution Research 24 (10): 9089–9101. 29 Fan, J., Cui, Y., Wan, M. et al. (2014). Lipid accumulation and biosynthesis genes response of the oleaginous Chlorella pyrenoidosa under three nutrition stressors. Biotechnology for Biofuels 7 (1): 17. 30 Mohan, S.V. and Devi, M.P. (2014). Salinity stress induced lipid synthesis to harness biodiesel during dual mode cultivation of mixotrophic microalgae. Bioresource Technology 165: 288–294.

References

31 He, Q., Yang, H., Wu, L., and Hu, C. (2015). Effect of light intensity on physiological changes, carbon allocation and neutral lipid accumulation in oleaginous microalgae. Bioresource Technology 191: 219–228. 32 Atta, M., Idris, A., Bukhari, A., and Wahidin, S. (2013). Intensity of blue LED light: a potential stimulus for biomass and lipid content in fresh water microalgae Chlorella vulgaris. Bioresource Technology 148: 373–378. 33 Han, F., Pei, H., Hu, W. et al. (2016). Beneficial changes in biomass and lipid of microalgae Anabaena variabilis facing the ultrasonic stress environment. Bioresource Technology 209 (Supplement C): 16–22. 34 Pancha, I., Chokshi, K., George, B. et al. (2014). Nitrogen stress triggered biochemical and morphological changes in the microalgae Scenedesmus sp. CCNM 1077. Bioresource Technology 156: 146–154. 35 Ren, H.-Y., Liu, B.-F., Ma, C. et al. (2013). A new lipid-rich microalga Scenedesmus sp. strain R-16 isolated using Nile red staining: effects of carbon and nitrogen sources and initial pH on the biomass and lipid production. Biotechnology for Biofuels 6 (1): 143. 36 Mujtaba, G., Choi, W., Lee, C.-G., and Lee, K. (2012). Lipid production by Chlorella vulgaris after a shift from nutrient-rich to nitrogen starvation conditions. Bioresource Technology 123: 279–283. 37 Wang, T., Tian, X., Liu, T. et al. (2017). A two-stage fed-batch heterotrophic culture of Chlorella protothecoides that combined nitrogen depletion with hyperosmotic stress strategy enhanced lipid yield and productivity. Process Biochemistry 60: 74–83. 38 Chu, F.-F., Chu, P.-N., Shen, X.-F. et al. (2014). Effect of phosphorus on biodiesel production from Scenedesmus obliquus under nitrogen-deficiency stress. Bioresource Technology 152: 241–246. 39 Batan, L.Y., Graff, G.D., and Bradley, T.H. (2016). Techno-economic and Monte Carlo probabilistic analysis of microalgae biofuel production system. Bioresource Technology 219: 45–52. 40 Lee, J., Cho, D.-H., Ramanan, R. et al. (2013). Microalgae-associated bacteria play a key role in the flocculation of Chlorella vulgaris. Bioresource Technology 131: 195–201. 41 Powell, R.J. and Hill, R.T. (2014). Mechanism of algal aggregation by Bacillus sp. strain RP1137. Applied and Environmental Microbiology 80 (13): 4042–4050. 42 Hom-Diaz, A., Jaén-Gil, A., Bello-Laserna, I. et al. (2017). Performance of a microalgal photobioreactor treating toilet wastewater: pharmaceutically active compound removal and biomass harvesting. Science of the Total Environment 592: 1–11. 43 Zhou, W., Min, M., Hu, B. et al. (2013). Filamentous fungi assisted bio-flocculation: a novel alternative technique for harvesting heterotrophic and autotrophic microalgal cells. Separation and Purification Technology 107: 158–165. 44 Muradov, N., Taha, M., Miranda, A.F. et al. (2015). Fungal-assisted algal flocculation: application in wastewater treatment and biofuel production. Biotechnology for Biofuels 8 (1): 24.

189

190

7 Utilization of Microalgae and Thraustochytrids

45 Li, Y., Xu, Y., Liu, L. et al. (2016). First evidence of bioflocculant from Shinella albus with flocculation activity on harvesting of Chlorella vulgaris biomass. Bioresource Technology 218: 807–815. 46 Ndikubwimana, T., Zeng, X., Murwanashyaka, T. et al. (2016). Harvesting of freshwater microalgae with microbial bioflocculant: a pilot-scale study. Biotechnology for Biofuels 9 (1): 47. 47 Aljuboori, A.H.R., Uemura, Y., and Thanh, N.T. (2016). Flocculation and mechanism of self-flocculating lipid producer microalga Scenedesmus quadricauda for biomass harvesting. Biomass and Bioenergy 93: 38–42. 48 Lv, J., Guo, J., Feng, J. et al. (2016). A comparative study on flocculating ability and growth potential of two microalgae in simulated secondary effluent. Bioresource Technology 205: 111–117. 49 Tran, N.-A.T., Seymour, J.R., Siboni, N. et al. (2017). Photosynthetic carbon uptake induces autoflocculation of the marine microalga Nannochloropsis oculata. Algal Research 26: 302–311. 50 Salim, S., Kosterink, N.R., Tchetkoua Wacka, N.D. et al. (2014). Mechanism behind autoflocculation of unicellular green microalgae Ettlia texensis. Journal of Biotechnology 174: 34–38. 51 Ubeda, B., Galvez, J.A., Michel, M., and Bartual, A. (2017). Microalgae cultivation in urban wastewater: Coelastrum cf. pseudomicroporum as a novel carotenoid source and a potential microalgae harvesting tool. Bioresource Technology 228: 210–217. 52 Ho, S.-H., Chiu, S.-Y., Kao, C.-Y. et al. (2017). Ferrofluid-assisted rapid and directional harvesting of marine microalgal Chlorella sp. used for biodiesel production. Bioresource Technology 244 (SI): 1337–1340. 53 Lee, K., Lee, S.Y., Praveenkumar, R. et al. (2014). Repeated use of stable magnetic flocculant for efficient harvest of oleaginous Chlorella sp. Bioresource Technology 167: 284–290. 54 Seo, J.Y., Praveenkumar, R., Kim, B. et al. (2016). Downstream integration of microalgae harvesting and cell disruption by means of cationic surfactant-decorated Fe3 O4 nanoparticles. Green Chemistry 18 (14): 3981–3989. 55 Wang, T., Yang, W.-L., Hong, Y., and Hou, Y.-L. (2016). Magnetic nanoparticles grafted with amino-riched dendrimer as magnetic flocculant for efficient harvesting of oleaginous microalgae. Chemical Engineering Journal 297: 304–314. 56 Chen, L., Li, R., Ren, X., and Liu, T. (2016). Improved aqueous extraction of microalgal lipid by combined enzymatic and thermal lysis from wet biomass of Nannochloropsis oceanica. Bioresource Technology 214: 138–143. 57 Lee, I. and Han, J.-I. (2015). Simultaneous treatment (cell disruption and lipid extraction) of wet microalgae using hydrodynamic cavitation for enhancing the lipid yield. Bioresource Technology 186: 246–251. 58 Gilbert-Lopez, B., Mendiola, J.A., Fontecha, J. et al. (2015). Downstream processing of Isochrysis galbana: a step towards microalgal biorefinery. Green Chemistry 17 (9): 4599–4609. 59 McKennedy, J., Önenç, S., Pala, M., and Maguire, J. (2016). Supercritical carbon dioxide treatment of the microalgae Nannochloropsis oculata for the

References

60

61

62

63

64

65

66

67 68

69

70

71

72

73

production of fatty acid methyl esters. The Journal of Supercritical Fluids 116 (Supplement C): 264–270. Taher, H., Al-Zuhair, S., Al-Marzouqi, A.H. et al. (2014). Supercritical carbon dioxide extraction of microalgae lipid: Process optimization and laboratory scale-up. The Journal of Supercritical Fluids 86: 57–66. Taher, H., Al-Zuhair, S., Al-Marzouqi, A.H. et al. (2014). Effective extraction of microalgae lipids from wet biomass for biodiesel production. Biomass and Bioenergy 66: 159–167. Garoma, T. and Janda, D. (2016). Investigation of the effects of microalgal cell concentration and electroporation, microwave and ultrasonication on lipid extraction efficiency. Renewable Energy 86: 117–123. Chen, C.-L., Huang, C.-C., Ho, K.-C. et al. (2015). Biodiesel production from wet microalgae feedstock using sequential wet extraction/transesterification and direct transesterification processes. Bioresource Technology 194: 179–186. Yoo, G., Yoo, Y., Kwon, J.-H. et al. (2014). An effective, cost-efficient extraction method of biomass from wet microalgae with a functional polymeric membrane. Green Chemistry 16 (1): 312–319. Guo, H., Chen, H., Fan, L. et al. (2017). Enzymes produced by biomass-degrading bacteria can efficiently hydrolyze algal cell walls and facilitate lipid extraction. Renewable Energy 109: 195–201. Lai, Y.S., De Francesco, F., Aguinaga, A. et al. (2016). Improving lipid recovery from Scenedesmus wet biomass by surfactant-assisted disruption. Green Chemistry 18 (5): 1319–1326. Wu, C., Xiao, Y., Lin, W. et al. (2017). Surfactants assist in lipid extraction from wet Nannochloropsis sp. Bioresource Technology 243: 793–799. Huang, W.-C. and Kim, J.-D. (2017). Simultaneous cell disruption and lipid extraction in a microalgal biomass using a nonpolar tertiary amine. Bioresource Technology 232: 142–145. Guldhe, A., Singh, P., Kumari, S. et al. (2016). Biodiesel synthesis from microalgae using immobilized Aspergillus niger whole cell lipase biocatalyst. Renewable Energy 85: 1002–1010. Carrero, A., Vicente, G., Rodriguez, R. et al. (2015). Synthesis of fatty acids methyl esters (FAMEs) from Nannochloropsis gaditana microalga using heterogeneous acid catalysts. Biochemical Engineering Journal 97: 119–124. Veillette, M., Giroir-Fendler, A., Faucheux, N., and Heitz, M. (2017). Esterification of free fatty acids with methanol to biodiesel using heterogeneous catalysts: from model acid oil to microalgae lipids. Chemical Engineering Journal 308: 101–109. Guldhe, A., Singh, P., Ansari, F.A. et al. (2017). Biodiesel synthesis from microalgal lipids using tungstated zirconia as a heterogeneous acid catalyst and its comparison with homogeneous acid and enzyme catalysts. Fuel 187: 180–188. Loong, T.C. and Idris, A. (2017). One step transesterification of biodiesel production using simultaneous cooling and microwave heating. Journal of Cleaner Production 146: 57–62.

191

192

7 Utilization of Microalgae and Thraustochytrids

74 Park, J.Y., Park, M.S., Lee, Y.C., and Yang, J.W. (2015). Advances in direct transesterification of algal oils from wet biomass. Bioresource Technology 184: 267–275. 75 Bwapwa, J.K., Anandraj, A., and Trois, C. (2017). Possibilities for conversion of microalgae oil into aviation fuel: a review. Renewable and Sustainable Energy Reviews 80: 1345–1354. 76 Wang, H.Y., Bluck, D., and Van Wie, B.J. (2014). Conversion of microalgae to jet fuel: process design and simulation. Bioresource Technology 167: 349–357. 77 Yang, X.Y., Guo, F., Xue, S., and Wang, X. (2016). Carbon distribution of algae-based alternative aviation fuel obtained by different pathways. Renewable and Sustainable Energy Reviews 54: 1129–1147. 78 Fortier, M.O.P., Roberts, G.W., Stagg-Williams, S.M., and Sturm, B.S.M. (2014). Life cycle assessment of bio-jet fuel from hydrothermal liquefaction of microalgae. Applied Energy 122: 73–82. 79 Byreddy, A.R., Gupta, A., Barrow, C.J., and Puri, M. (2015). Comparison of cell disruption methods for improving lipid extraction from thraustochytrid strains. Marine Drugs 13 (8): 5111–5127. 80 Gupta, A., Singh, D., Byreddy, A.R. et al. (2016). Exploring omega-3 fatty acids, enzymes and biodiesel producing thraustochytrids from Australian and Indian marine biodiversity. Biotechnology Journal 11 (3): 345–355. 81 Chang, K.J.L., Dunstan, G.A., Abell, G.C.J. et al. (2012). Biodiscovery of new Australian thraustochytrids for production of biodiesel and long-chain omega-3 oils. Applied Microbiology and Biotechnology 93 (5): 2215–2231. 82 Chen, W., Ma, L., Zhou, P.-p. et al. (2015). A novel feedstock for biodiesel production: the application of palmitic acid from Schizochytrium. Energy 86: 128–138. 83 Chang, K.J.L., Nichols, C.M., Blackburn, S.I. et al. (2014). Comparison of thraustochytrids Aurantiochytrium sp., Schizochytrium sp., Thraustochytrium sp., and Ulkenia sp. for production of biodiesel, long-chain omega-3 oils, and exopolysaccharide. Marine Biotechnology 16 (4): 396–411. 84 Chang, K.J.L., Dumsday, G., Nichols, P.D. et al. (2013). High cell density cultivation of a novel Aurantiochytrium sp. strain TC 20 in a fed-batch system using glycerol to produce feedstock for biodiesel and omega-3 oils. Applied Microbiology and Biotechnology 97 (15): 6907–6918. 85 Thyagarajan, T., Puri, M., Vongsvivut, J., and Barrow, C.J. (2014). Evaluation of bread crumbs as a potential carbon source for the growth of Thraustochytrid species for oil and omega-3 production. Nutrients 6 (5): 2104–2114. 86 Chang, K.J.L., Rye, L., Dunstan, G.A. et al. (2015). Life cycle assessment: heterotrophic cultivation of thraustochytrids for biodiesel production. Journal of Applied Phycology 27 (2): 639–647. 87 Singh, D., Barrow, C.J., Puri, M. et al. (2016). Combination of calcium and magnesium ions prevents substrate inhibition and promotes biomass and lipid production in thraustochytrids under higher glycerol concentration. Algal Research 15: 202–209.

References

88 Lowrey, J., Armenta, R.E., and Brooks, M.S. (2016). Recycling of lipid-extracted hydrolysate as nitrogen supplementation for production of thraustochytrid biomass. Journal of Industrial Microbiology & Biotechnology 43 (8): 1105–1115. 89 Lowrey, J., Armenta, R.E., and Brooks, M.S. (2016). Sequential recycling of enzymatic lipid-extracted hydrolysate in fermentations with a thraustochytrid. Bioresource Technology 209: 333–342. 90 Lowrey, J., Brooks, M.S., and Armenta, R.E. (2016). Nutrient recycling of lipid-extracted waste in the production of an oleaginous thraustochytrid. Applied Microbiology and Biotechnology 100 (10): 4711–4721. 91 Quinn, J.C. and Davis, R. (2015). The potentials and challenges of algae based biofuels: a review of the techno-economic, life cycle, and resource assessment modeling. Bioresource Technology 184: 444–452. 92 Weinberg, J., Kaltschmitt, M., and Wilhelm, C. (2012). Analysis of greenhouse gas emissions from microalgae-based biofuels. Biomass Conversion and Biorefinery 2 (2): 179–194. 93 Breuer, G., Lamers, P.P., Janssen, M. et al. (2015). Opportunities to improve the areal oil productivity of microalgae. Bioresource Technology 186: 294–302. 94 Zhang, T.-Y., Hu, H.-Y., Wu, Y.-H. et al. (2016). Promising solutions to solve the bottlenecks in the large-scale cultivation of microalgae for biomass/bioenergy production. Renewable and Sustainable Energy Reviews 60: 1602–1614. 95 Delrue, F., Setier, P.-A., Sahut, C. et al. (2012). An economic, sustainability, and energetic model of biodiesel production from microalgae. Bioresource Technology 111: 191–200. 96 Hoffman, J., Pate, R.C., Drennen, T., and Quinn, J.C. (2017). Techno-economic assessment of open microalgae production systems. Algal Research 23: 51–57. 97 Richardson, J.W., Johnson, M.D., and Outlaw, J.L. (2012). Economic comparison of open pond raceways to photo bio-reactors for profitable production of algae for transportation fuels in the Southwest. Algal Research 1 (1): 93–100. 98 Davis, R., Aden, A., and Pienkos, P.T. (2011). Techno-economic analysis of autotrophic microalgae for fuel production. Applied Energy 88 (10): 3524–3531. 99 Adarme-Vega, T.C., Lim, D.K., Timmins, M. et al. (2012). Microalgal biofactories: a promising approach towards sustainable omega-3 fatty acid production. Microbial Cell Factories 11 (1): 96. 100 Marchan, L.F., Chang, K.J.L., Nichols, P.D. et al. (2017). Taxonomy, ecology and biotechnological applications of thraustochytrids: a review. Biotechnology Advances 36 (1): 26–46. 101 Cepák, V., Pˇribyl, P., Kohoutková, J., and Kaštánek, P. (2014). Optimization of cultivation conditions for fatty acid composition and EPA production in the eustigmatophycean microalga Trachydiscus minutus. Journal of Applied Phycology 26 (1): 181–190. 102 Chen, C.-Y., Chen, Y.-C., Huang, H.-C. et al. (2013). Engineering strategies for enhancing the production of eicosapentaenoic acid (EPA) from an isolated microalga Nannochloropsis oceanica CY2. Bioresource Technology 147 (Supplement C): 160–167.

193

194

7 Utilization of Microalgae and Thraustochytrids

103 Hamilton, M.L., Haslam, R.P., Napier, J.A., and Sayanova, O. (2014). Metabolic engineering of Phaeodactylum tricornutum for the enhanced accumulation of omega-3 long chain polyunsaturated fatty acids. Metabolic Engineering 22: 3–9. 104 Huang, X., Huang, Z., Wen, W., and Yan, J. (2013). Effects of nitrogen supplementation of the culture medium on the growth, total lipid content and fatty acid profiles of three microalgae (Tetraselmis subcordiformis, Nannochloropsis oculata and Pavlova viridis). Journal of Applied Phycology 25 (1): 129–137. 105 Rodolfi, L., Biondi, N., Guccione, A. et al. (2017). Oil and eicosapentaenoic acid production by the diatom Phaeodactylum tricornutum cultivated outdoors in Green Wall Panel (GWP®) reactors. Biotechnology and Bioengineering 114 (10): 2204–2210. 106 Qiao, H., Cong, C., Sun, C. et al. (2016). Effect of culture conditions on growth, fatty acid composition and DHA/EPA ratio of Phaeodactylum tricornutum. Aquaculture 452 (Supplement C): 311–317. 107 Liu, B., Liu, J., Sun, P. et al. (2015). Sesamol enhances cell growth and the biosynthesis and accumulation of docosahexaenoic acid in the microalga Crypthecodinium cohnii. Journal of Agricultural and Food Chemistry 63 (23): 5640–5645. 108 Gupta, A., Singh, D., Barrow, C.J., and Puri, M. (2013). Exploring potential use of Australian thraustochytrids for the bioconversion of glycerol to omega-3 and carotenoids production. Biochemical Engineering Journal 78: 11–17. 109 Liu, Y., Tang, J., Li, J. et al. (2014). Efficient production of triacylglycerols rich in docosahexaenoic acid (DHA) by osmo-heterotrophic marine protists. Applied Microbiology and Biotechnology 98 (23): 9643–9652. 110 Patil, K.P. and Gogate, P.R. (2015). Improved synthesis of docosahexaenoic acid (DHA) using Schizochytrium limacinum SR21 and sustainable media. Chemical Engineering Journal 268: 187–196. 111 Qu, L., Ren, L.-J., Sun, G.-N. et al. (2013). Batch, fed-batch and repeated fed-batch fermentation processes of the marine thraustochytrid Schizochytrium sp. for producing docosahexaenoic acid. Bioprocess and Biosystems Engineering 36 (12): 1905–1912. 112 Borowitzka, M.A. (2013). High-value products from microalgae – their development and commercialisation. Journal of Applied Phycology 25 (3): 743–756. 113 Gupta, A., Barrow, C.J., and Puri, M. (2012). Omega-3 biotechnology: thraustochytrids as a novel source of omega-3 oils. Biotechnology Advances 30 (6): 1733–1745. 114 Li, J., Liu, R., Chang, G. et al. (2015). A strategy for the highly efficient production of docosahexaenoic acid by Aurantiochytrium limacinum SR21 using glucose and glycerol as the mixed carbon sources. Bioresource Technology 177 (Supplement C): 51–57. 115 Lung, Y.-T., Tan, C.H., Show, P.L. et al. (2016). Docosahexaenoic acid production from crude glycerol by Schizochytrium limacinum SR21. Clean Technologies and Environmental Policy 18 (7): 2209–2216. 116 Yokoyama, R. and Honda, D. (2007). Taxonomic rearrangement of the genus Schizochytrium sensu lato based on morphology, chemotaxonomic

References

117

118

119

120

121

122

123

124

125

126

127

128

characteristics, and 18S rRNA gene phylogeny (Thraustochytriaceae, Labyrinthulomycetes): emendation for Schizochytrium and erection of Aurantiochytrium and Oblongichytrium gen. nov. Mycoscience 48 (4): 199–211. Boelen, P., van Dijk, R., Sinninghe Damsté, J.S. et al. (2013). On the potential application of polar and temperate marine microalgae for EPA and DHA production. AMB Express 3 (1): 26. Camacho-Rodríguez, J., González-Céspedes, A.M., Cerón-García, M.C. et al. (2014). A quantitative study of eicosapentaenoic acid (EPA) production by Nannochloropsis gaditana for aquaculture as a function of dilution rate, temperature and average irradiance. Applied Microbiology and Biotechnology 98 (6): 2429–2440. Kim, K., Jung Kim, E., Ryu, B.-G. et al. (2013). A novel fed-batch process based on the biology of Aurantiochytrium sp. KRS101 for the production of biodiesel and docosahexaenoic acid. Bioresource Technology 135 (Supplement C): 269–274. Chang, G., Gao, N., Tian, G. et al. (2013). Improvement of docosahexaenoic acid production on glycerol by Schizochytrium sp. S31 with constantly high oxygen transfer coefficient. Bioresource Technology 142 (Supplement C): 400–406. Ethier, S., Woisard, K., Vaughan, D., and Wen, Z. (2011). Continuous culture of the microalgae Schizochytrium limacinum on biodiesel-derived crude glycerol for producing docosahexaenoic acid. Bioresource Technology 102 (1): 88–93. Yang, Y.-H., Du, L., Hosokawa, M. et al. (2017). Fatty acid and lipid class composition of the microalga Phaeodactylum tricornutum. Journal of Oleo Science 66 (4): 363–368. Gross, M., Henry, W., Michael, C., and Wen, Z. (2013). Development of a rotating algal biofilm growth system for attached microalgae growth with in situ biomass harvest. Bioresource Technology 150: 195–201. Gross, M., Mascarenhas, V., and Wen, Z. (2015). Evaluating algal growth performance and water use efficiency of pilot-scale revolving algal biofilm (RAB) culture systems. Biotechnology and Bioengineering 112 (10): 2040–2050. Lee, S.H., Oh, H.M., Jo, B.H. et al. (2014). Higher biomass productivity of microalgae in an attached growth system, using wastewater. Journal of Microbiology and Biotechnology 24 (11): 1566–1573. Ozkan, A., Kinney, K., Katz, L., and Berberoglu, H. (2012). Reduction of water and energy requirement of algae cultivation using an algae biofilm photobioreactor. Bioresource Technology 114: 542–548. Gao, F., Yang, Z.-H., Li, C. et al. (2015). A novel algal biofilm membrane photobioreactor for attached microalgae growth and nutrients removal from secondary effluent. Bioresource Technology 179: 8–12. Hoh, D., Watson, S., and Kan, E. (2016). Algal biofilm reactors for integrated wastewater treatment and biofuel production: a review. Chemical Engineering Journal 287: 466–473.

195

196

7 Utilization of Microalgae and Thraustochytrids

129 Carreres, B., de Jaeger, L., Springer, J. et al. (2017). Draft genome sequence of the oleaginous green alga Tetradesmus obliquus UTEX 393. Genome Announcements 5 (3): e01449-16. 130 Ji, X.-J., Mo, K.-Q., Ren, L.-J. et al. (2015). Genome sequence of Schizochytrium sp. CCTCC M209059, an effective producer of docosahexaenoic acid-rich lipids. Genome Announcements 3 (4): e00819-15. 131 Liu, B., Ertesvåg, H., Aasen, I.M. et al. (2016). Draft genome sequence of the docosahexaenoic acid producing thraustochytrid Aurantiochytrium sp. T66. Genomics Data 8 (Supplement C): 115–116. 132 Starkenburg, S., Polle, J., Hovde, B. et al. (2017). Draft nuclear genome, complete chloroplast genome, and complete mitochondrial genome for the biofuel/bioproduct feedstock species Scenedesmus obliquus strain DOE0152z. Genome Announcements 5 (32): e00617-17. 133 Zhao, X., Dauenpen, M., Qu, C., and Qiu, X. (2016). Genomic analysis of genes involved in the biosynthesis of very long chain polyunsaturated fatty acids in Thraustochytrium sp. 26185. Lipids 51 (9): 1065–1075. 134 Ye, C., Qiao, W., Yu, X. et al. (2015). Reconstruction and analysis of the genome-scale metabolic model of schizochytrium limacinum SR21 for docosahexaenoic acid production. BMC Genomics 16 (1): 799. 135 Gao, C., Wang, Y., Shen, Y. et al. (2014). Oil accumulation mechanisms of the oleaginous microalga Chlorella protothecoides revealed through its genome, transcriptomes, and proteomes. BMC Genomics 15 (1): 582. 136 Li, J., Han, D., Wang, D. et al. (2014). Choreography of transcriptomes and lipidomes of Nannochloropsis reveals the mechanisms of oil synthesis in microalgae. The Plant Cell 26 (4): 1645–1665. 137 Ren, L., Hu, X., Zhao, X. et al. (2017). Transcriptomic analysis of the regulation of lipid fraction migration and fatty acid biosynthesis in Schizochytrium sp. Scientific Reports 7: 3562. 138 Courchesne, N.M.D., Parisien, A., Wang, B., and Lan, C.Q. (2009). Enhancement of lipid production using biochemical, genetic and transcription factor engineering approaches. Journal of Biotechnology 141 (1): 31–41. 139 De Bhowmick, G., Koduru, L., and Sen, R. (2015). Metabolic pathway engineering towards enhancing microalgal lipid biosynthesis for biofuel application – a review. Renewable and Sustainable Energy Reviews 50: 1239–1253. 140 Daboussi, F., Leduc, S., Maréchal, A. et al. (2014). Genome engineering empowers the diatom Phaeodactylum tricornutum for biotechnology. Nature Communications 5: 3831. 141 Trentacoste, E.M., Shrestha, R.P., Smith, S.R. et al. (2013). Metabolic engineering of lipid catabolism increases microalgal lipid accumulation without compromising growth. Proceedings of the National Academy of Sciences 110 (49): 19748–19753. 142 Yan, J., Cheng, R., Lin, X. et al. (2013). Overexpression of acetyl-CoA synthetase increased the biomass and fatty acid proportion in microalga Schizochytrium. Applied Microbiology and Biotechnology 97 (5): 1933–1939.

References

143 Xue, J., Niu, Y.-F., Huang, T. et al. (2015). Genetic improvement of the microalga Phaeodactylum tricornutum for boosting neutral lipid accumulation. Metabolic Engineering 27 (Supplement C): 1–9. 144 Singh, P., Liu, Y., Li, L., and Wang, G. (2014). Ecological dynamics and biotechnological implications of thraustochytrids from marine habitats. Applied Microbiology and Biotechnology 98 (13): 5789–5805. 145 Jia, B., Song, Y., Wu, M. et al. (2016). Characterization of long-chain acyl-CoA synthetases which stimulate secretion of fatty acids in green algae Chlamydomonas reinhardtii. Biotechnology for Biofuels 9 (1): 184. 146 da Silva, T.L., Gouveia, L., and Reis, A. (2014). Integrated microbial processes for biofuels and high value-added products: the way to improve the cost effectiveness of biofuel production. Applied Microbiology and Biotechnology 98 (3): 1043–1053. 147 Foley, P.M., Beach, E.S., and Zimmerman, J.B. (2011). Algae as a source of renewable chemicals: opportunities and challenges. Green Chemistry 13 (6): 1399–1405. 148 Li, J., Liu, Y., Cheng, J.J. et al. (2015). Biological potential of microalgae in China for biorefinery-based production of biofuels and high value compounds. New Biotechnology 32 (6): 588–596. 149 Trivedi, J., Aila, M., Bangwal, D.P. et al. (2015). Algae based biorefinery – how to make sense? Renewable and Sustainable Energy Reviews 47 (Supplement C): 295–307. 150 Yen, H.-W., Hu, I.C., Chen, C.-Y. et al. (2013). Microalgae-based biorefinery – from biofuels to natural products. Bioresource Technology 135 (Supplement C): 166–174. 151 Zhu, L. (2015). Biorefinery as a promising approach to promote microalgae industry: an innovative framework. Renewable and Sustainable Energy Reviews 41 (Supplement C): 1376–1384. 152 Dong, T., Yu, X., Miao, C. et al. (2015). Selective esterification to produce microalgal biodiesel and enrich polyunsaturated fatty acid using zeolite as a catalyst. RSC Advances 5 (103): 84894–84900. 153 Vermuë, M., Eppink, M., Wijffels, R., and van den Berg, C. (2017). Multi-product microalgae biorefineries: from concept towards reality. Trends in Biotechnology 36 (2): 216–227.

197

199

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective Goldy De Bhowmick and Ajit K. Sarmah Department of Civil and Environmental Engineering, The Faculty of Engineering, The University of Auckland, Private Bag 92019, Auckland 1142, New Zealand

8.1 Introduction Over the years, the first-generation bioenergy strategy has mainly focused on starch, sugar, vegetable or animal oil-based feedstock using conventional technologies for energy generation [1]. These methods received tremendous global criticism due to their direct competition with consumable food resources. However, to circumvent these issues, the second-generation strategy was adopted using nonedible or waste vegetable oils and terrestrial agricultural wastes such as straw leaves and lumber but was incompatible due to their low availability and issues related with indirect emission and carbon debt from land clearance [1–3]. Hence, a more sustainable feedstock is pondered upon. In this regard, microalgae emerged as one of the most promising renewable and alternative energy feedstock in response to energy crisis, global warming, and climate change [3, 4]. The advantages of using microalgae as a favorable feedstock for bioenergy generation [5–7] include: ●

● ● ● ● ●

●

High percentage of oil yield (60 times higher than soybean, 15 times more productive than Jatropha, and 5 times more than palm oil per acre of land on an annual basis). High photosynthetic ability. High biomass productivity and yield per acre of cultivation. Do not compete for arable land. Nonfood based resource. Do not require fresh water and are capable of growing in brackish, saline, and wastewaters. Potential recycling of carbon from CO2 -rich flue gas emissions from stationary sources, including power plants and other industrial emitters.

However, economically viable production of conventional bioenergies from algae is still not realistic due to horrendous capital and operational investments [8]. A recent techno-economic analysis demonstrated that reducing algae harvesting/ Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

200

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

cultivation cost to make biofuel economy viable is exceedingly difficult [9]. Therefore, multitasking that aims at maximizing value-derived products from different microalgal components is expected to promote the algal biofuel economics. Furthermore, the highly diverse and unique and complex chemical structure of algal biomass make it a usable form of bio-resource to produce a cascade of high-value low-volume products along with biofuels within the biorefinery contexts [10]. However, challenges lie in identifying issues at the interface between production and conversion processes, discovering novel compounds and establishing a link with scaled conversion process characteristics and respective market opportunities for different bio-products [6, 7, 9]. Thus this chapter focuses primarily on pertinent issues of algal energy and bio-product formation in order to establish a successful biorefinery. Although numerous published research studies are available on algal biomass-based biorefineries, report on identifying issues at the interface level between production and conversion processes while establishing a link with scaled conversion process characteristics and respective market opportunities for different bio-products is limited. The goal of this chapter is to shed light on the various aspects of algal bio-product formation along with fuel generation in a biorefinery concept. This study would help in understanding the global energy requirement and availability to meet the demand along with the development of other prominent value-added products. Identifying factors that influence value-added product formation is important for improving their quality and quantity.

8.2 Current Status of Algal Energy and Bio-product Formation In recent years, among the various renewable energy sources, biomass-based bioenergy production alone accounted for 10% of the total primary energy supply (TPES) followed by 2.3% from hydro energy sources [4, 11]. Interestingly, in comparison to other forms of renewable energy, biofuels derived from algae have been found to store the energy chemically and also to be used in engines and transportation infrastructure after blending them with petroleum diesel to various degrees [11]. Based on these evidences, biofuels derived from algae will be able to diversify the global energy source, while gradually replacing the fossil fuel [2, 11]. As of now, microalgae have been pursued as a superior raw material for biodiesel production alone all over the world and are considered as an alternative fuel in diesel engine [1, 6, 11]. Evidences indicate that fuel properties of pure microalgae biodiesel and its blend with diesel fuel humbly satisfy the European Biodiesel Standards (EN 14214). However, the cetane number of microalgal biodiesel was found to be low; this can be compensated by mixing microalgae biodiesel with diesel fuel [11]. Importantly, algal biofuel has been found to play a significant role in reducing the global emissions of GHG while improving the global transportation fuel [6]. According to recent estimation, renewable energy is expected to be the major global energy generator by 2070. For instance, the European Energy Commission

8.2 Current Status of Algal Energy and Bio-product Formation

has set a target for aviation industries to use 2 million tonnes of biofuel by 2020 with a reduction of 3% of GHG emission in Europe [12]. If the plan is successful then there will be a drastic reduction in the dependence of fossil fuel usage with a 60% reduction of GHG by 2050. In order to fulfill the target, aviation industries have already implemented the use of special algae-based hydrocarbons as additives. At the moment, a “drop in” replacement program is being highlighted to replace the jet fuels. One such example is the production of first ever algae-based jet fuel by Solazyme (a California-based renewable oil producer) that passed all the tests including the most challenging ASTM D1655 specification for aviation turbine fuel [12]. The major oil-producing countries have set specific future goals and targets for using biofuel and renewable forms of energy rather than fossil fuels. For instance, countries like UAE and USA have proposed to turn down 10 and 20% of their transportation fuel by using biofuel by 2020 and 2022, respectively [12, 13]. Additionally, the renewable transport fuel obligation statistics of United Kingdom (department of transportation) ensures that if the consistency of renewable fuel proportional quality is maintained, then a certificate called Renewable Transport Fuel Certificates (RTFC) will be awarded to the suppliers [12, 13]. Furthermore, the Renewable Energy Directive (RED) has announced the promotion of usage of advanced algal biofuels by 2020. In order to facilitate more markets for biofuel producers, a minimum of 15p/L for RTFCs was implemented [12, 13]. Such advancements in using algae based biofuel seconds the ideology of replacing fossil fuel that promises to preserve our environment while maintaining a sustainable energy sector. A rough estimation of the current worldwide commercial production of microalgae biomass from different species is as follows: roughly 10 000 tons/year and $10 000/ton for Spirulina; 4000 tons/year and $20 000/ton for Chlorella; 1000 tons/year and $20 000/ton for Dunaliella; and 200 tons/year and $100 000/ton for Haematococcus [14]. At a smaller scale, Tetraselmis and Nannochloropsis are cultivated commercially but have not reached the target of several hundreds of tons yet [14]. The raceway pond cultivation system offers half the initial investments than closed bioreactor systems at the expense of lower productivities, dilute cultures with more volume to process, contamination problem and heavy rain that may interfere with proper operation[15]. Though closed systems need more investment, they offer several advantages such as high biomass productivity; more controlled environment; and more degrees of freedom in designing, construction, operation, and implication. However, large-scale cultivation in closed system involves considerable constraints. For instance, overheating, oxygen buildup, and biofouling are a big problem as they can be lethal to microalgae [15]. Currently, the major commercial products are based on harvesting and drying of biomass or on extracting and purifying special lipids such as omega-3 fatty acids or pigments such as astaxanthin. At the moment, the only pigments on market are β-carotene from Dunaliella salina and astaxanthin from Haematococcus pluvialis [14, 15]. Another pigment, phycocyanin from Arthrospira is currently commercially used as a food colorant and in cosmetics in China, Japan, and Thailand [14]. However, no report has been found yet on the applicability of the remaining

201

202

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

biomass that is still rich in lipids (e.g., γ-linolenic acid), proteins (membrane proteins), and carbohydrates.

8.3 Analysis of Conversion Methods Biomass cultivation and harvesting cost can be further reduced by considering complete valorization of the algal cell components rather than relying on lipid fractions only. Using or following efficient simultaneous extraction and conversion methods of almost all algal components into fuels and value-added chemicals would aid in reducing the biofuel production cost, while lowering the risks to stakeholders [16]. Discovering novel promising platform chemicals or bio-products would provide a rational means for upgrading detailed biomass production cost and biomass composition. Specific understanding of biomass composition would provide a framework to identify critical compositional characteristics with suitable conversion techniques [9]. At the moment, the possibility of generating these bio-products is at the hypothetical stage. Therefore, we would like to present a basic understanding of the main aspects considered for high-value product formation.

8.3.1

Dynamics of Algal Biomass Composition

Dynamics of algal biomass composition needs proper understanding in order to progress for the creation of new platform chemicals from the existing ones. In this regard, we focus on the three popular and important genera of microalgae including Chlorella, Scenedesmus, and Nannochloropsis. The three basic macromolecular biomass constituents namely lipids, carbohydrates, and proteins at specific percentages for a particular species would play a huge role in selecting a particular type of bio-product refinement. Subdivision of each of these macromolecular structures depends variedly on biomass growth-specific stages and physiological culturing conditions [9, 17]. Recently, Laurens et al. [9] reported a distinctive accumulation pattern of the three macromolecular constituents for the three genera during different phases of their growth cycle. It was reported by the authors that for all the three genera, protein accumulation was inversely proportional to that of carbohydrates and lipids, whereas for Scenedesmus carbohydrate synthesis did not affect lipid accumulation but for the other two genera lipid accumulation increased only when carbohydrate content decreased. Thus, the observation indicated that the microalgae preferably chose carbohydrate as the primary storage form in their early life cycle and from their mid-log phase preferred to convert them into lipid, thereby helping to sustain or increase their longevity. This compositional shift can help in understanding the recovery time for a particular product. For instance, if increasing the lipid content is the main focus then rather going for starvation mode if the cultures are detained in their early log phase for a longer period of time then perhaps, more carbohydrate can be produced and simultaneously more lipid can be generated. This phenomenon can happen if we keep providing additional nutrients

8.3 Analysis of Conversion Methods

not in batch mode but in intermittent mode, thus giving the culture more time to acclimatize. As the basic nature of the microalga is to synthesize carbohydrate first, therefore, by changing such cultivation methodology is believed to enhance lipid accumulation process at a rapid rate. Perhaps, when the culture would enter a fascinating log phase, then the cultivation strategy should be that no addition of nutrients takes place and rather a batch process is followed. Another important aspect could be considering the value of derived products against the extra cultivation time need to maximize the yield of a target product. Laurens et al. [9] stated that few techno-economic analysis modeled methods such as minimum fuel-selling price (MFSP) and total capital investment (TCI) showed consistent engineering feasibility-level outputs. Perhaps, they emphasized that biomass feedstock cost is a function of biomass productivity cost and also found that the productivity cost remained constant throughout the nutrient depletion period [9]. Insights taken from this study suggest that if the formation and extraction of several component parts are allowed, then the productivity cost may be lowered down and therefore, the overall production cost can be lowered.

8.3.2

Conversion Routes

The basic aim of a biorefinery is to combine various feedstock to produce energy and other platform chemicals employing various transformation routes via process integration, while reducing the environmental impacts and consuming less fossilized energy during the process [18]. Common routes for conversion of biomass are through thermochemical conversion techniques such as gasification, combustion, and pyrolysis and biochemical conversion techniques such as pretreatment, hydrolysis, physical, chemical, or biological [18]. Sometimes combining both thermochemical and biochemical conversion can result in higher yield of the product; for example in case of biofuel production, a combined biochemical process is mostly preferred [18]. Along with the conversion technique, feedstock type is also very important. Choosing the correct type of conversion method for a particular type of feedstock would result in greater yield or productivity. For instance, if lipids are the target and algae is the feedstock, then hydrothermal carbonization (HTC) could be a better option as this method does not require any pretreatment or drying step. On the other hand, if biochar production from lignocellulosic biomass is the target then slow pyrolysis is a better option. The mentioned conversion techniques are discussed earlier by De Bhowmick et al. [4]. Interestingly, each of these conversion methods results in the release of other by-products in the process such as four main outputs are observed, while subjecting the algal biomass for HTC. The main gas released (CO2 ) could be usefully trapped in cylinders and recycled back for the next batch of cultivation. Further, the aqueous phase containing nitrogen, phosphorous, and other organic components could also be recycled as nutrient for the next batch of run. Furthermore, the char produced during the process could be possibly used for soil amendment or as adsorbent [17]. All these examples

203

204

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

discussed fall under the three main concepts of biorefinery approach and they are: (i) hierarchy; (ii) sequencing; and (iii) integration [18]. The decomposition of elements belonging to a biorefinery as feedstock, products, and technology falls under the hierarchy concept. Similarly, the logical step-by-step synthesis relating to technological pathway with the product is considered as a sequencing concept. Finally, integration of complex portfolios associated with integration of feedstock, technologies and products completes the biorefinery approach methodology.

8.3.3

Product Yield and Market Value

The current situation of commercialization of products from microalgae biorefinery is in its infancy with a single product-centric focus. Most recently, Parieto et al. [19] proposed the idea of using multi-objective mixed integer nonlinear programming (MINLP)-based methods to optimize the production of a value-added product from the single product-centric target. The study showed that the production of methanol from glycerol, which was a by-product produced during oil transesterification reaction and the integrated technique, resulted in an operational cost of $0.05–$0.16 L−1 higher than the conventional way of production advantageously without compromising on the environmental fronts. Sawaengsak et al. [20] also hypothesized that it is possible to incur higher profitable margin by integrating the production of value-added products from the mainstream target product generation. However, the study also indicated that in order to make the net present ˇ cek et al. value positive, the CAPEX and OPEX must be assessed first. Likewise, Cuˇ [21] also developed a multi-period synthesis and optimization model, wherein the capital cost calculation showed a positive impact on market completion between fuel and food processing requirements. Assessing various biorefinery products, it has been observed that biorefinery of functional proteins requires sophisticated techniques resulting in high production cost, whereas biorefinery of commodity chemicals such as biofuels, chemicals, and food/feed results in much lower production cost (€0.9–€1.1/kg). Desai et al. [22] reported a case study wherein it was found that by using a two-phase aqueous extraction method, the cost of recovery was significantly reduced. Therefore, in order to build a good logistics strategy for the product, the yield recovery is a very important aspect to look into. All these examples suggest that by choosing wisely and appropriately the right technique for a particular product, the production cost could be successfully reduced, while gaining on yield recovery rate. In order to realize the potential commercial success of algae products, two main aspects need special attention: (i) exploitation of biomass for bulk commodity production and (ii) special markets for these products. Ruiz et al. [23] analyzed the European Union microalgae product market and found that the market for bulk chemicals (biopolymers, bio-lubricants, bio-solvents, and surfactants) has a substantial demand with a rise in 2.4 × 106 tons/year. In contrast, when the value of biomass in different markets were assessed, it was found that the lowest revenue was collected for biofuels (€0.3/kg), whereas considering the food additive potential of microalgae biomass the market value was three times higher with a greater

8.4 Competent Applications of Algae

revenue of more than €2/kg. More interesting results were obtained when the market value of natural pigments such as lutein, beta-carotene, and astaxanthin were assessed and it was found that the market volume reached almost $1 billion. It is these high-value products that would help to sustain the market risks in terms of commercializing the microalgae feedstock. In another study, Budzianowski [10] reported that the market for bio-products is developing dynamically and is a rise of about 20% is expected in the compounded annual growth rate (CAGR). Primarily, it has been identified that bio-products with two times higher selling prices than their production cost are of great potential for emerging companies. It has been assumed that the new bio-products remaining in their early development phase will still have growing market size with high selling prices and therefore, the market entrance of a new product could still be overwhelming. Another opportunity identified by Budzianowski [10] was the introduction of a new bio-product that does not have any competition with the existing products and would help to promote economies in the new area. However, the actual market value will depend mainly on technological developments with real solid policies supporting the bioeconomy. Similar strategies are recommended for biocosmetics, bionutrients, and biopharmaceutical industry with perishable replacement of algae/plants extracts.

8.4 Competent Applications of Algae Every energy counts a great deal with the dramatic increase in global energy supply both economically and socially [24]. The International Energy Agency (IEA) estimated approximately 53% increment in global energy consumption by 2030 and the total energy consumption of developing countries will supposedly exceed with that of the developed countries [8]. With such high demand for energy generation, the dependence on the duration of fossil resource consumption is absolutely not optimistic [8]. Therefore, algae have come into the limelight of research in an effort to search for sustainable development of renewable energy, while controlling environmental pollution. Few of the potential applications of algae discussed recently by De Bhowmick et al. [17] are depicted in Figure 8.1 and listed below: ●

●

● ●

Utilizing flue gas as a source of CO2 will provide a very promising alternative in reducing greenhouse gas emissions. Wastewater treatment by utilizing wastewater contaminants like ammonium nitrogen, nitrate, and phosphorous as a nutrient source for cultivating algal biomass. The remnant biomass can be used as livestock feed/organic fertilizers. High-value additive products such as (i) mineralized carbon for stable sequestration, (ii) algal surfactants, and (iii) hydrocarbons are of great importance for potential application of algae.

One of the key advantages of using high CO2 content feeding stream such as flue/flaring gases containing 5–15% CO2 is that it allows more efficient capturing of CO2 than terrestrial plants, which typically absorb only 0.03–0.06% [24]. Few studies

205

206

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

Flue gas CO2

Microalgal cell factory

Wastewater CO2 remediation sequestration

Figure 8.1

Biofuel

Biochar High-value products

Organic fertilizer

Reduced greenhouse gas emission

Competent applications of microalgal feedstock.

have shown that microalgal species like Monoraphidium minutum, Chlorococcum littorale, and Chlorella strains from hot spring were able to efficiently fix flue gas CO2 10–50 times higher than other terrestrial plants and were also able to assimilate CO2 from soluble carbonates [24]. Recorded bio-fixation rate of 53.29% for 6% (v/v) flue gas CO2 and 45.61% for 12% (v/v) flue gas CO2 was achieved by using Spirulina sp. Additionally, up to 18% (v/v) flue gas CO2 bio-fixation was reported for S. obliquus and C. kessleri strains [24]. The life cycle analysis (LCA) showed that co-firing coal and microalgae helped in reducing methane and CO2 and that actually helped in lowering the net SOx and NOx particulates [24]. Among all the potential applications, the integration of algal biomass cultivation using wastewater is the most promising solution, since the algae biofuel production appears to be strongly economically convenient only in conjunction with wastewater treatment [17]. Utilizing wastewater offers the feasibility to recycle these nutrients (organic contaminants and heavy metals) into the biomass, while reducing the treatment cost and helping in releasing oxygen-rich effluents into water bodies [17, 24]. Using a wide variety of wastewater has resulted in proven accumulation of lipids, proteins, starch, and other hydrocarbons, all of which can be efficiently converted into biofuels [25]. Perhaps, if 50% (i.e. approximately 495 billion m3 ) of the consumed water turned into wastewater is used for algal biomass cultivation, there is potential to generate about 247 million tons of biomass equating up to 37 million tons of oil [24]. The idea of cultivating algae is not new, but it has been observed that approximately 717 research papers were published in the year 2010 indicating the exponential growth of microalgae bioenergy research. Most of the studies have focused on algae culturing methods and cultivation systems, systems engineering, metabolic and genetic engineering, harvesting/dewatering measures, biofuel conversion technologies, life cycle assessment, and policy implications [8, 24]. Despite grabbing increasing interest, algal technology has witnessed a serious and obvious dilemma regarding its bulk commercial production [4, 17]. The dilemma stems around overwhelming capital and operational costs. Nonetheless, the inevitable question that arises: Is it too luxurious to cultivate microalgae only for biofuel applications? Taking into account the lack of availability of commercial-scale data

8.5 Biorefinery and Integrated Approaches

for biofuel production cost, several studies have sought to produce an overall estimation of the biofuel production cost [8]. Importantly, the estimated reports vary widely due to several factors such as (i) various production capacity scales; (ii) different product portfolios; (iii) various production routes and motivations; (iv) minimization of discounted break even time; and (v) maximization of return on event [8]. Additionally, the design of a production plant can also affect the overall production cost, making the comparison scale difficult [8, 17]. Therefore, a proper reliable production estimation cost is a dire need in order to understand the complex logistics and navigate accordingly.

8.5 Biorefinery and Integrated Approaches Biorefinery is an industrial process, wherein the biomass is converted into a range of biochemical materials and energy products. It is a facility that is actually analogous to current oil refinery, where multiple fuels and various products are produced from fossil oils [4]. The biorefinery chain comprises of pretreatment and separation of primary biomass components and their subsequent conversion to secondary components through a cascade of events [4]. Major utilization and recycling efforts are being made to prevent the raw material resource loss. From the view point of high-value products, microalgae are the most appropriate source for production of fine chemicals, animal feed, human nutrition, pharmaceuticals, and cosmetics [4, 8]. Additionally, considering the environmental impacts, microalgae are a resource-efficient feedstock [8]. However, due to significantly low biomass productivity following the autotrophic cultivation method, algal bioprocessing has shifted its focus toward mixotrophic and heterotrophic modes by implementing several multiple strategy-based biorefinery approaches [26]. Integrating the biorefinery concept with wastewater treatment will provide efficient utilization of residual waste, reducing the overall production cost while favoring sustainable economics. The residual/remnant biomass after lipid/pigments extraction can be subjected to various biochemical and thermochemical conversion processes such as fermentation, anaerobic digestion, and production of biochar [27]. Other integrated approaches include ecological engineering/synthetic ecology wherein recovery of more products will be enhanced using artificial biomimetic systems. It is visualized that the existing algal biorefinery system would certainly benefit from the beneficial effects of the bacteria for algal growth promotion [26]. Few such examples are: (i) two-stage production of value-added chemicals using algae and methane-oxidizing bacteria developed for simultaneous production of biodiesel and bioplastics [28]; (ii) co-cultivation of algae with growth-enhancing bacteria , which resulted in approximately 10–70% enhancement in algal growth rate [29]; (iii) algal-bacterial interaction helping in flocculation by quorum sensing [30]; (iv) along with bio-flocculation, application of algicidal bacteria, which was observed to help in lipid extraction process due to their algal cell lysis property [31].

207

208

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

Another integrated approach could be bioelectrogenesis, wherein chemical energy is transformed into electricity through a cascade of redox reactions under anaerobic conditions [26]. Apart from power generation, microbial fuel cells (MFCs) with minimal modifications based on their functional utility can be variedly used as bioelectrochemical treatment system (BET), bioelectrochemical system (BES), microbial electrolysis cell (MEC), and microbial desalination cell (MDC) [26]. Apparently, the nutrients formed during the breakdown process are resealed as complex molecules that can be used for gardening as bio-fertilizers. Likewise, integrating dark fermentation and MFCs with MEC will aid in high recovery of biogas resulting in complete treatment of wastes [26]. Additionally, biological mineralization via microbial degradation through plants and algae is proposed to be one of the reliable and long-term methods for nutrient removal in constructed wetlands [26]. Perhaps, this process will enhance the growth of protein-rich aquatic biomass and ferns that can serve as additional means of nutrient removal while resulting in simultaneous production of biofuels, bio-fertilizers, and animal feed [26]. Amalgamation of various approaches will result in developing a successful conceptual biorefinery model that can be realized to its full potential through concerted efforts of multidisciplinary scientists, engineers, and economists working as one team.

8.6 Technological Issues: Pros and Cons The central aspect of the algal process system is choosing the right algae species and growth technology. Major issues of the algal process system stem around photosynthetic efficiency, sunlight requirement, degassing system, and versatility of the selected strain and energy requirement. Though the open raceway ponds are a much cheaper option, due to significant drawbacks in perfecting the growth leaves the use of this system is a debatable issue. On the other hand, closed systems are found to be twice as expensive considering per gallon of lipid obtained and are the most energy-intensive systems as far as power is concerned [5–7]. In addition to cultivation system, selection of strain is also very important in streamlining the process [6, 7]. The Aquatic Species Program by the US Department of Energy evaluated thousands of species based on few characteristics such as high growth rates, auto-flocculation, size, environmental tolerance, and robustness [5]. In conjunction with desirable characteristics, the research findings have further shown that the lipid content in most of the species ranged from 20 to 30% of its dry cell weight but under some conditions may rise up to 70% [5]. However, increasing the lipid content by stressing the cells could result in retarded growth, which is again debatable. Therefore, there are a lot of challenges that require special attention and emphasis while developing a product from microalgae. Harvesting, dewatering, and drying of algal biomass are primary steps in concentrating the biomass while removing large volumes of water. Centrifugation, membrane filtration, flocculation, (electro-)flotation, coagulation, and sedimentation are few of the well-known methods of harvesting the biomass [4]. Except for auto-flocculation, almost all the techniques are cumbersome and energy-intensive

8.6 Technological Issues: Pros and Cons

especially because microalgae can be just a few microns in size. Among all these harvesting techniques, flotation and thickening is comparatively cheap but is able to reduce 10% of the solid biomass concentration followed by filtration that can effectively further dry the biomass up to 40%, however this techniques does requires heavy maintenance cost [5]. Centrifugation also returns a relatively dry product of 32% solids but requires lot of energy for its operation [5]. Ruiz et al. [23] estimated the total CAPEX (capital cost) and OPEX (operational cost) and found that the cost of microalgae production varied with the production system but showed different trends in operation of open and closed systems. If wastewater is used instead of freshwater, then it played the most influential role in open system, minimizing the raw material requirement up to 17–23% and energy usage up to 14%. Based on a statistical estimation, it was observed that the total production cost was drastically reduced in open ponds (up to 30–40%) when wastewater was used. Though the microalgae biorefinery is in its infancy and the existing few are heading toward a single product-centric commercial focus, the products are simply based on harvesting, drying, and extraction techniques. For example, Ruiz et al. [23] reported that a 100-ha cultivation system in Spain with a bulk production facility dedicated to fuels, chemicals, or food resulted in €1/kg of biorefinery cost. In this case study, the two major identifiable economic bottlenecks were (i) energy involved in cell disruption (roughly, ∼1 kW (h kg) heat dissipated corresponds to €0.15/kg), and (ii) heat and energy used to dry biomass, extracting lipid, and recovering solvent altogether roughly corresponding to €0.18/kg. To counter these disruption techniques, theoretically up to 90% reduction in energy can be achieved by using novel disruption techniques, which include hydrodynamic cavitation, electrical fields, electromagnetic field, and acoustic cavitation. Similarly, for solvent extraction of lipids, less energy-intensive techniques are supercritical fluids and switchable solvents. Additionally, a case study in Spain showed that purification of high-value products such as omega-3 fatty acids, pigments, and water-soluble proteins resulted in €3/kg in overall production cost [23]. It has been estimated that biorefinery of bulk commodity markets such as biofuel, chemical, and food/feed leads up to €0.9–€1.1/kg in overall production cost. DOE of National Algal Biofuels Technology roadmaps instructed that if bio-products can be produced along with fuels for approximately $0.67–$2.2/kg at a rate of 10 000–1 000 000 tonnes/year the candidate can be considered as a good option [9]. In another case study, it was found that the biorefinery that focused on H2 via supercritical fluid extraction (SFE), oil and pigment production became the most commercially successful one. One such example is Anabaena sp. biorefinery including H2 production by autotrophic and dark fermentative pathway by Enterobacter aerogenes [1]. Campenni et al. [32] demonstrated Chlorella protothecoides biorefinery, wherein the algae were grown autotrophically under salinity and luminosity stress condition for ethanol or hydrogen production while bearing sufficient amount of lipid and carotenoids. Similarly, Olguin [33] studied an integrated biorefinery system for biogas, biodiesel, and biohydrogen production. Spirogyra biorefinery for biohydrogen and pigment production with 62% of reduction in

209

210

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

overall energy consumption was reported by Pacheco et al. [34]. Another promising hydrothermal microwave pyrolysis microalgae biorefinery concept devoid of water removal requirement was proposed by Budarin et al. [35]. However, in spite of so many microalgae biorefinery proposals over the past, majority of the models was not actually commercialized due to the potential costs of the technology.

8.7 Life Cycle Assessment LCA is the primary and most commonly used environmental assessment tool weighing the aspects of algal bioprocesses [36]. Among all the processes, biodiesel production has been the common focus of the LCA studies. Impactful LCA studies can be observed for energy consumption and greenhouse gas emissions [36]. Other common LCA indicators include water demand, land occupation, and eutrophication potential of the process [36]). Recent studies indicate that the techno-economic assessments of microalgal biorefineries are the essential element for strategic planning and decision-making that helps in evaluating the project value [36]. Apparently, the first detailed economic evaluation provided a comprehensive CAPEX and OPEX (per barrel of oil produced) of common open raceway ponds [37]. An exhaustive update including novel reactor configurations and sensitivity analysis is therefore required to present a more feasible option. Recent LCA studies have shown that economic performances of algal systems has been improved by producing high-value products/compounds, while considering the production of biodiesel from lipid fractions, animal feed from protein fractions, and electricity from carbohydrate fractions [38]. Importantly, representative product substitutes must be included along with considering consequences related to GHG emission or savings from by-products and land use. Gnansounou and Raman [38] emphasized on algal systems as being a better system than fossilized systems that were also found to be more environmentally friendly than the respective reference system [38]. Overall, algal systems with biodiesel, protein, and succinic acid were potentially found to be an alternative renewable option reducing environmental impacts while simultaneously generating employment opportunities in developing countries like India [38]. Souza et al. [39] reported life cycle fossil energy use and GHG emissions of biofuels produced from integrated algal biorefinery system and suggested that by replacing diesel by algae biodiesel, substantial improvement in life cycle GHG profile of sugarcane ethanol was observed. Around 10–50% improvement compared to traditional Brazilian sugarcane ethanol distillery was achievable. Another assessment on economic feasibility and environmental impact of open-lagoon wastewater treatment plant with a rotating algal biofilm reactor cultivation facility and HTL conversion plants showed that cultivating algae were found to be very active in reducing environmental impacts. In addition, the sensitivity analysis suggested algal productivity to be most significantly affecting fuel prices, while showcasing the importance of using optimization techniques in order to increase the biomass productivity [39]. However, direct comparisons of studies are difficult based only on differences in

8.8 Techno-Economic Analysis (TEA)

system boundaries and core LCA assumptions. Perhaps, assumptions with regard to co-product allocation methods, electrical energy sourcing, and life cycle inventory data dramatically impacted the outcome [40]. Additionally, differences in processing pathways and unrealistic validation of sub-processing models with small-scale data contribute to the large variability in reported data. As such there is a need to understand the sub-process technologies on systems level and then integrate the modeling efforts. Utilization of consistent system boundaries will facilitate the comparison of processing technologies providing better understanding of the impact and limitations of biofuels derived from microalgae on the metrics of economic feasibility, resource demand, and environmental concerns [40]. Technically, the net energy ratio (NER), GHG emissions, water consumption and wastewater integration are the main concerns while working out the LCA. Gaining on each and every point will certainly raise the bar for considering microalgae cultivation for various products along with fuel generation in a biorefinery model. Therefore, an improved necessary economically viable and sustainable system is required to make the microalgae technology successful.

8.8 Techno-Economic Analysis (TEA) Techno-economic analysis (TEA) is another foundation tool used for evaluating the feasibility of a particular process that is capable of analyzing alternative processing technologies, while helping in generating production pathways that are commercially viable [40, 41]. Different cultivation systems such as open ponds, flat panel, and vertical and horizontal photobioreactor systems have been used to analyze the commercial viability of the biorefinery system [41]. According to a study by Chew et al. [41], an efficient biorefinery system utilizing remnant biomass and lipids as nutrient-rich source for animal feeds and vegetable oils would serve excellent economic and environmental purposes. Additionally, they also reported that simple construction-based open ponds performed better on lowering the capital expenditure and operating cost but the photobioreactors were more advantageous considering microalgae productivity, nutrient uptake efficiency, reaching high culture density and easy controllability. Another TEA performed by Thomassen et al. [42] suggested that using specialized membrane-based recycling medium option installed in open ponds could result in feasible and profitable techno-economic assessment. They highlighted few issues such as land occupancy and excessive water consumption lead to scarcity of water affecting water cost and additional requirement of water treatment process. Further, the lack of public agreement and associated uncertainties restricted the economic viability of microalgae fuels at the present situation [43]. Mostly, producing microalgae includes the harvesting equipment and pond/reactor cost. Free CO2 , low raw material cost, and recycled nutrient recovery make the process still feasible [41]. However, in reality production of microalgae for biofuel is not able to compete with the use of fossil fuel easily despite the availability of the technology with improved market economics. Besides, the increment in oil prices and incentives on

211

212

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

carbon-intensive projects may help in accelerating the adoption of renewable fuels from microalgae with enhanced sustainability and life cycle implications.

8.9 Futuristic Options Biosyngas-based refinery, pyrolysis-based refinery, hydrothermal upgrading-based refinery, and fermentation-based refinery are the four main types of boirefinery systems, which are shown in Figure 8.2. The main purpose of a biorefinery is to integrate the production of commodity chemicals and high-value products along with fuel and energy generation, while optimizing resources and minimizing waste usage up to maximum benefit [1]. Various possibilities of algae-based biorefinery via process integration options are shown schematically in Figure 8.3. Other than the integrated processes suggested in Figure 8.2, several achievable algae-biorefinery frameworks include the following (Adapted from Trivedi et al. [1]; De Bhowmick et al. [4]): ● ●

●

● ● ●

Water recyclability can considerably improve economics. Co-production of value-added by-products either directly or indirectly with food production sector. Few algal species were encouraged for fermentative biogas production via combined biorefinery where the species acted as a substrate for fermentation. Anaerobic fermentation holds potential as well. Highly controlled process parameters are required to reciprocate the same result. LCA for any product must be considered to understand its feasibility.

Hydrothermal

Power/Heat

Materials

Biosyngas

BIOREFINERY

Chemicals Transportation fuels

Pyrolysis

Figure 8.2

Current biorefinery types.

Fermentation

References

Non-fossil based methanol

Dairy industry Co-production Co-pyrolysis Value-added biochar

Agricultural waste

Co-production

Algae-based biorefinery (integration options)

Aquaculture

Inland-based animal production system for proteins

Co-production

Lignocellulosic industry

Cellulase/hemicellulase enzymes for hydrolysis

Figure 8.3

Process integration options for algae-based biorefinery system.

Although numerous reports are available on microalgae biorefinery, none of them are envisioned for simultaneous production of similarly large volumes of multiple products to avoid saturation of any one particular market. Laurens et al. [9] rightly identified three pivotal criteria for viable biorefinery: (i) price of the bio-derived product should be the primary driver of the existing biorefinery; (ii) price is still the primary driver in the case of identical functional performances of new commodity products; and (iii) priority should be given to entirely new materials with unique beneficial functional properties [9]. However, due to unpredictable market volume and price target for new novel products from algae, the last criterion is apparently the most difficult to fulfill of all the three. Today, dozens of start-up companies are aiming at commercializing algal fuels and bio-products, but according to the editorial survey (2011) the current production methods need significant improvement [23, 44]. Overall, the technology for lipids is well established and robust but fractionation of major components such as proteins and carbohydrates from lipids and simultaneous recovery of commodity chemicals require a major breakthrough. Development of cheaper novel filtration systems for fractionation, extraction using novel solvents such as ionic liquids/surfactants, blending of algal lipid with existing petrodiesel, reusability of waste water, in-depth LCA, optimization in cultivating algae, and reducing GHG emissions are the need of the hour.

References 1 Trivedi, J., Aila, M., Bangwal, D.P. et al. (2015). Algae based biorefinery – how to make sense? Renewable and Sustainable Energy Reviews 47: 295–307. 2 Cheah, W.Y., Show, P.L., Chang, J.S. et al. (2015). Biosequestration of atmospheric CO2 and flue gas-containing CO2 by microalgae. Bioresource Technology 184: 190–201.

213

214

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

3 Zhu, L.D. and Hiltunen, E. (2016). Application of livestock waste compost to cultivate microalgae for bioproducts production: a feasible framework. Renewable and Sustainable Energy Reviews 54: 1285–1290. 4 De Bhowmick, G., Sarmah, A.K., and Sen, R. (2018). Lignocellulosic biorefinery as a model for sustainable development of biofuels and value added products. Bioresource Technology 247: 1144–1154. 5 Gebreslassie, B.H., Waymire, R., and You, F. (2013). Sustainable design and synthesis of algae-based biorefinery for simultaneous hydrocarbon biofuel production and carbon sequestration. Process Systems Engineering 59: 1599–1621. 6 De Bhowmick, G., Koduru, L., and Sen, R. (2015). Metabolic pathway engineering towards enhancing microalgal lipid biosynthesis for biofuel application – a review. Renewable and Sustainable Energy Reviews 50: 1239–1253. 7 De Bhowmick, G., Vegesna, N., and Sen, R. (2015). Process design for augmentation and spectrofluorometric quantification of neutral lipid by judicious doping of pathway intermediate in the culture of marine Chlorella variabilis for biodiesel application. Bioresource Technology 198 A: 781–788. 8 Zhu, L. (2015). Biorefinery as a promising approach to promote microalgae industry: an innovative framework. Renewable and Sustainable Energy Reviews 41: 1376–1384. 9 Laurens, L.M.L., Markham, J., Templeton, D.W. et al. (2017). Development of algae biorefinery concepts for biofuels and bioproducts; a perspective on process-compatible products and their impact on cost-reduction. Energy and Environmental Science 10: 1716–1738. 10 Budzianowski, W.M. (2017). High-value low-volume bioproducts coupled to bioenergies with potential to enhance business development of sustainable biorefineries. Renewable and Sustainable Energy Reviews 70: 793–804. 11 Zhang, X., Rong, J., Chen, H. et al. (2014). Current status and outlook in the application of microalgae in biodiesel production and environmental protection. Frontiers in Energy Research 19 https://doi.org/10.3389/fenrg.2014.00032. 12 Adeniyi, O.M., Azimov, U., and Burluka, A. (2018). Algae biofuel: current status and future applications. Renewable and Sustainable Energy Reviews 90: 316–335. 13 Koizumi, T. (2015). Biofuels and food security. Renewable and Sustainable Energy Reviews 52: 829–841. 14 Eppink, M.H.M., Olivieri, G., Reith, H. et al. (2017). From current algae products to future biorefinery practices: a review. In: Advances in Biochemical Engineering/Biotechnology, vol. 166 (eds. K. Wagemann and N. Tippkötter), 1–25. Cham: Springer https://doi.org/10.1007/10_2016_64. 15 Gong, M. and Bassi, A. (2016). Carotenoids from microalgae: a review of recent developments. Biotechnology Advances 34: 1396–1412. 16 Dong, T., Knoshaug, E.P., Lieve, R.D. et al. (2016). Combined algal processing: a novel integrated biorefinery process to produce algal biofuels and bioproducts. Algal Research 19: 316–323. 17 De Bhowmick, G., Sarmah, A.K., and Sen, R. (2019). Zero-waste algal biorefinery for bioenergy and biochar: a green leap towards achieving energy and environmental sustainability. Science of the Total Environment 650: 2467–2482.

References

18 Moncada, J.B., Aristizábal, V.M., and Cardona, C.A. (2016). Design strategies for sustainable biorefineries. Biochemical Engineering Journal 116: 122–134. 19 Parieto, C.V.G., Ramos, F.D., Estrada, V. et al. (2017). Optimization of an integrated algae-based biorefinery for the production of biodiesel, astaxanthin and PHB. Energy 139: 1159–1172. 20 Sawaengsak, W., Silalertruksa, T., Bangviwat, A. et al. (2014). Life cycle cost of biodiesel production from microalgae in Thailand. Energy for Sustainable Development 18: 67–74. ˇ cek, L., Martín, M., Grossmann, I.E., and Kravanja, Z. (2014). Multi-period 21 Cuˇ synthesis of optimally integrated biomass and bioenergy supply network. Computers and Chemical Engineering 66: 57–70. 22 Desai, R.K., Streefland, M., Wijffels, R.H. et al. (2014). Extraction and stability of selected proteins in ionic liquid based aqueous two phase systems. Green Chemistry 16: 2670–2679. 23 Ruiz, J., Olivieri, G., de Vree, J. et al. (2016). Towards industrial products from microalgae. Energy and Environmental Science 9: 3036–3043. 24 Suganya, T., Varman, M., Masjuki, H.H. et al. (2016). Macroalgae and microalgae as a potential source for commercial applications along with biofuels production: a biorefinery approach. Renewable and Sustainable Energy Reviews 55: 909–941. 25 Nayak, M., Karemore, A., and Sen, R. (2016). Performance evaluation of microalgae for concomitant wastewater bioremediation, CO2 biofixation and lipid biosynthesis for biodiesel application. Algal Research 16: 216–223. 26 Mohan, S.V., Nikhil, G.N., Chiranjeevi, P. et al. (2016). Waste biorefinery models towards sustainable circular bioeconomy: critical review and future perspectives. Bioresource Technology 215: 2–12. 27 De Bhowmick, G., Sarmah, A.K., and Sen, R. (2018). Production and characterization of a value added biochar mix using seaweed, rice husk and pine sawdust: a parametric study. Journal of Cleaner Production 200: 641–656. 28 Van der Ha, D., Nachtergaele, L., Kerckhof, F.M. et al. (2012). Conversion of biogas to bioproducts by algae and methane oxidizing bacteria. Environmental Science and Technology 46: 13425–13431. 29 Cho, D.H., Ramanan, R., Heo, J. et al. (2015). Organic carbon, influent microbial diversity and temperature strongly influence algal diversity and biomass in raceway ponds treating raw municipal wastewater. Bioresource Technology 191: 481–487. 30 Lee, J., Cho, D.H., Ramanan, R. et al. (2013). Microalgae-associated bacteria play a key role in the flocculation of Chlorella vulgaris. Bioresource Technology 131: 95–201. 31 Lenneman, E.M., Wang, P., and Barney, B.M. (2014). Potential application of algicidal bacteria for improved lipid recovery with specific algae. FEMS Microbiology Letters 354: 102–110. 32 Campenni, L., Nobre, B.P., Santos, C.A. et al. (2013). Carotenoids and lipids production of autotrophic microalga Chlorella protothecoides under nutritional, salinity and luminosity stress conditions. Applied Microbiology and Biotechnology 97: 1383–1393.

215

216

8 Pertinent Issues of Algal Energy and Bio-Product Development A Biorefinery Perspective

33 Olguin, E.J. (2012). Dual purpose microalgae-bacteria-based systems that treat wastewater and produce biodiesel and chemical products within a biorefinery. Biotechnology Advances 30: 1031–1046. 34 Pacheco, R., Ferreira, A.F., Pinto, T. et al. (2015). The production of pigments & hydrogen through a Spirogyra sp. biorefinery. Energy Conversion and Management 89: 789–797. 35 Budarin, V., Ross, A.B., Biller, P. et al. (2012). Microalgae biorefinery concept based on hydrothermal microwave pyrolysis. Green Chemistry (12): 3251–3254. 36 Pérez-López, P., Montazeri, M., Feijoo, G. et al. (2018). Integrating uncertainties to the combined environmental and economic assessment of algal biorefineries: a Monte Carlo approach. Science of the Total Environment 626: 762–775. 37 Benemann, J.R. and Oswald, W.J. (1996). Systems and Economic Analysis of Microalgae Ponds for Conversion of CO2 to Biomass. Pittsburg, PA: Pittsburgh Energy Technology Center. 38 Gnansounou, E. and Raman, J.K. (2016). Life cycle assessment of algae biodiesel and its co-products. Applied Energy 161: 300–308. 39 Souza, S.P., Gopal, A.R., and Seabra, J.E.A. (2015). Life cycle assessment of biofuels from an integrated Brazilian algae-sugarcane biorefinery. Energy 81: 373–381. 40 Quinn, J.C. and Davis, R. (2015). The potentials and challenges of algae based biofuels: a review of the techno-economic, life cycle, and resource assessment modeling. Bioresource Technology 184: 444–452. 41 Chew, K.W., Yap, J.Y., Show, P.L. et al. (2017). Microalgae biorefinery: high value products perspectives. Bioresource Technology 229: 53–62. 42 Thomassen, G., Vila, U.E., Dael, M.V. et al. (2016). A techno-economic assessment of an algal-based biorefinery. Clean Technologies and Environmental Policy 18: 1849–1862. 43 Manganaro, J.L., Lawal, A., and Goodall, B. (2015). Techno-economics of microalgae production and conversion to refinery-ready oil with co-product credits. Biofuels Bioproducts and Biorefining 9: 760–777. 44 Chisti, Y. and Yan, J. (2011). Energy from algae: current status and future trends algal biofuels – a status report. Applied Energy 88: 3277–3279.

217

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater Dong Wei, Bin Du, and Qin Wei Collaborative Innovation Center for Green Chemical Manufacturing and Accurate Detection, University of Jinan, Jinan, PR China

9.1 Introduction Over the recent decades, nitrogen-containing wastewater has been increasingly produced in various ways, such as from fertilizer wastes, packing plants, feed processing industries, pharmaceutical factories, aquaculture industries, and landfill leachates [1]. Activated sludge (AS) system is widely used in municipal and industrial wastewater treatments via anoxic and aerobic processes for biological nitrogen and phosphorus removal. However, a variety of disadvantages exist in current AS systems, such as sludge bulking, high sludge treatment cost, and large treatment area [2]. As a special case of biofilm, granular sludge has attracted much attention because of its more advantageous physicochemical properties compared to AS during the treatment of wastewater. Granular sludge includes two forms: anaerobic and aerobic granular sludge (AnGS and AGS) [3]. Compared to flocculent sludge, granular sludge has much denser microbial structure, higher sludge biomass, better treatment performance and separation capacity, etc. Based on operational evaluation, De Bruin et al. [4] considered AGS bioreactors to be an attractive alternative to conventional AS technology in primary as well as post treatment. Due to the difference in dominant microbial populations of the two sludges, AnGS is widely used in high-concentration organic industrial wastewater treatment, whereas AGS is used mainly for treating municipal wastewater [5, 6]. In the recent years, one of the biggest drawbacks of sludge-based processes has been their large excess sludge production. As a by-product of sewage treatment, excess sludge produced from the sewage treatment plant for domestic wastewater treatment often contains a large number of bacteria, microorganisms, parasites, suspended and colloidal substances, also containing N, P, and other nutrients. The disposal of excess sludge in sewage treatment plant is often expensive [7]. As a carbon-rich material, excess sludge has been widely applied in the synthesis and preparation of porous activated carbons (ACs). So far, many activation-influencing Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

218

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

factors have been investigated for improving the characteristics of sludge-based AC, including activation type, activating agent, pyrolysis/activation temperature, and impregnation ratio [8]. Moreover, sludge-based AC as biosorbent has been applied to remove various organic and inorganic pollutants with the advantages of low cost, high efficiency, and large adsorption capacity [9, 10]. Recent literatures have also focused on the modification of sludge-based AC surface aimed at enhancing the adsorption capacity of various pollutants [11]. Granular sludge, including anaerobic and aerobic forms, has a better settling ability and higher adsorption performance than those of AS. The different types of pollutants sorption onto granular sludge-based biosorbent has been extensively studied in view of adsorption condition and relative mechanism. In particular, extracellular polymeric substances (EPSs) have important significance in regard to the formation, structural stability, and physicochemical property of granular sludge as a three-dimensional matrix [12]. Thereby, recent studies have also reported the important role of EPS in the biosorption of heavy metal from aqueous solution onto AGS and AnGS [13]. Moreover, many spectroscopic approaches have been used for in-depth investigation of the interaction between EPS and pollutions for fully understanding the biosorption mechanism. Additionally, EPSs are also regarded as effective bioflocculants and adsorbents that are extensively used in wastewater treatment due to their abundant functional groups and binding sites [14]. In this chapter, the resource utilization of sludge and its potential environmental applications for the treatment of various pollutants is systematically reviewed. Additionally, the preparation, characteristics, and biosorption performance of sludge-based carbon are discussed. Also, the biosorption behavior and mechanism of granular sludge for heavy metal and organic pollutions are summarized. Especially, the role of EPS in the granular sludge biosorption process and their environmental applications are described.

9.2 Types of Sludge in Wastewater Treatment Process 9.2.1

Activated Sludge

The AS system has been widely applied for treating various types of wastewaters by microbial metabolism under different operating conditions [15]. AS is mainly composed of bacteria, protozoa and metazoans, inorganic substances, organic substances that are not decomposed by microorganisms, and residues that are metabolized by microorganisms themselves. AS is successfully applied for treating organic matter, nitrogen, and phosphorus from wastewater through the metabolic activities of microorganisms. The first full-scale AS system was built in 1913 at the Davyhulme sewage treatment in Manchester, United Kingdom. Generally, two separate phases (aeration and sludge settlement) are the main representatives of the suspended-growth aerobic AS system. Biological nitrogen removal is consisted of aerobic nitrification and anoxic

9.2 Types of Sludge in Wastewater Treatment Process

denitrification, in which NH4 + –N is converted to NO3 − –N and further forms molecular nitrogen. Since nitrification and denitrification require aerobic and anoxic environmental conditions, two separate units are generally operated in the actual process. The principle of biological P removal is to use phosphate accumulating organisms (PAO) to release and accumulate phosphorus under anaerobic and aerobic conditions to form biological sludge rich in phosphorus. The phosphorus-enriched sludge is discharged from the AS system by sedimentation to achieve the effect of phosphorus removal from the wastewater system.

9.2.2

Granular Sludge

9.2.2.1 Anaerobic Granular Sludge

Compared with flocculent sludge, granular sludge in form of AGS and AnGS is regarded as a special case of biofilm. AnGS is based on the biogranulation technology developed in the early 1980s. It is composed of methanogens, acetogens, and hydrolyzed bacteria under high hydraulic shear in the anaerobic reactor. The sedimentation of AnGS is superior to the self-aggregate of activated sludge flocs [16]. Most of the AnGS is black or gray with a relatively regular spherical or ellipsoidal structure. The surface of mature AnGS is clear, and the diameter varies from 0.1 to 5 mm. In contrast to AGS, AnGS system exhibits the advantages of higher volume load and lower operation cost when treating high concentration organic wastewater. Liu and Tay [3] reviewed the anaerobic and aerobic biogranulation processes, certifying that many full-scale AnGS units have been operated in large-scale applications, such as in the production of alcohol, starch, soy protein, beer, and xylose and in papermaking. At present, various types of AnGS reactors including upflow anaerobic sludge blanket (UASB) reactor, expanded granular sludge bed (EGSB) reactor, and internal circulation (IC) anaerobic reactor have been extensively developed. UASB is one of the most widely used reactors in high-efficiency anaerobic reactors, which is characterized by low energy consumption, low cost and bio-energy. The UASB reactor mainly includes components such as the inlet system, water distributor, reactor body, and three-phase separator. Three-phase separator is the most important unit of the UASB reactor, which could collect the biogas generated from the reaction chamber and separate the AnGS. After the 1990s, EGSB and IC reactor were developed based on the UASB reactor. The EGSB reactor utilizes an additional effluent cycle to create a high rate of rise inside the reactor, which could also treat low concentrations of organic wastewater, such as municipal wastewater, under low temperature. The IC reactor relies on a large amount of biogas generated by the anaerobic biological process itself for sufficient circulation and mixing of the internal mixture to achieve a higher organic load. However, there still remains a certain amount of organic matter (biochemical oxygen demand [BOD] or COD) in the wastewater after anaerobic biological treatment, and aerobic biological treatment should be carried out to meet the discharge standard.

219

220

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

9.2.2.2 Aerobic Granular Sludge

In the recent years, AGS has been successfully cultivated and applied by feeding a wide variety of wastewaters containing organics, nitrogen, and phosphorus [17–20]. Aerobic granular sludge-based system is regarded as a cost-effective and environment-friendly technology once applied on a large scale in the future [4]. Laboratory-scale application: It is well known that AGS is much easier to cultivate and stably operate in the batch mode (e.g. sequencing batch reactor [SBR]) than the continuous flow mode, which may be attributed to the SBR configuration for repetition of feast and famine conditions to achieve self-immobilization process. Until now, many operational parameters, such as hydrodynamic shear force [21], organic loading rate [22], settling time [23], operational period, etc. [24], affecting the formation of AGS have been investigated by using synthetic or practical wastewaters at laboratory-scale conditions. Wei et al. [25] studied the simultaneous nitrogen and phosphorus removal in a lab-scale AGS reactor treating high-strength ammonia wastewater. The SBR was sequentially operated four cycles per day for six hours in each cycle. The influent COD concentration of the SBR changed from 600 to 2000 mg/L to increase the organic load. The NH4 + –N and TP concentrations in the influent were fixed at 200 and 15 mg/L, respectively. After continuous operation for the whole 150 days, the AS gradually formed AGS under the selective pressure of shortened settling time. The seed sludge was obtained from the aeration tank with a typical fluffy, irregular, loose morphology and brown color. After successful aerobic granulation, the diameter of mature AGS was in the range of 2−3 mm (Figure 9.1). According to the observation of scanning electron microscopy (SEM), cocci were widely observed around the granules suggesting the good effluent water treatment effect (Figure 9.1f). The mixed liquor suspended solids (MLSS) concentration and sludge volume index (SVI) value of the AGS-SBR changed to 14.52 g/L and 30.32 mL/g, respectively. Nitrogen and phosphorus treatment were simultaneously improved through increasing the ratio of influent COD/N. The total nitrogen (TN) and total phosphorus (TP) removal efficiencies of AGS reactor under ideal conditions average at 89.8 and 77.5%, respectively. Pilot-scale application: AGS has been successfully cultured using the pilot-scale SBR with low-concentration domestic wastewater and industrial wastewater. Compared with the cultivation of AGS in the laboratory, the pilot-scale cultivation faces the challenge of unstable influent wastewater quality, and this therefore increases the difficulty of aerobic granulation in real-life conditions. A pilot-scale SBR with an effective volume of 1470 L was operated in a soy protein production company for the treatment of the effluent of an IC reactor [26]. The influent COD concentration of SBR was altered by the addition of raw soy protein wastewater as an external carbon source. After 90 days of continuous operation, the MLSS and SVI values of the SBR improved at 7.02 g/L and 42.99 mL/g through shortening the settling time of the reactor. According to the morphology observation, the diameters of mature aerobic granules varied in the range of 1.2–2.0 mm. By

9.2 Types of Sludge in Wastewater Treatment Process

(a)

(b)

200 μm

200 μm

(d)

(c)

500 µm

500 µm

(e)

(f)

10 mm

Figure 9.1 Morphology observation of AGS during the granulation process: (a) seed sludge on day 1; (b) sludge on day 17; (c) sludge on day 67; (d) sludge on day 120; (e) sludge on day 134; and (f) SEM image of microstructure on day 147. Source: Wei et al. [25]; reproduced with permission from Elsevier.

increasing the COD/N ratio stepwise, the AGS reactor showed excellent COD and NH4 + –N performances with removal efficiencies over 93 and 98%, respectively. As shown in Figure 9.2, Ni et al. [27] investigated the aerobic granular sludge formation process in a pilot-scale SBR by using low-strength municipal wastewater. The SBR was operated in successive cycles of four hours with HRT of six to eight hours. After continuous operation for 300 days, the COD and NH4 + –N removal efficiencies in the pilot-scale SBR reactor average at 90 and 95%, respectively. The SBR consisted of 85% granular sludge with MLSS of 9.5 g/L. The organic loading rate (OLR) and settling time were the two major selection pressures for sludge granulation.

221

222

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

Influent 6

2

4 Effluent 3 1. Air pump

2. Influent pump

3. Mass-flow controller 4. Aerator 5. Electromagnetic valve 6. Time controller

Figure 9.2 The pilot-scale SBR reactor for cultivating aerobic granules with low-strength municipal wastewater in Zhuzhuanjing WWTP. Source: Ni et al. [27]; reproduced with permission from Elsevier.

9.3 Sludge-Based Activated Carbon for Wastewater Treatment Excess AS, a by-product from numerous WWTPs, is usually discharged from the secondary sedimentation tank as a growth of microbial metabolism during wastewater treatment. Carbonization with excess sludge as raw material could realize the sludge resource utilization [28]. Recent literatures have focused on the potential application of sludge-based activated carbons (ACs) for various pollutions sorption from aqueous media [29–32].

9.3.1

Production Method

The activation methods have important effects on the physicochemical structure and adsorption property of prepared AC. In the recent years, physical activation, chemical activation, and physical–chemical activation have been the main three methods for AC production [33]. Physical activation uses an oxygen-containing gas such as water vapor, flue gas, or a mixed gas thereof as an activator at high temperature. Chemical activation refers to a method in which a chemical (ZnCl2 , H3 PO4 , KOH, etc.) is added into the raw material and then heated under the protection of an inert

9.3 Sludge-Based Activated Carbon for Wastewater Treatment

gas for activation and carbonization. Compared with the physical activation method, chemical activation has the advantages of short time, low temperature, and easy control of activation reaction, which has become one of the main effective methods for preparing high-performance AC [34]. Generally, the physicochemical characteristics, such as surface area, pore volume, and pore size, of prepared AC from sludge are influenced by various activation parameters [35]. 9.3.1.1 ZnCl2

As an inorganic chemical activator for AC production, ZnCl2 can react with cellulose to create a rich pore structure caused by electrolytic called “swelling” in the molecular structure of cellulose, while also could change the type and amount of functional groups on the surface of adsorbent. The formation of small elemental crystallites could also promote the porous structure of the prepared AC. Shi et al. [32] reported the utilization of AnGS for AS preparation by using ZnCl2 as activator. The raw AnGS was first dried to constant weight at 105 ∘ C, and then the particle size of the dehydrate sludge (DS) was controlled at 0.5–2.0 mm. Afterward, 10 g of DS was activated with 5 mol/L ZnCl2 (25 mL) as the activator, and the ZnCl2 -loaded DS samples were pyrolyzed in a quartz tube under N2 atmosphere. The heating rate and terminal temperature were kept at 15 ∘ C/min and 650 ∘ C, respectively. After carbonization, the cool, clean (3 mol/L HCl solution), and dry carbonated product was finally used for the next step of adsorption. The pHPZC (pH of zero point charge) of AnGS-AC was low at 3.0, indicating that there were a lot of negative-charge functional groups existing on the surface of AnGS-AC. This means that the positively charged pollutants could be more easily adsorbed by AnGS-AC through the effect of electrostatic attraction. According to the result of X-ray photoelectron spectroscopy (XPS), the O-atomic percentages in the AnGS-AC and commercial activated carbon (CAC) were at 15.81 and 11.54%, respectively. However, the N-atomic percentages in CAC and AnGS-AC were 0.91 and 7.80%, respectively, suggesting AnGS-AC contained more carboxyl, hydroxyl, and amino groups than those of CAC. Since those functional groups were important binding sites in the adsorption process, AGS-AC had a much stronger sorption performance in the actual process [36, 37]. 9.3.1.2 H3 PO4

The preparation of AC by phosphoric acid (H3 PO4 ) activation is a complex physicochemical reaction process, in which the raw precursors are transformed into solid carbon with certain pore structures and surface chemical properties by heat treatment under the action of H3 PO4 . It has the advantages of less pollution, higher yield and less impact on the environment [38]. Liu et al. [39] evaluated H3 PO4 as activator for influence of carbonation process by using flocculent sludge and granular sludge as raw precursors. Flocculent sludge included aerobic activated sludge (AS) and anaerobic sludge (AnS), whereas granular sludge included AGS and AnGS. The collected four kinds of sludge were firstly dried in the oven to constant weight for 24 hours. The mass ratio of sludge to H3 PO4 solution (45%) was controlled at 1 : 2 for 10 hours. After pretreatment, the four sludges were loaded in crucibles placed in

223

224

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

Table 9.1

Comparison of the properties of prepared ACs from sludge.

Samples

S BET (m2 /g)

S mic (m2 /g)

S ext (m2 /g)

V mic (cm3 /g)

V tot (cm3 /g)

Dp (nm)

AS-AC

62.9

23.7

39.2

0.0106

0.1068

6.79

AnS-AC

198.9

107.6

91.3

0.0490

0.1284

2.58

AGS-AC

506.6

301.4

205.2

0.1382

0.3995

2.72

AnGS-AC

242.4

125.6

116.8

0.0570

0.1627

2.69

BET surface area (SBET ); external surface area (Sext ); micropore surface area (Smic ), micropore volume (V mic ), total pore volume (V tot ), and average pore diameter (Dp ). Source: Liu et al. [39].

a muffle furnace with a heating rate of 25 o C/min to 550 o C. Lastly, the carbonated products were cooled and cleaned before further use. Table 9.1 summarizes the comparison of the relative parameters, such as SBET , pore volume, pore area, and diameter (Dp ), of the prepared four kinds of carbons. It was found that AGS-AC had the largest specific surface area (506.6 m2 /g) than other ACs, followed by AnGS-AC (242.4 m2 /g), AnS-AC (198.9 m2 /g), and AS-AC (62.9 m2 /g). The values of two groups of granular sludge ACs were higher than those of flocculent sludge ACs. The average pore diameters (Dp ) of AS-AC, AnS-AC, AGS-AC, and AnGS-AC were 6.79, 2.58, 2.72, and 2.69 nm, respectively. It can be seen that sludge structures could lead to different physical and chemical properties of the synthesized ACs. According to the observation of SEM micrographs (Figure 9.3), the four kinds of carbonated products expressed different structures and sizes of pores on the carbon surface. The surface roughness of the carbonized product using granular sludge as raw material greatly increased the specific surface area, indicating that the gasification of phosphate mainly occurred inside the particle [40].

9.3.2

Treatment of Dye Wastewater

9.3.2.1 MG Sorption onto Sludge-Based ACs

Malachite green (MG) is a toxic triphenylmethane chemical that is both a dye and a bactericidal and parasitic chemical that can cause cancer [41]. MG could cause Saprolegniasis of fish body, and therefore it is forbidden to be added in the field of pollution-free aquaculture. Moreover, MG has acute and chronic toxic effects on freshwater fish, which can produce teratogenic, carcinogenic, and mutagenic effects on mammals and humans through the food chain [42]. Liu et al. [39] prepared four kinds of ACs by using sewage sludges as raw materials, namely AS, AnS, AGS, and AnGS, and evaluated their adsorption performances for MG treatment. The amounts of MG adsorbed onto ACs increased with initial MG concentration, which might be attributed to the change of pollutant concentration gradient [43]. The sorption data of MG onto AC samples were better fitted by Langmuir model with the qm values at 97.74, 114.5, 148.4, and 283.79 mg/g for AS-AC, AnS-AC, AnGS-AC, and AGS-AC, respectively. The results suggested

9.3 Sludge-Based Activated Carbon for Wastewater Treatment

(a)

(b)

4 μm

(c)

4 μm

(d)

4 μm

4 μm

Figure 9.3 SEM micrographs of carbonated products: AS-AC (a); AnS-AC (b); AGS-AC (c); AnGS-AC (d). Source: Liu et al. [39]. Reproduced from Ref. [39] with permission from the Royal Society of Chemistry.

that the granular sludge-derived ACs had better sorption capacities than flocculent sludge-derived ACs. Additionally, the four kinds of sludge-derived ACs by H3 PO4 as activator expressed much higher dye adsorption capacities than other reported adsorbents [41, 42, 44, 45]. 9.3.2.2 Mineral Acid Modification of AGS-Derived AC for MG Sorption

The sorption and separation property of AC depend on the pore structure and surface chemical property. The methods for chemical surface modification of AC include oxidation modification, reduction modification, acid–base modification, and electrochemical modification. Acid modification can improve the oxygen-containing functional groups of the AC and therefore increase the target pollutant removal ability. Zhang et al. [46] reported a novel AGS-derived AC by the ZnCl2 as the activator and further modified by two kinds of mineral acid (nitric acid [NA] and sulfuric acid [SA]). Data implied that AC-NA and AC-SA had lower surface area of 838.3 and 960 m2 /g than that of AC (1003.8 m2 /g). Moreover, the total pore volumes (V tot ) of two modified carbons (0.498 and 0.531 cm3 /g for AC-NA and AC-SA) were lower than that of AC. However, the average pore diameters (Dp ) of three kinds of carbons had no significant difference.

225

226

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

MG sorption performance was investigated to evaluate the adsorption properties of the three kinds of carbons by using batch experiments. Compared to controlled AC sample, the MG sorption efficiency of AC-NA and AC-SA had no significant change regardless of the influent pH values. The three kinds of carbons reached sorption equilibrium at 240 minutes and better fitted to the pseudo second-order kinetic model. The MG adsorption behavior of the three carbons was well described by the Langmuir model. The adsorption capacities of MG onto AC-NA, AC-SA, and AC were 303.03, 284.90, and 269.54 mg/g, respectively. Therefore, mineral acid modification exhibited more oxygenic functional groups that determined the adsorption performance. According to Fourier transform infrared spectrometry (FTIR) and zeta potential, the surface functionality played a more crucial role than pore structure during the adsorption process.

9.3.3

Treatment of Heavy Metal-Contained Wastewater

With the increasing exploitation, processing, and commercial manufacturing activities, many heavy metals have entered into soil, water, atmosphere, and organisms. Even if the concentration of heavy metal is low, it can accumulate in algae and sediments. Heavy metals could interact strongly with proteins and various enzymes in the human body, which in turn affects the body’s metabolism and disrupts the functioning of the body [47]. Heavy metals could persist, accumulate, and migrate after entering the environment or ecosystem, which is significantly necessary prior to discharging them into the receiving water. At present, heavy metal removal methods are mainly divide into chemical precipitation, dissolution, dialysis, electrolysis, reverse osmosis, distillation, adsorption, etc. [48, 49]. Adsorption is widely used because of its advantages such as simple operation, large treatment capacity, recyclable heavy metal ions, and recyclable adsorbents, especially in the treatment of low-concentration metal wastewater. As a common adsorbent, sludge-based carbon is widely used in various types of heavy metal wastewater treatments because of its low cost, resource abundance as well as green feasible. 9.3.3.1 Heavy Metal Sorption onto Sludge-Based AC

Wang et al. [50] compared sewage sludge-based AC by using H3 PO4 or ZnCl2 as the chemical activating agents. The weights of activating agent (H3 PO4 and ZnCl2 ) to raw material were controlled at 1 : 2 and 1 : 1.5, respectively. It was found that the yield of the prepared carbon by H3 PO4 was higher than that prepared by ZnCl2 . By using batch adsorption experiment, both kinds of ACs were found to have the highest Cu(II) treatment efficiency at pH 5.0. The maximum adsorption capacities obtained from the Langmuir isotherm were 7.73 and 10.56 mg/g for H3 PO4 - and ZnCl2 -activated ACs, respectively. Cu(II) desorption from H3 PO4 - or ZnCl2 -activated AC by using H2 SO4 were around 75 and 65%, respectively. Lu et al. [51] studied the main sorption mechanisms of Pb(II) onto sludge-derived AC (SDAC) and their related contributions. The pyrolysis temperature and heating rate were controlled at 550 ∘ C and 10 ∘ C/min, respectively. The sorption capacities

9.3 Sludge-Based Activated Carbon for Wastewater Treatment

Cation release

Co-precipitation Sludgederived biochar

Surface precipitation (amorphorse) Physical adsorption Functional groups complexation

Other

Ca(II)/Ca2+, (I)/K+, Mg(II)/Mg2+, Na(I)/Na+ H2O PO43–, SiO42–, CO32– Fe, Al, Mn ...

Figure 9.4 Conceptual illustration of Pb adsorption mechanism on sludge-derived AC. Source: Lu et al. [51]. Reproduced with permission from Elsevier.

of Pb(II) onto SDAC gradually increased from 16.11 to 30.88 mg/g in the pH range of 2–5. Pb(II) sorption onto SDAC increased with contact time and was well fitted by the pseudo second-order model. According to the results of various spectral characterization methods, Pb(II) sorption may mainly involve the electrostatic outer-sphere complexation, coprecipitation and inner-sphere complexation, surface complexation with free carboxyl functional groups as well as surface complexation (Figure 9.4). 9.3.3.2 Cu(II) Sorption onto AGS-AC in the Presence of HA and FA

The sorption performance of heavy metal onto sludge AC is influenced by many parameters, including adsorbent dosage, pH value, contact time, coexisting substance, and temperature [52, 53]. Effluent organic matter (EfOM) is a complex organic compound produced from biochemical effluents from secondary

227

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

sedimentation tank with the main fractions of humic acid (HA) and fulvic acid (FA) [54]. The HA and FA can interact with the surface of the adsorbent and therefore affect the physicochemical structure of the adsorbent, further affecting the adsorption effect [55]. The sorption performance of Cu(II) onto AGS prepared carbon was investigated in the absence and presence of EfOM by Wei et al [56]. The raw EfOM was collected at the secondary settling tank before the disinfection step in a WWTP treating municipal wastewater. AGS obtained from the laboratory was firstly carbonized with the addition of ZnCl2 for the preparing AGS-AC. The BET surface area of AGS-AC was measured to be 1175.1 m2 /g. According to FTIR, the main functional groups of AGS-AC were identified as –OH at 3436 cm−1 , C=O at 1631 cm−1 , and C–O–C at 1212 cm−1 . The zeta potentials of AGS-AC at pH 2.0 and 10.0 were 11.6 to −4.7 mV, respectively. It was observed that the addition of EfOM could reduce the equilibrium time of Cu(II) sorption onto AGS-AC from 300 to 60 minutes. The presence of EfOM in sorption process improved the Cu(II) removal efficiency from 57.4 to 76.8%. The calculated qm values from Langmuir isotherm model were 18.5 and 20.0 mg/g for AGS-AC in the absence and presence of EfOM, respectively. As shown in Figure 9.5, the raw EfOM sample contained two main fluorescence peaks (A and B) located at Ex/Em of 330/414 and 250/423 nm, respectively, representing humic-like and 400

400

–2.000

–2.000

360

360

60.75

60.75 123.5

320

186.3 249.0

280

311.8

Ex (nm)

Ex (nm)

123.5

320

186.3 249.0

280

311.8

374.5

240

374.5

437.3

300

350

(a)

400 450 Em (nm)

500

550

240

437.3

500.0

300

350

(b)

400 450 Em (nm)

500

550

–5.500

–5.500

360

9.938

360

9.938

25.38

320

40.81 56.25

280

71.69

240

102.6

Ex (nm)

25.38

320

40.81

280

71.69

240

102.6

56.25

87.13

87.13

300 (c)

500.0

400

400

Ex (nm)

228

350

400 450 Em (nm)

500

550

118.0

118.0

300 (d)

350

400 450 Em (nm)

500

550

Figure 9.5 Changes in 3D-EEM spectra of EfOM samples from batch biosorption processes: (a) raw EfOM; (b) raw EfOM in the presence of 10 mg/L Cu(II); (c) raw EfOM in the presence of 20 mg AGS-AC; and (d) raw EfOM in the presence of 20 mg AGS-AC and 10 mg/L Cu(II). Source: Wei et al. [56]; reproduced with permission from Elsevier.

9.4 Granular Sludge Biosorbent Applied for Wastewater Treatment

fulvic-like substances. The fluorescence intensities of both peaks were significantly reduced after the addition of Cu(II) and AGS-AC, suggesting that HA and FA could react with both adsorbent and adsorbate and that altered the sorption characteristics of Cu(II) onto AGS-AC. Adsorption of HA and FA from EfOM onto the surface of the adsorbent could increase more binding sites and negative charges of the AGS-AC, thereby increasing the metal adsorption capacity through chemical complexation and electrostatic adsorption.

9.4 Granular Sludge Biosorbent Applied for Wastewater Treatment Since AS is of low cost, freely available, and rich in functional groups, it has been successfully applied as an effective biosorbent to treat various types of wastewater [57]. However, in the actual operation process, the AS adsorbent has the disadvantages of poor solid–liquid separation effect after adsorption is completed [58]. Compared with conventional AS, granular sludge is a kind of biosorbent that is widely developed, which has the advantages of porous microbial structure, abundant adsorption sites, and excellent settling performance [26]. At present, granular sludge biosorbents have been widely applied for the treatment of toxic organics, heavy metals, and dyes [59]. EPS are distributed on the surface of sludge in a three-dimensional matrix that keeps the stable and dense structure of granular sludge [60]. Protein (PN), polysaccharides (PS), humic acid, and fulvic acid are the main components of microbial EPS, usually affecting the physicochemical properties of sludge [61]. EPSs are the main hindrance for protecting the structure of granular sludge when in the presence of toxic substances. EPSs also have a certain adsorption capacity due to their large amount of binding sites [62]. Hence, it is postulated that EPS could play a certain role to remove contaminants in the granular sludge biosorption processes. Recent literatures have reported the contribution of EPS in the biosorption of dye or heavy metals by applying AGS/AnGS as biosorbent. Moreover, the combined uses of chemical and spectroscopy analyses were investigated to elucidate the mechanism between EPS and target pollutants for better understanding the key EPS component for contribution to the target pollutant removal in biosorption systems.

9.4.1

Treatment of Dye Wastewater

9.4.1.1 Role of EPS in Aerobic Granular Sludge for MB Sorption

Wei et al. [63] studied the role and mechanism of EPS in biosorption of MB wastewater onto AGS through batch adsorption experiment. The adsorption of MB by AGS could be divided into sludge adsorption and EPS adsorption. The total MB sorption onto AGS (sludge + EPS) was firstly measured after batch adsorption experiment. Then, EPS was extracted from dye-loaded AGS for measuring MB concentration and recognized as adsorbed by EPS. The MB adsorbed by sludge was determined by subtracting the amount adsorbed by EPS from the total adsorption amount of AGS.

229

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater 400

400 Sludge + EPS Sludge EPS

200

200

100

100

0

0 0

(a)

Sludge + EPS Sludge EPS

300 qc (mg/g)

300 qc (mg/g)

230

300

600

900

Ce (mg/L)

1200

1500

0

(b)

300

600

900

1200

1500

Ce (mg/L)

Figure 9.6 Langmuir (a) and Freundlich (b) adsorption isotherms fit of MB adsorption onto AGS. Source: Wei et al. [63]; reproduced with permission from Elsevier.

By examining the sorption time, it can be found that the contact time of MB sorption onto AGS to equilibrium was 60 minutes. After adsorption equilibrium, the MB sorption onto EPS, sludge, and AGS was 9.38, 80.72 and 90.10%, respectively. It indicated that the interaction between EPS and MB had a removal effect during sorption process. Figure 9.6 shows the simulating curves of the Langmuir and Freundlich models of MB sorption onto AGS. It was observed that MB sorption onto sludge agreed well with the Langmuir isotherm model in both the absence and presence of EPS, and the correlation coefficients of MB sorption onto sludge + EPS, sludge, and EPS were 0.9825, 0.9694, and 0.9811, respectively. The calculated maximum adsorption capacity (qm ) of MB onto AGS was 389.6 mg/g, which was much higher than the previous biosorbents’ values of 158 and 16.56 mg/g reported by McKay et al. [64] and Bulut and Ayd𝚤n [65], respectively. 9.4.1.2 Biosorption of Dye Wastewater and Photocatalytic Regeneration of AGS

After sorption saturation, the used biosorbents lose the ability to further adsorb and become harmful substances, which should be properly regenerated to reduce the disposal risk. Chemical regeneration, such as using alkali or acid solution, is a common process that enables the saturated dye-loaded biosorbent to be reused. But the desorption solution should be further treated. Till now, photocatalytic oxidation has achieved extensive development in the field of activated carbon regeneration, which has the advantages of high efficiency, low energy, and time saving. By combining biosorption and photocatalytic degradation technology, the dye is first efficiently adsorbed onto the surface of the adsorbent, and subsequently photocatalytic oxidation tends to treat the concentrated dye solution for eliminating the secondary pollution under UV light [66, 67]. Huang et al. [68] reported the dye wastewater biosorption onto AGS and then the photocatalytic regeneration and degradation of AGS with acid TiO2 hydrosol. Acid TiO2 hydrosol was successfully prepared and characterized by using SEM, XRD, and FTIR for simultaneous dye-loaded biosorbent recovery and dye degradation. Two kinds of dye target pollutants, methyl orange (MO) and crystal violet (CV), were selected to evaluate the biosorbent property and photocatalytic regeneration effect. AGS biosorbent was first collected from the SBR system and cleaned to remove

9.4 Granular Sludge Biosorbent Applied for Wastewater Treatment

80 Decolorization efficiency (%)

Figure 9.7 Regeneration of AGS by using different eluants: (a) MO and (b) CV. Source: Huang et al. [68]; reproduced with permission from Springer Nature.

60

40

20

0 NaOH (a)

Acid TiO2 EDTA Hydrosol

HCI

HCI NaOH Hydrosol

EDTA

Decolorization efficiency (%)

50 40 30 40 10 0 Acid TiO2 (b)

surface ions. Through the batch adsorption result, the maximum adsorption uptake capacities (qm ) of MO and CV onto AGS were high, 741.6 and 323.6 mg/g, respectively. Therefore, it can be concluded that AGS can be considered as an efficient biosorbent and easy to separate from aqueous solution for the practical treatment of these two dyes. As shown in Figure 9.7, four kinds of eluents (i.e. acid TiO2 hydrosol, NaOH, HCl, and EDTA) were compared to evaluate the desorption effect of the dye-loaded AGS. After washing with acid TiO2 hydrosol or NaOH, the adsorbed MO could be easily eluted from AGS. However, acid TiO2 hydrosol showed better CV desorption performance than that of other desorption eluents. The elution efficiencies of MO and CV by using acid TiO2 hydrosol were achieved at 50.5 and 43.4%, respectively, suggesting that half of the dyestuffs can be desorbed from the surface of AGS biosorbent into aqueous solution and by further conducting photocatalytic degradation. The result indicated that the synthetic acidic TiO2 hydrogel had excellent desorption effects on both cationic and anionic dye-loaded adsorbents. However, the traditional eluent can only desorb the dye from the adsorbent, which could not further solve the problem of secondary pollution of the concentrated wastewater. The prepared acid TiO2 hydrogel has excellent adsorbent regeneration capabilities, and also can further reduce dye contamination by photocatalytic degradation. Then, the regenerated AGS could be easily reused in the next

231

232

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

adsorption cycling runs by centrifugation and washing (0.01 mol/L NaOH) from a hydrosol system. After separation, MO and CV dyestuffs can be easily degraded from aqueous solution at 240 and 300 minutes in the presence of irradiation, respectively, suggesting the TiO2 hydrogel had a highly efficient catalytic activity.

9.4.2

Treatment of Heavy Metal-Contained Wastewater

9.4.2.1 Zn(II) Sorption onto AGS

As a widely distributed metal element in nature, Zn(II) is generally generated and discharged from industrial activities include chemical processes, zinc mining, smelting, machinery manufacturing, galvanizing, instrumentation, and paper manufacturing [69]. In a recent research, EPS for Zn(II) binding and the relative mechanism during its sorption process onto AGS was evaluated [70]. It was found from the contact time that the sludge and EPS contribution for Zn (II) sorption was 78.88 and 1.84% at equilibrium time, respectively. Zn(II) biosorption onto AGS followed the pseudo second-order kinetic model, suggesting that chemisorption as the main driving force was the rate-limiting step in the biosorption process between sorbent and sorbate. By data fitting, Zn(II) sorbed by AGS in the presence and absence of EPS better fitted to Freundlich equation, which is a fairly satisfactory exponential isotherm that has widely been applied in batch experiments [71]. The maximum adsorption capacity (qm ) of Zn(II) onto AGS was calculated to be 64.44 mg/g, which was higher than that of other reported adsorbents [71–74]. The mechanism of adsorption of heavy metals by AGS may include the following processes: (i) positively charged heavy metals and negatively charged cell walls occuring commonly in the sorption process; (ii) cell walls or EPS containing abundant functional groups providing rich sorption sites for heavy metal sorption; (iii) the ion exchange process with light metal ions (Ca2+ , K+ , Na+ ); and (iv) chemical precipitation [75]. Moreover, AGS has the advantages of excellent separation effect and dense structure that make it more feasible from treated effluent after the sorption process. 9.4.2.2 Cu(II) Sorption onto AGS

Huang et al. [76] assessed the biosorption performance of Cu(II) onto AGS by considering the EPS binding process as one kind of sorption mechanism. At the equilibration time, Cu(II) adsorbed by EPS accounted for approximately 11.6% of the total removal efficiency, suggesting the presence of EPS on the surface of AGS had a certain complex ability for Cu(II) during biosorption process. Regardless of the presence and absence of EPS, the sorption isotherms of Cu(II) onto AGS were better fitted by the Langmuir model. The qm of Cu(II) sorption onto AGS was 123.2 mg/g, which was similar to the literature value for Cu(II) removal by Ca/Al-layered double hydroxides (CA-LDH) after graphene oxide (GO) coagulation (122.7 mg/g) [77]. Specifically, EPS sorption contributed much higher to Cu(II) removal performance (82.85%) at low Cu(II) concentration than equilibrium. The maximum adsorption capacity (qm ) value was 114.01 mg/g of Cu(II) sorption onto AGS in the absence of EPS, and EPS contributed only 7.49% of total Cu(II) removal.

9.4 Granular Sludge Biosorbent Applied for Wastewater Treatment

1000

[Cu] = 0 mg/L [Cu] = 20 mg/L [Cu] = 40 mg/L [Cu] = 60 mg/L [Cu] = 80 mg/L [Cu] = 100 mg/L [Cu] = 120 mg/L [Cu] = 140 mg/L [Cu] = 160 mg/L [Cu] = 200 mg/L

Fluorescence intensity

800

600 400

200 0 240

260

280

300

320

340

Wavelength (nm)

Figure 9.8 Changes in synchronous fluorescence spectra of released SMP as a function of contact time. Source: Huang et al. [76]; reproduced with permission from Elsevier.

Although EPSs contribute to Cu(II) removal from aqueous solution, soluble microbial products (SMPs) may be produced under the stress of heavy metal as by-products during the biosorption process, which would be a ignored secondary pollution. The released SMP was characterized by using synchronous fluorescence spectra as a function of contact time. As presented in Figure 9.8, the release of SMP showed an increasing trend with increasing exposure time. Three distinctive regions of protein-like, fulvic-like, and humic-like fluorescence fractions corresponded to the wavelengths from 250 to 300 nm, 300 to 380 nm, and 380 to 550 nm, respectively [78]. Synchronous spectra of SMP had the highest fluorescence intensity at 375 nm than the peaks at 275 and 460 nm, suggesting that the fulvic-like fluorescence fraction was more sensitive to the Cu(II) addition. 9.4.2.3 Ni(II) Sorption onto AGS/AnGS

High levels of Ni(II) are discharged from nickel mining and smelting, production and processing of alloy steel, coal and oil burning, electroplating and nickel plating production process, which may mainly reduce the fertility of human, and also has the effect of teratogenesis and mutagenesis [79, 80]. Thereby, the maximum amount of Ni(II) in drinking water of World Health Organization (WHO) is 0.1 mg/L [81]. Moreover, the existence and accumulation of Ni(II) can affect the efficiency of biological removal efficiency in wastewater treatment. Li et al. [13] comparatively evaluated the roles of EPS in biosorption process of Ni(II) onto AGS and AnGS by using batch experiment. AGS and AnGS with average size of 2 mm were collected from lab-scale SBR and full-scale UASB, respectively. The sorption data of Ni(II) were analyzed by using Langmuir, Freundlich, and Fritz–Schlunder models, as shown in Figure 9.9. Ni(II) sorption onto AGS and AnGS was better fitted to Langmuir and Freundlich isotherm models, respectively. It was found from Langmuir model that AGS had a higher Ni(II) adsorption capacity (65.77 mg/g) than that of AnGS (54.18 mg/g).

233

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

60

50 40

40 30 Langmuir of sludge with EPS Freundlich of sludge with EPS Fritz-schlunder of sludge with EPS Langmuir of sludge without EPS Freundlich of sludge without EPS Fritz-schlunder of sludge without EPS

20 10 0 0

(a)

200

400 600 ce(mg/L)

800

20

25 20 15 10 5 25 50 75 100 150 250 500 600 700 C0(mg/L)

Langmuir of sludge with EPS Freundlich of sludge with EPS Fritz-schlunder of sludge with EPS Langmuir of sludge without EPS Freundlich of sludge without EPS Fritz-schlunder of sludge without EPS

10 0 0

200

400

600

800

ce(mg/L)

(b)

EPS sorption contribution (%)

30

(c)

30

1000

35

0

qe (mg/g)

qe (mg/g)

50

EPS sorption contribution (%)

234

35 30 25 20 15 10 5 0

(d)

25 50 75 100 150 250 500 600 700 C0(mg/L)

Figure 9.9 Sorption isotherms fit of Ni(II) onto AGS (a) and AnGS (b) with and without EPS; contribution of EPS for AGS (c) and AnGS (d) on Ni(II) removal efficiency. Source: Li et al. [13]; reproduced with permission from Elsevier.

The interaction mechanism between EPS and Ni(II) was studied by using a combined spectra approach. The PS and PN contents were reduced in the presence of Ni(II) in both AGS and AnGS. Through the observation of 3D-EEM, there were two and three fluorescence peaks identified in the raw EPS from AGS and AnGS, which were assigned to PN-like peaks and humic-like fluorescence peaks [63, 82, 83]. Their fluorescence intensities significantly decreased with different degrees after the Ni(II) migration during the biosorption process. Through the result of synchronous fluorescence spectra, the fluorescence intensities of two kinds of EPS were obviously reduced with increasing Ni(II) concentration, and the quenched types of AGS and AnGS in the presence of Ni(II) belonged to dynamic and static quenching, respectively. Two-dimensional correlation spectroscopy (2D-COS) further descripted that the fluorescence changes took place with the increased Cu(II) concentration in the humic-like region > PN-like regions in AGS and fulvic-like fraction > PN-like and humic-like fractions in AnGS. It was observed from FTIR spectra that hydroxyl band of PS and amide-I bands of PN from two kinds of EPS were the key functional groups that caused Ni(II) biosorption. 9.4.2.4 Magnetic Modification of AnGS for Pb(II) and Cu(II) Removal

It is noteworthy that the powder of adsorbent materials is difficult to separate from treated wastewater due to its small particle size in sorption process. In order to solve

9.4 Granular Sludge Biosorbent Applied for Wastewater Treatment

Fe3O4

Magnet

Chitosan(CS)

Magnetic separatation

Magnetic chitosan (M-CS)

AnGS

Adsorption Magnetic chitosan/anaerobic granular sludge (M-CS-AnGS)

Figure 9.10 Schematic illustration of the synthesis of M-CS-AnGS and its application for the removal of Pb(II) and Cu(II). Source: Liu et al. [11]; reproduced with permission from Elsevier.

the problem, magnetic nano adsorbents have been widely focused on to achieve rapid separation under the condition of external magnetic field. A novel magnetic chitosan/anaerobic granular sludge (M-CS-AnGS) composite was synthesized to enhance the metal sorption performance and separation effect [11]. Chitosan (CS), as a product of natural high polymer chitin deacetylation, contains abundant hydroxyl and amino groups in the molecular chain, which enables it to form stable chelates with most metal ions. It has the advantages of wide raw material, low price, no secondary pollution, and being easily degradable. By using external magnetic field, the as-prepared M-CS-AnGS composite can be separated easily from the sorption system after Pb(II) and Cu(II) sorption, as depicted in Figure 9.10. It was found that the addition of CS and Fe3 O4 can effectively improve the adsorption capacities of AnGS. Through batch experiments, the optimum pH values were selected at 6.0 and 3.0 for Pb(II) and Cu(II) removal, respectively. The adsorption kinetics of Pb(II) and Cu(II) onto M-CS-AnGS were better fitted with pseudo second-order kinetic model and Langmuir isotherm with the maximum adsorption capacities of 97.97 and 83.65 mg/g, respectively. Surface complexation and electrostatic attraction were the main adsorption mechanisms for heavy metal removal by M-CS-AnGS according to the observation of FTIR, zeta potential, and XPS analysis. In addition, two kinds of heavy metals could be effectively eluted from M-CS-AnGS by using Na2 EDTA (0.1 mol/L). The regenerated M-CS-AnGS for Pb(II) and Cu(II) over four cycles was 94.81 and 92.58%, respectively.

9.4.3

Treatment of Multicomponent Contaminants

At present, industrial wastewater contains a large amount of coexisting pollutants rather than a single pollutant, such as the presence of inorganic and organic toxic

235

236

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

compounds [84]. These hazardous pollutants exist or interact with a variety of chemical states, resulting in a comprehensive impact on the environment and ecology [85]. Wastewater normally contains more than a single adsorbate in solution and diverse approaches, including biosorption, membrane filtration, and electro-chemical strategies, have been employed to treat these multicomponent practical situations [86]. However, the practical wastewater containing multiple components (i.e. heavy metals–dyes) has the characteristics of wide influence, high pollution intensity, and being highly difficult to treat. Shi et al. [87] applied AnGS for competitive biosorption of MB and Pb(II), and also optimized the experimental conditions by using response surface methodology (RSM). It was found that the coexisted of MB declined the Pb(II) sorption amount in the binary system as a function of contact time. The coexisted ions reduced the adsorption capacity of granular sludge for pollutants in competitive biosorption systems. The optimized conditions for multicomponent treatment from RSM were 338.38 mg/L Pb(II), 340.65 mg/L MB, and contact time of 122.61 minutes. Sun et al. [88] studied the competitive biosorption removal of Zn(II) and Co(II) by AGS in binary-metal system. The sorption capacities of Co(II) and Zn(II) onto AGS were 55.25 and 62.50 mg/g in single-metal system, whereas they were 54.05 and 56.50 mg/g in binary-metal system. It was found from FTIR and XPS that alcohols and carboxylates were the main binding sites for two kinds of heavy metals.

9.5 Applications of EPS Extracted from Sludge for Wastewater Treatment 9.5.1

Bioflocculant

EPSs are secreted by microorganisms as a general metabolism product of microbial aggregates in the formation of AGS or AnGS [60]. Most notably, the organic components of the extracellular polymer can change the surface characteristics and the physical property of granular sludge. EPS is commonly divided into soluble EPS (S-EPS) and bound EPS (B-EPS), and B-EPS generally has a two-layer structure. The outer and inner layers of B-EPS are loosely bound EPS (LB-EPS) and tightly bound EPS (TB-EPS) [89]. Moreover, many environmental factors and system parameters may change the main components and structure of the two kinds of EPS. EPS from sludge can be applied as an underlying bioflocculant due to the main components of PN and PS. As reported by Sun et al. [90], the bioflocculant was prepared by using hydrochloric acid from excess sludge, and 99.5% of flocculating rate was achieved for 4 g/L kaolin clay when the flocculant concentration was 3.0% (v/v). Moreover, PS occupied the main proportion of purified sludge bioflocculant. According to the performance test, the sludge-based bioflocculant showed moderate thermostability. Liu et al. [91] investigated the bioflocculant from backwashing sludge, and found that the bioflocculant had lost its activity at pH 11.0. In contrast, bioflocculants had higher flocculating activity when the extracted pH of sludge was 5.0. At the optimum condition, kaolin clay achieved the best flocculating rate of 92.67%. In another report, amino groups and hydroxyl groups existed in the

9.5 Applications of EPS Extracted from Sludge for Wastewater Treatment

molecules of bioflocculant prepared from excess sludge by using XPS and FTIR. The sludge bioflocculant first caused the suspension instability of kaolin through charge neutralization, and then promoted the aggregation of suspended particles as the main flocculation mechanism [92].

9.5.2

Biosorbent for the Removal of Various Pollutants

Heavy metal sorption: EPSs have been obtained by different extraction methods from various sources, such as pure bacterial strains, sludge, bacteria, algae, and fungi [93, 94]. In the recent years, EPSs have been commonly applied as an efficient biosorbent for treating wastewater containing heavy metal since their abundant binding sites [95]. The influences of sorption conditions have been comprehensively evaluated for the interactions between EPS and heavy metal. At present, there are also some related mechanisms during metal sorption onto EPS, such as ion-exchange, complexation and surface precipitation. Guibaud et al. [96] applied EPS from anaerobic granular biofilms by cationic exchange resin, and found that EPS had higher sorption ability for Pb (II) than Cd (II) and this changed with increase in pH. Wen et al. [97] reported that the adsorption capacities of Pb(II), Cd(II), and Zn(II) onto EPS from AGS were high: 1587.3, 1470.6, and 1123.6 mg/g respectively. Zhang et al. [98] extracted EPS from waste sludge after short-time aerobic digestion as biosorbent for Cu(II) treatment, finding that Cu(II) sorption onto EPS was mainly caused by physical and chemical interaction, with especially electrostatic attraction being the most important factor. Moreover, extensive analytical methods have been developed in the recent years to confirm the binding mechanism between heavy metal and EPS, including FTIR, fluorescence spectroscopy, and SEM. Sun et al. [88] applied XPS and FTIR to study the interaction mechanism of Zn (II) and Co (II) with reactive chemical groups on EPS, finding that alcohol, carboxyl, and amino groups were the main sites for metal binding in the EPS sorption process. Yin et al. [99] found that ion exchange had a certain contribution in the EPS adsorption process of Cu(II) and Cd (II) by using energy dispersive X-ray (EDX). Fluorescence spectroscopy has the advantages of high sensitivity, being nondestructive, and selectivity, and is an important method for studying the interaction between heavy metal and EPS in the fields of chemistry, environment, and biochemistry. Sheng et al. [100] studied the interaction between Cu(II) and EPS via 3D-EEM, obtaining information on excitation spectrum, emission spectrum, fluorescence intensity, and fluorescence quenching of EPS trend with the increase of Cu(II) concentration. Zhang et al. [69] investigated the complexation between Hg(II) and natural biofilm EPS via fluorescence quenching titration, finding that two protein-like fluorescence peaks were observed in the raw EPS and their intensities were strongly influenced by pH and Hg(II) concentration. Organic pollutant sorption: Organic pollutants, such as antibiotics, pesticides, chlorophenol, and dyes can be effectively adsorbed by EPS. Gao et al. [101] investigated the biosorption process of four kinds of dyes onto aerobic granules biosorbent, in which LB-EPS, TB-EPS, and residual sludge (the sludge left after EPS extraction) contributed to the dye removal efficiency. It was found that the functional groups of

237

238

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

EPS and dyes’ chemical structures played the main role in the biosorption mechanisms. For reactive brilliant blue KN-R (KN-R) biosorption, the binding sites were provided by lipid and phosphate groups. Amino acids, amine, carboxyl, phosphate, and lipid played an important role in Congo red (CR) adsorption. The lipid fractions were responsible for the binding of reactive brilliant red K-2G (RBR). The biosorption of MG was mostly contributed by phosphate groups. Wei et al. [102] evaluated the interaction mechanism between 4-CP and EPS from AGS via spectroscopic approaches. Protein-like and visible humic-like substances were the main identified peaks from raw EPS by EEM, and their intensities reduced from 1013.01 and 272.37 to 640.25 and 216.58 a.u. in the presence of 4-CP, respectively. It was found from synchronous fluorescence spectra that tryptophan (Trp) was the main site for EPS fluorescence quenching. According to the calculation of Stern– Volmer equation, the type of quenching between EPS and 4-CP was static quenching rather than dynamic quenching with the formation constant (K A ) of 0.07 × 104 L/mol.

9.6 Conclusion In summary, the amount of sludge produced by sewage treatment is increasing rapidly. The proper disposal and resource utilization of sewage sludge has been paid much attention around the world. In the recent decades, the application of sludge-based carbon biosorbent has been widely investigated in view of operational parameters such as pretreatment method, pH value, pollutant concentrations, and other coexisting ions. The result proves that sludge-based biosorption is a promising, stable, and easy operation for treating various pollutants from wastewater. In comparison with bioflocs, granular sludge-based biosorbent has better adsorption performance and separation effect in real wastewater treatment. The findings are significant to reveal the resource utilization of sewage sludge as biosorbent, and provide useful information on understanding the interaction mechanism between biosorbent and pollutants. Further investigation is needed to apply the sludge-based biosorbent in full-scale wastewater treatment systems.

References 1 Dong, H., Zhang, K., Han, X. et al. (2017). Achievement, performance and characteristics of microbial products in a partial nitrification sequencing batch reactor as a pretreatment for anaerobic ammonium oxidation. Chemosphere 183: 212–218. 2 Liu, Z., Liang, Z., Wu, S., and Liu, F. (2015). Treatment of municipal wastewater by a magnetic activated sludge device. Desalination and Water Treatment 53 (4): 909–918. 3 Liu, Y. and Tay, J.H. (2004). State of the art of biogranulation technology for wastewater treatment. Biotechnology Advances 22 (7): 533–563.

References

4 De Bruin, L.M.M., De Kreuk, M.K., Van Der Roest, H.F.R. et al. (2004). Aerobic granular sludge technology: an alternative to activated sludge? Water Science and Technology 49: 1–7. 5 Morgenroth, E., Sherden, T., Van Loosdrecht, M. et al. (1997). Aerobic granular sludge in a sequencing batch reactor. Water Research 31: 3191–3194. 6 Schmidt, J.E. and Ahring, B.K. (1996). Granular sludge formation in upflow anaerobic sludge blanket (UASB) reactors. Biotechnology and Bioengineering 49: 229–246. 7 Wei, Y., Van Houten, R.T., Borger, A.R. et al. (2003). Minimization of excess sludge production for biological wastewater treatment. Water Research 37: 4453–4467. 8 Chen, X., Jeyaseelan, S., and Graham, N. (2002). Physical and chemical properties study of the activated carbon made from sewage sludge. Waste Management 22: 755–760. 9 Aksu, Z. and Gönen, F. (2004). Biosorption of phenol by immobilized activated sludge in a continuous packed bed: prediction of breakthrough curves. Process Biochemistry 39: 599–613. 10 Rafatullah, M., Sulaiman, O., Hashim, R. et al. (2010). Adsorption of methylene blue on low-cost adsorbents: a review. Journal of Hazardous Materials 177: 70–80. 11 Liu, T., Han, X., Wang, Y. et al. (2017). Magnetic chitosan/anaerobic granular sludge composite: synthesis, characterization and application in heavy metal ions removal. Journal of Colloid and Interface Science 508: 405–414. 12 Liu, Y.Q., Liu, Y., and Tay, J.H. (2004). The effects of extracellular polymeric substances on the formation and stability of biogranules. Applied Microbiology and Biotechnology 65: 143–148. 13 Li, N., Wei, D., Wang, S. et al. (2017). Comparative study of the role of extracellular polymeric substances in biosorption of Ni (II) onto aerobic/anaerobic granular sludge. Journal of Colloid and Interface Science 490: 754–761. 14 More, T.T., Yadav, J.S.S., Yan, S. et al. (2014). Extracellular polymeric substances of bacteria and their potential environmental applications. Journal of Environmental Management 144: 1–25. 15 Chan, Y.J., Chong, M.F., Law, C.L. et al. (2009). A review on anaerobic–aerobic treatment of industrial and municipal wastewater. Chemical Engineering Journal 155: 1–18. 16 Yan, Q., Yu, D., Wang, Z. et al. (2008). Phenol inhibition and restoration of the bioactivity of anaerobic granular sludge. Applied Biochemistry and Biotechnology 150: 259–265. 17 Adav, S.S., Chen, M.Y., Lee, D.J. et al. (2007). Degradation of phenol by aerobic granules and isolated yeast Candida tropicalis. Biotechnology and Bioengineering 96 (5): 844–852. 18 Cassidy, D.P. and Belia, E. (2005). Nitrogen and phosphorus removal from an abattoir wastewater in a SBR with aerobic granular sludge. Water Research 39: 4817–4823.

239

240

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

19 Qin, L. and Liu, Y. (2006). Aerobic granulation for organic carbon and nitrogen removal in alternating aerobic–anaerobic sequencing batch reactor. Chemosphere 63: 926–933. 20 Xu, H., Tay, J.H., Foo, S.K. et al. (2004). Removal of dissolved copper (II) and zinc (II) by aerobic granular sludge. Water Science and Technology 50: 155–160. 21 Tay, J.H., Liu, Q.S., and Liu, Y. (2002). Aerobic granulation in sequential sludge blanket reactor. Water Science and Technology 46: 13–18. 22 Kim, I.S., Kim, S.M., and Jang, A. (2008). Characterization of aerobic granules by microbial density at different COD loading rates. Bioresource Technology 99: 18–25. 23 Qin, L., Liu, Y., and Tay, J.H. (2004). Effect of settling time on aerobic granulation in sequencing batch reactor. Biochemical Engineering Journal 21: 47–52. 24 Wan, J., Bessière, Y., and Spérandio, M. (2009). Alternating anoxic feast/aerobic famine condition for improving granular sludge formation in sequencing batch airlift reactor at reduced aeration rate. Water Research 43: 5097–5108. 25 Wei, D., Shi, L., Yan, T. et al. (2014). Aerobic granules formation and simultaneous nitrogen and phosphorus removal treating high strength ammonia wastewater in sequencing batch reactor. Bioresource Technology 171: 211–216. 26 Wei, D., Qiao, Z., Zhang, Y. et al. (2013). Effect of COD/N ratio on cultivation of aerobic granular sludge in a pilot-scale sequencing batch reactor. Applied Microbiology and Biotechnology 97: 1745–1753. 27 Ni, B.J., Xie, W.M., Liu, S.G. et al. (2009). Granulation of activated sludge in a pilot-scale sequencing batch reactor for the treatment of low-strength municipal wastewater. Water Research 43: 751–761. 28 Li, W., Yue, Q., Tu, P. et al. (2011). Adsorption characteristics of dyes in columns of activated carbon prepared from paper mill sewage sludge. Chemical Engineering Journal 178: 197–203. 29 Monsalvo, V.M., Mohedano, A.F., and Rodriguez, J. (2011). Activated carbons from sewage sludge: application to aqueous-phase adsorption of 4-chlorophenol. Desalination 277: 377–382. 30 Ong, S.A., Uchiyama, K., Inadama, D. et al. (2010). Treatment of azo dye Acid Orange 7 containing wastewater using up-flow constructed wetland with and without supplementary aeration. Bioresource Technology 101 (23): 9049–9057. 31 Rozada, F., Otero, M., Moran, A. et al. (2005). Activated carbons from sewage sludge and discarded tyres: production and optimization. Journal of Hazardous Materials 124: 181–191. 32 Shi, L., Zhang, G., Wei, D. et al. (2014). Preparation and utilization of anaerobic granular sludge-based biochar for the adsorption of methylene blue from aqueous solutions. Journal of Molecular Liquids 198: 334–340. 33 Nor, N.M., Lau, L.C., Lee, K.T. et al. (2013). Synthesis of activated carbon from lignocellulosic biomass and its applications in air pollution control – a review. Journal of Environmental Chemical Engineering 1: 658–666. ´ B.V., Jovanovic, ´ V.M., Stevanovic, ´ S.I. et al. (2014). Characteriza34 Kaludjerovic, tion of nanoporous carbon fibrous materials obtained by chemical activation of plane tree seed under ultrasonic irradiation. Ultrasonics Sonochemistry 21: 782–789.

References

35 Hadi, P., Barford, J., and McKay, G. (2013). Toxic heavy metal capture using a novel electronic waste-based material mechanism, modeling and comparison. Environmental Science & Technology 47: 8248–8255. 36 Comte, S., Guibaud, G., and Baudu, M. (2006). Relations between extraction protocols for activated sludge extracellular polymeric substances (EPS) and complexation properties of Pb and Cd with EPS: part II. Consequences of EPS extraction methods on Pb2+ and Cd2+ complexation. Enzyme and Microbial Technology 38: 246–252. 37 Guibaud, G., Comte, S., Bordas, F. et al. (2005). Comparison of the complexation potential of extracellular polymeric substances (EPS), extracted from activated sludges and produced by pure bacteria strains, for cadmium, lead and nickel. Chemosphere 59: 629–638. 38 Liu, H., Wang, X., Zhai, G. et al. (2012). Preparation of activated carbon from lotus stalks with the mixture of phosphoric acid and pentaerythritol impregnation and its application for Ni(II) sorption. Chemical Engineering Journal 209: 155–162. 39 Liu, T., Wei, D., Zhang, G. et al. (2016). A comparison of the influence of flocculent and granular structure of sludge on activated carbon: preparation, characterization and application. RSC Advances 6 (90): 87353–87361. 40 Teng, H., Yeh, T.S., and Hsu, L.Y. (1998). Preparation of activated carbon from bituminous coal with phosphoric acid activation. Carbon 36 (9): 1387–1395. 41 Nethaji, S., Sivasamy, A., Thennarasu, G. et al. (2010). Adsorption of Malachite Green dye onto activated carbon derived from Borassus aethiopum flower biomass. Journal of Hazardous Materials 181: 271–280. 42 Hameed, B.H. and El-Khaiary, M.I. (2008). Malachite green adsorption by rattan sawdust: isotherm, kinetic and mechanism modeling. Journal of Hazardous Materials 159: 574–579. 43 Chowdhury, S., Mishra, R., Saha, P. et al. (2011). Adsorption thermodynamics, kinetics and isosteric heat of adsorption of malachite green onto chemically modified rice husk. Desalination 265: 159–168. 44 Leng, L., Yuan, X., Huang, H. et al. (2015). Bio-char derived from sewage sludge by liquefaction: characterization and application for dye adsorption. Applied Surface Science 346: 223–231. 45 Zhang, J., Li, Y., Zhang, C., and Jing, Y. (2008). Adsorption of malachite green from aqueous solution onto carbon prepared from Arundo donax root. Journal of Hazardous Materials 150 (3): 774–782. 46 Zhang, G., Shi, L., Zhang, Y. et al. (2015). Aerobic granular sludge-derived activated carbon: mineral acid modification and superior dye adsorption capacity. RSC Advances 5 (32): 25279–25286. 47 Wang, J. and Chen, C. (2009). Biosorbents for heavy metals removal and their future. Biotechnology Advances 27 (2): 195–226. 48 Fang, X., Li, J., Li, X. et al. (2017). Internal pore decoration with polydopamine nanoparticle on polymeric ultrafiltration membrane for enhanced heavy metal removal. Chemical Engineering Journal 314: 38–49.

241

242

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

49 Hua, M., Zhang, S., Pan, B. et al. (2012). Heavy metal removal from water/wastewater by nanosized metal oxides: a review. Journal of Hazardous Materials 211: 317–331. 50 Wang, X., Liang, X., Wang, Y. et al. (2011). Adsorption of copper (II) onto activated carbons from sewage sludge by microwave-induced phosphoric acid and zinc chloride activation. Desalination 278 (1–3): 231–237. 51 Lu, H., Zhang, W., Yang, Y. et al. (2012). Relative distribution of Pb2+ sorption mechanisms by sludge-derived biochar. Water Research 46: 854–862. 52 Inyang, M.I., Gao, B., Yao, Y. et al. (2016). A review of biochar as a low-cost adsorbent for aqueous heavy metal removal. Critical Reviews in Environmental Science and Technology 46: 406–433. 53 Qian, L., Zhang, W., Yan, J. et al. (2016). Effective removal of heavy metal by biochar colloids under different pyrolysis temperatures. Bioresource Technology 206: 217–224. 54 Shon, H., Vigneswaran, S., and Snyder, S.A. (2006). Effluent organic matter (EfOM) in wastewater: constituents, effects, and treatment. Critical Reviews in Environmental Science and Technology 36: 327–374. 55 Yang, K. and Xing, B. (2009). Sorption of phenanthrene by humic acid-coated nanosized TiO2 and ZnO. Environmental Science & Technology 43 (6): 1845–1851. 56 Wei, D., Ngo, H.H., Guo, W. et al. (2018). Biosorption performance evaluation of heavy metal onto aerobic granular sludge-derived biochar in the presence of effluent organic matter via batch and fluorescence approaches. Bioresource Technology 249: 410–416. 57 Göbel, A., Thomsen, A., McArdell, C.S. et al. (2005). Occurrence and sorption behavior of sulfonamides, macrolides, and trimethoprim in activated sludge treatment. Environmental Science & Technology 39 (11): 3981–3989. 58 Gao, J., Zhang, Q., Su, K. et al. (2010). Biosorption of Acid Yellow 17 from aqueous solution by non-living aerobic granular sludge. Journal of Hazardous Materials 174: 215–225. 59 Adav, S.S., Lee, D.J., Show, K.Y. et al. (2008). Aerobic granular sludge: recent advances. Biotechnology Advances 26: 411–423. 60 Sheng, G.P., Yu, H.Q., and Li, X.Y. (2010). Extracellular polymeric substances (EPS) of microbial aggregates in biological wastewater treatment systems: a review. Biotechnology Advances 28 (6): 882–894. 61 Guo, W., Ngo, H.H., and Li, J. (2012). A mini–review on membrane fouling. Bioresource Technology 122: 27–34. 62 Liu, Y., Lam, M., and Fang, H. (2001). Adsorption of heavy metals by EPS of activated sludge. Water Science and Technology 43 (6): 59–66. 63 Wei, D., Wang, B., Ngo, H.H. et al. (2015). Role of extracellular polymeric substances in biosorption of dye wastewater using aerobic granular sludge. Bioresource Technology 185: 14–20. 64 McKay, G., Porter, J., and Prasad, G. (1999). The removal of dye colours from aqueous solutions by adsorption on low–cost materials. Water, Air, and Soil Pollution 114: 423–438.

References

65 Bulut, Y. and Ayd𝚤n, H. (2006). A kinetics and thermodynamics study of methylene blue adsorption on wheat shells. Desalination 194: 259–267. 66 Chan, H.W., Chan, C.M., Ang, P.O. et al. (2007). Integrated biosorption and photocatalytic oxidation treatment of di(2–ethylhexyl)phthalate. Journal of Applied Phycology 19: 745–753. 67 Dhodapkar, R., Rao, N.N., Pande, S.P. et al. (2007). Adsorption of cationic dyes on jalshakti, super absorbent polymer and photocatalytic regeneration of the adsorbent. Reactive & Functional Polymers 67: 540–548. 68 Huang, X., Wei, D., Yan, L. et al. (2018). High-efficient biosorption of dye wastewater onto aerobic granular sludge and photocatalytic regeneration of biosorbent by acid TiO2 hydrosol. Environmental Science and Pollution Research 25: 27606–27613. 69 Zhang, D., Pan, X., Mostofa, K.M. et al. (2010). Complexation between Hg (II) and biofilm extracellular polymeric substances: an application of fluorescence spectroscopy. Journal of Hazardous Materials 175: 359–365. 70 Wei, D., Li, M., Wang, X. et al. (2016). Extracellular polymeric substances for Zn (II) binding during its sorption process onto aerobic granular sludge. Journal of Hazardous Materials 301: 407–415. 71 Pehlivan, E., Cetin, S., and Yan𝚤k, B. (2006). Equilibrium studies for the sorption of zinc and copper from aqueous solutions using sugar beet pulp and fly ash. Journal of Hazardous Materials 135: 193–199. 72 Bhattacharya, A., Mandal, S., and Das, S. (2006). Adsorption of Zn (II) from aqueous solution by using different adsorbents. Chemical Engineering Journal 123: 43–51. 73 Kandah, M.I. (2004). Zinc and cadmium adsorption on low-grade phosphate. Separation and Purification Technology 35 (1): 61–70. 74 Tunali, S. and Akar, T. (2006). Zn (II) biosorption properties of Botrytis cinerea biomass. Journal of Hazardous Materials 131: 137–145. 75 Xu, H. and Liu, Y. (2008). Mechanisms of Cd2+ , Cu2+ and Ni2+ biosorption by aerobic granules. Separation and Purification Technology 58: 400–411. 76 Huang, L., Li, M., Si, G. et al. (2018). Assessment of microbial products in the biosorption process of Cu(II) onto aerobic granular sludge: extracellular polymeric substances contribution and soluble microbial products release. Journal of Colloid and Interface Science 527: 87–94. 77 Yao, W., Wang, J., Wang, P. et al. (2017). Synergistic coagulation of GO and secondary adsorption of heavy metal ions on Ca/Al layered double hydroxides. Environmental Pollution 229: 827–836. 78 Chen, W., Habibul, N., Liu, X.Y. et al. (2015). FTIR and synchronous fluorescence heterospectral two–dimensional correlation analyses on the binding characteristics of copper onto dissolved organic matter. Environmental Science & Technology 49: 2052–2058. 79 Borba, C.E., Guirardello, R., Silva, E.A. et al. (2006). Removal of nickel(II) ions from aqueous solution by biosorption in a fixed bed column: experimental and theoretical breakthrough curves. Biochemcal Engineering Journal 30: 184–191.

243

244

9 Resource Utilization of Sludge and Its Potential Environmental Applications for Wastewater

80 Lata, H., Garg, V.K., and Gupta, R.K. (2008). Sequestration of nickel from aqueous solution onto activated carbon prepared from Parthenium hysterophorus L. Journal of Hazardous Materials 157: 503–509. 81 Sun, Y., Yue, Q., Gao, B. et al. (2014). Adsorption and cosorption of ciprofloxacin and Ni(II) on activated carbon-mechanism study. Journal of the Taiwan Institute of Chemical Engineers 45: 681–688. 82 Chen, W., Westerhoff, P., Leenheer, J.A. et al. (2003). Fluorescence excitation−emission matrix regional integration to quantify spectra for dissolved organic matter. Environmental Science & Technology 37: 5701–5710. 83 Yamashita, Y. and Tanoue, E. (2003). Chemical characterization of protein-like fluorophores in DOM in relation to aromatic amino acids. Marine Chemistry 82: 255–271. 84 Körbahti, B.K., Artut, K., Geçgel, C. et al. (2011). Electrochemical decolorization of textile dyes and removal of metal ions from textile dye and metal ion binary mixtures. Chemical Engineering Journal 173: 677–688. 85 Fu, F. and Wang, Q. (2011). Removal of heavy metal ions from wastewaters: a review. Journal of Environmental Management 92: 407–418. 86 Wang, S. and Ariyanto, E. (2007). Competitive adsorption of malachite green and Pb ions on natural zeolite. Journal of Colloid and Interface Science 314: 25–31. 87 Shi, L., Wei, D., Ngo, H.H. et al. (2015). Application of anaerobic granular sludge for competitive biosorption of methylene blue and Pb (II): fluorescence and response surface methodology. Bioresource Technology 194: 297–304. 88 Sun, X.F., Wang, S.G., Zhang, X.M. et al. (2009). Spectroscopic study of Zn2+ and Co2+ binding to extracellular polymeric substances (EPS) from aerobic granules. Journal of Colloid and Interface Science 335: 11–17. 89 Li, X.Y. and Yang, S. (2007). Influence of loosely bound extracellular polymeric substances (EPS) on the flocculation, sedimentation and dewaterability of activated sludge. Water Research 41: 1022–1030. 90 Sun, J., Zhang, X., Miao, X. et al. (2012). Preparation and characteristics of bioflocculants from excess biological sludge. Bioresource Technology 126: 362–366. 91 Liu, W., Yuan, H., Yang, J. et al. (2009). Characterization of bioflocculants from biologically aerated filter backwashed sludge and its application in dying wastewater treatment. Bioresource Technology 100: 2629–2632. 92 Zhang, X., Sun, J., Liu, X. et al. (2013). Production and flocculating performance of sludge bioflocculant from biological sludge. Bioresource Technology 146: 51–56. 93 Comte, S., Guibaud, G., and Baudu, M. (2006). Biosorption properties of extracellular polymeric substances (EPS) resulting from activated sludge according to their type: soluble or bound. Process Biochemistry 41: 815–823. 94 Kenney, J.P. and Fein, J.B. (2011). Importance of extracellular polysaccharides on proton and Cd binding to bacterial biomass: a comparative study. Chemical Geology 286: 109–117.

References

95 Zhang, D., Wang, J., and Pan, X. (2006). Cadmium sorption by EPSs produced by anaerobic sludge under sulfate-reducing conditions. Journal of Hazardous Materials 138: 589–593. 96 Guibaud, G., Bhatia, D., d’Abzac, P. et al. (2012). Cd (II) and Pb (II) sorption by extracellular polymeric substances (EPS) extracted from anaerobic granular biofilms: evidence of a pH sorption-edge. Journal of the Taiwan Institute of Chemical Engineers 43: 444–449. 97 Wen, L., Zhang, J., Jin, Y. et al. (2015). Adsorption of Pb(II), Cd(II) and Zn(II) by extracellular polymeric substances extracted from aerobic granular sludge: efficiency of protein. Journal of Environmental Chemical Engineering 3: 1223–1232. 98 Zhang, Z., Zhou, Y., Zhang, J. et al. (2014). Copper (II) adsorption by the extracellular polymeric substance extracted from waste activated sludge after short-time aerobic digestion. Environmental Science and Pollution Research 21: 2132–2140. 99 Yin, Y., Hu, Y., and Xiong, F. (2011). Sorption of Cu (II) and Cd (II) by extracellular polymeric substances (EPS) from Aspergillus fumigatus. International Biodeterioration & Biodegradation 65: 1012–1018. 100 Sheng, G.P., Xu, J., Luo, H.W. et al. (2013). Thermodynamic analysis on the binding of heavy metals onto extracellular polymeric substances (EPS) of activated sludge. Water Research 47: 607–614. 101 Gao, J.F., Zhang, Q., Wang, J.H. et al. (2011). Contributions of functional groups and extracellular polymeric substances on the biosorption of dyes by aerobic granules. Bioresource Technology 102: 805–813. 102 Wei, D., Wang, Y., Wang, X. et al. (2015). Toxicity assessment of 4-chlorophenol to aerobic granular sludge and its interaction with extracellular polymeric substances. Journal of Hazardous Materials 289: 101–107.

245

247

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery Mian Hu 1,2 , Dabin Guo 1,2 , Yingqun Ma 1 , and Yu Liu 1,3 1 Advanced Environmental Biotechnology Centre, Nanyang Environment & Water Research Institute, Nanyang Technological University, Singapore, Singapore 2 School of Environmental Science & Engineering, Huazhong University of Science and Technology, Wuhan, China 3 School of Civil and Environmental Engineering, Nanyang Technological University, Singapore, Singapore

10.1 Introduction Nowadays, the generation of sewage sludge is increasing dramatically with global urbanization and economic development, which has brought a new challenge to the solid waste management [1, 2]. In general, sewage sludge with a large volume produced is rich in many different kinds of organic matter, which could be used as a potential resource for energy production via conventional and emerging treatment technologies [2]. Moreover, sewage sludge also contains abundant nutrients, e.g. nitrogen (N) and phosphorus (P), providing a unique opportunity for resource recovery [3, 4]. It has been believed that recovery of energy and resource is the ideal way toward a sustainable sewage sludge management. The traditional sewage sludge treatment (e.g. landfilling and land-farming) has been gradually phased out due to limited availability of land, emission of hazardous gases, and public health-associated environmental risks [5]. Thermal-chemical treatment has been commonly employed for managing sewage sludge, due to its advantages of energy and resource recovery, significant volume reduction, effective destruction of pathogens, etc. [6, 7]. The potential resources recoverable from sewage sludge may include N, P, and various species of metals [8, 9].In fact, the rapid development of the thermal-chemical process is largely driven by the emerging concept of “Waste-to-Energy and Resource.” However, a comprehensive and systematic analysis of such processes is required in terms of engineering feasibility and economic viability. Therefore, this chapter aims to offer a comprehensive picture of the state of the art of thermal-chemical technology for energy and resource recovery from sewage sludge. The chapter begins with a brief summary of the current status of sewage sludge and its impact on environmental sustainability and then introduces the sewage sludge characteristics, followed by a discussion on different thermal-chemical methods for energy and resource recovery from sewage Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

sludge. Meanwhile, the technology limitations and challenges are also outlined, together with the perspectives and directions forward.

10.2 Sewage Sludge and Its Impact on Environmental Sustainability According to UN-Habitat’s statistics (Figure 10.1), more than 6.5 million tons of dry sewage sludge were produced per year in the United States, and about 3.0 and 2.0 million tons generated annually in China and Japan, respectively [10]. These figures indeed stand on a rising slope due to the expansion of sewage treatment capacity in many developing countries. According to the levels of economic development, the sewage sludge management is largely country-specific. For example, in Germany, United Kingdom and United States, the reuse and disposal rate of sewage sludge through the major ways (e.g. landfill, incineration, and land-farming) already reached up to 84.1, 99.8, and 90%, respectively (Figure 10.1). It was reported that 54.7% of sewage sludge in Germany was used for energy recovery via incineration, while 78.3% of sewage sludge in United Kingdom was channeled to land-farming[5]. Comparatively, 97% of sewage sludge was disposed of mainly through landfills in China [5]. It should be clearly realized that the improper disposal of sewage sludge is not only a waste of organic resources, but also may cause serious secondary pollutions to water bodies, land, and atmosphere. Resource recovery from biosolids Energy recovery Industrial proceses

7000

Horticulture and landscaping

Forestry

5000 4000 3000 2000 1000

Possible alternative ways

2.5 2.0 1.5 1.0 0.5 0.0

l ga ay ey da zil lic nd ary nia kia r tu orw urk na Brapub inlaung ove ova Po N T Ca re F H Sl Sl h ec Cz

0

Incinerationa

er

m

an y

s m ly al y y a zil lic d y ia ia A na an nd do Ita tug vwa rkenad Bra ub lan gar en ak US hi ap r r Tu a p n n v v C J rla ng re Fi Hu Slo Slo C Po No he ki t d ch Ne ite ze n C U G

Major ways

Land reclamation

Landfilla

Estimated sludge production (×103 dry metric tons/year)

6000 Estimated sludge production (×103 dry metric tons/year)

248

Landfarminga

a

p.c. of sewage sludge utilization

USA: 28% China: 97%

USA: 15% China: 0.36%

USA: 47% China: 2.47%

Japan: 31% Germany: 0%

Japan: 0% Germany: 54.7%

Japan: 2% Germany: 29.4%

UK: 0.4%

UK: 21.1%

UK: 78.3%

Figure 10.1 Sludge generation rate in different countries and current practice in sewage sludge disposal. Date based on the statistics of UN-Habitat. Sources: Fijalkowski et al. [5], Zhen et al. [10]. Reproduced with permission from Elsevier; Zhen et al. [10]. © John Wiley & Sons.

Pretreatment

Raw wastewater

Sewage sludge

Atmospheric deposition

Biological process

Ammonla volatization

PM2.5 NOx SOx CO2

Combustion

Greenhouse gas Lechate

Sewage sludge

Lechate Underground water Landfill Steam

Dioxin

SOx

NOx, PM2.5

N, P, K Organic pollutants

Fertilizor factory

Industry

Urban run-off

Heavy metals pathogens

CO2

Heavy metals accumulation

Wastewater

Drying and incineration Surface run-off Plants

Heavy Organic Pathogen metals pollutants Land use

Figure 10.2

Farmland

Sewage sludge

Drain

Nutrient adsorption uptake release

Groundwater Decomposition

Impact of different sewage sludge disposal methods on environmental sustainability.

release

250

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

Figure 10.2 illustrates the impact of different methods for sewage sludge disposal on environmental sustainability. Landfill as one of the most traditional technologies for solid waste treatment has the drawbacks of generation of harmful gases and leachates, which are difficult to manage. Sewage sludge has also been commonly employed as a fertilizer and soil amendment for agriculture. However, the hazardous components in sewage sludge including heavy metals, organic contaminants, and pathogens may eventually deteriorate the quality of soil and cause a potential safety risk. As for sludge incineration, the exhaust gases released into the environment can lead to acid rain formation and global climate change, while the heavy metals concentrated in the ashes may enter the environment through leaching. Obviously, the traditional sewage sludge treatment methods, to a great extent, would lead to the serious deterioration of the environmental quality. Thus, an environmentally sustainable and economically viable approach is urgently needed for future sewage sludge management.

10.3 Characterization of Sewage Sludge To explore potential environmentally sustainable and economically viable approaches, it is necessary to briefly summarize the composition of the sewage sludge. In general, as shown in Figure 10.3, sewage sludge is a heterogeneous mixture of organics, inorganics, microorganisms, and water, of which nontoxic organic matter accounts for about 60% on a dry mass basis [11], with abundant plant nutrients (e.g. N, P, K) and multiple inorganic minerals. However, it is worth noting that sewage sludge often contains various toxic organic pollutants, heavy metals, and pathogenic microorganisms, which may cause potential secondary pollution and public health-associated environmental risks. Therefore, the valuable substances in sewage sludge should be recovered as energy and resource while the harmful substances should undergo appropriate treatment, controlled disposal and careful management. Generally, the moisture content in sewage sludge is up to 80% after mechanical dewatering [7], which seriously affects it residual energy. Wet sewage sludge usually shows negative residual energy from −0.89 to −1.91 MJ/kg while the dewatered and dried sewage sludge possesses positive residual energy from 0.24 to 12.44 MJ/kg [6]. Therefore, moisture content removal is necessary before using sewage sludge as fuel for energy recovery.

10.4 Thermal-Chemical Treatment of Sewage Sludge 10.4.1 Incineration 10.4.1.1 Typical Incineration Processes

Although rotary kiln, cyclone, and smelting furnaces have been commonly employed for incineration of sewage sludge, currently, both multiple hearth furnace (MHF) and fluidized bed furnace (FBF) are dominant in the market. A MHF

(polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), adsorbable organohalogens (AOX), pesticides, pharmaceuticals, hormones, etc.)

(Cellulose, lignin, soluble saccharide, grease and fats, protein)

Toxic organic matters (As, Cd, Cr, Hg, Pb, Zn, etc.)

Organic biomass composition

Heavy matters Organic

(Alcohol, acid, ester, hydrocarbon, humus, aromatic, compound)

inorganic Plant nutrients

Sewage sludge

Organic functional compounds

(N, P, K and its soluble states)

Inorganic minerals

Microorganism water (bacteria, viruses, protozoa along with other parasitic helminths) (free water, interstitial water, vicinal water, bound water)

Figure 10.3

Sewage sludge composition.

(Fe, Al, Ca, Si, etc. and their oxides, hydroxides)

252

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery off gas + fly ash

Sewage sludge

Freeboard

off gas + fly ash Additional fuel

Sewage sludge

Recycle gas

Fluid bed Distributor plate

(a)

Air

Air

(b)

Ash

Steam/heated fluld

Heat exchange

Hot combusted gas

Combustion air Hot gas

Waste heat recovery system

H2O Freebord

Fly ash

Electrostatic precipitator

Scrubber sludge

Wet scrubber

Dewatered Drying sewage sludge

Fluid bed

Heat

Distributor plate Preheated air Greenhouse gas

Chimney

Bottom ash Sewage sludge

(c)

Leachate Undergroud water Landfill

Figure 10.4 Typical incinerators available in the market. (a) Multiple hearth furnace; (b) fluidized bed furnace; and (c) a typical process for sewage sludge incineration. (Source: Reproduced with permission from Werther and Ogada [12] of Elsevier Publishing).

basically consists of a vertical cylindrical shell with multiple hearths, and a rotating shaft in the middle of the shell (Figure 10.4a). In this system, sewage sludge is fed from the top of furnace, while moving downward along the hearths by scrapers that are fixed on the rotating shaft. In the MHF system, sewage sludge undergoes four steps: (i) drying; (ii) pyrolysis; (iii) combustion; and (iv) ash-cooling [12]. It should be realized that in MHF, the internal energy in sewage sludge can be fully utilized by recycling the high-temperature flue gas in direct contact with fed sewage sludge, while additional auxiliary fuel would be needed for maintaining stable incineration of sewage sludge. As for FBF (Figure 10.4b), high-pressure air is introduced into the furnace from the bottom and moves upward through the entire sand bed, generating bubbling and flowing sands. These in turn lead to the excellent contact between sewage sludge and oxidant. Sewage sludge is quickly

10.4 Thermal-Chemical Treatment of Sewage Sludge

dried and burnt off in the bed, with the release of energy. The freeboard to serve as a post-combustion chamber may provide complete burnout of the combustible materials. Meanwhile, the fluidized bed material can eliminate thermal shock and reduce damage to the refractory, together with no rotating structure in the reactor, reducing the maintenance cost of the system. As shown in Figure 10.4c, a typical incineration process for sewage sludge includes drying, combustion, energy recovery, and cleaning. In general, dewatering and/or drying of sewage sludge have been commonly employed prior to incineration with the aim to increase the calorific value of sewage sludge. Then, sewage sludge together with compressed air is fed into the FBF to generate high-temperature flue gases. The hot gases and fine particulate inorganic matter can be then conveyed out of the furnace chamber, going through the energy recovery and cleaning units in sequence. As such, the energy can be recovered via a heat exchanger for steam generation or heated fluid; alternatively it can also be used as a heat source for preheating of combustion air and drying of sewage sludge. Bottom ash as a by-product can be collected from the furnace bottom, while the fly ash mixed with flue gas is captured in the cleaning unit. In order to meet the discharge standards, a multistage purification system including cyclone separator, wet scrubber, electrostatic precipitator, filter bag etc., is usually been adopted. 10.4.1.2 Performance–Cost–Benefit Analysis of Incineration Technology

Generally, incineration is a relatively mature technology for solid waste treatment. However, it has also been criticized that incineration has limited applicability attributed to its high cost and generation of hazardous gases. The cost associated with incineration mainly includes capital investment and operation cost. The capital investment is closely related to selection of incineration technology and its scale. For example, for an incineration plant of 1000 tons/d, the equipment cost may account for 48.2% of the total investment cost, while 21.5 and 26.0% for the construction and installation costs, respectively [13]. The capital cost with mechanical-grate incineration technology (MGIT) appears to be higher than that of circulating fluidized bed incineration technology (CFBIT). In the estimation of capital cost, the land cost is an uncertain factor, largely depending on local economic situation and governmental policy. On the contrary, the operation cost is mainly determined by the type of incineration technology adopted. For example, the operation cost of MGIT may range from US$16.3/ton to US$32.6/ton, and US$9.78/ton to US$19.56/ton for CFBIT [13]. The benefits of incineration may come from government subsidies, electric energy recovery, tax incentives, heat supplies, etc. In addition, incineration can also improve environmental sustainability of sewage sludge management with significant volume reduction, reduced emission of greenhouse gases, and small footprint.

10.4.2 Pyrolysis 10.4.2.1 Typical Pyrolysis Processes

Extensive effort has been dedicated to the development of the pyrolysis technology, and many different types of configurations for pyrolysis reactor are currently

253

254

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

available for fast pyrolysis of sewage sludge, including ablative system, stirred or moving beds, vacuum pyrolysis system, and fluidized beds [14]. By contrast, the fluidized bed reactor (FBR) is more advantageous to handle a wide spectrum of feedstocks (e.g. lignocellulosic biomass, sewage sludge) with rapid heating and better mass transfer among gas, particles, catalysts, etc. The bed is first heated by externally recycled gas and/or produced char/bio-oil combustion, followed by the second wave of heating through transferring the generated heat back to the bed. Moreover, a short residence time in FBR can also be easily achieved by controlling the fluidizing gas flow rate, while it should be realized that FBR suffers from its large reactor size and high construction cost [15]. The representative FBR includes bubbling fluidized bed, circulating fluidized bed, and entrained fluidized bed. Figure 10.5a shows the pyrolysis procedure of sewage sludge in bubbling fluidized bed reactor (BFBR). During the pyrolysis process, the derived char can be quickly eluted out by a cyclone. Due to the fast heating of sewage sludge and short residence time of gaseous products, the bio-oil derived in BFBR Syngas

Hot sand

Cleaning systems

Char + sand

Syngas

Combustor

Sewage sludge

Flue gas

Cooling systems

Cyclones

Char

Pyrolyser

Cleaning systems

Cooling systems

Cyclones Sewage sludge Fluid bed reactor

Syngas

Syngas

Tar

Tar Air Ash Syngas recycle

Syngas recycle for supplying heating

(a)

(b) Sewage sludge Syngas

Cyclones

Sand + char

Flue gas

Char combustor

Cooling systems

Cleaning systems

Syngas

Sand Hot sand

Syngas

Ash

Tar

(c)

Figure 10.5 (a) Bubbling fluidized bed reactor and integrated process; (b) circulating fluidized bed reactor and integrated process; (c) rotating cone pyrolysis reactor and integrated process. (Source: Reproduced with permission from Bridgwater [16] of Elsevier Publishing.)

10.4 Thermal-Chemical Treatment of Sewage Sludge

is of high quality and yield. The residence times of solids and pyrolysis gases are determined by the fluidizing gas flow rate, and char generally has a longer residence time than vapor. It had been shown that char contained a wide variety of catalytic metals that can be employed as an effective in situ catalyst for tar cracking/reforming [17]. Thus, fast and effective char separation is important for high-efficiency pyrolysis of sewage sludge. The circulating fluidized bed reactor (CFBR) has features similar to BFBR, but has the same residence time for both char and vapor. In the CFBR, the gas flow is intentionally set to be high enough for better recycling of sands (Figure 10.5b), while the char separated by cyclones is then combusted for sand reheating. Compared to the BFBR, the char content in bio-oils is much higher in CFBR. Rotating cone pyrolysis reactor (RCPR) is designed for intensified mixing and heat transfer between sewage sludge and heat carrier without the use of fluidizing gas leading to reduced operation cost (Figure 10.5c). In RCPR procedure, the sewage sludge is fed from the bottom of the rotating cone, and is further lifted to the rotating cone wall by a spiral motion with centrifugal force. Flash pyrolysis of sewage sludge can be achieved by the high-efficiency heat transfer between the wall and heated sand. The char and sand fall into the fluidized bed around the cone, and then move up to a separate fluidized bed combustor where the char is burned for heating the sand, and finally drops back into the rotating cone. This integrated process mainly includes a rotating cone pyrolyzer, a sand circulation unit, and a fluidized bed for char combustion. In addition to the pyrolysis processes discussed in this section, there are also a series of new developments in pyrolysis technology, including microwave-assisted pyrolysis[18], hydrothermal pyrolysis[19], catalytic pyrolysis[20], chemical looping[21], and iron ore reduction[22]. 10.4.2.2 Performance–Cost–Benefit Analysis of Pyrolysis Technology

Currently, pyrolysis has been considered as an advanced and practical technology for energy and resource recovery from sewage sludge. However, the properties of the pyrolysis products differ substantially, which can be attributed to the differences in composition and structure of sewage sludge that are derived from different wastewater treatment plants (WWTPs). In addition, the environmental sustainability and economic viability of pyrolysis of sewage sludge are largely determined by the process scale, selected technology, collection and transportation of sewage sludge, etc. Although sewage sludge is an unwanted by-product from WWTPs, the costs associated with its collection and transportation indeed is not negligible. For possible profit generation, the cost associated with feedstock should not exceed US$83/ton dry weight for slow pyrolysis and US$64/ton dry weight for fast pyrolysis [23].

10.4.3 Gasification 10.4.3.1 Typical Gasification Processes

Currently, fixed bed gasifier (FXBG) and fluidized bed gasifier (FBG) are the main options for gasification of sewage sludge. The FXBG has been widely used in the traditional gasification process, with the operating temperature of about

255

256

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery Sewage sludge

Sewage sludge

Sewage sludge

Syngas Fixed bed of sewage sludge

Fixed bed of sewage sludge Fixed bed of sewage sludge

Gasifying agent

Gasifying Gasifying agent agent

Syngas

Grate Ash

(a)

Gasifying agent

(b)

(c)

Syngas + Ash

Ash

Syngas

Syngas Syngas

Bed materials Sewage sludge Ash

Cylones

Cylones

Syngas

Bed materials Sewage sludge

Ash

(d)

Gasifying agent

(e)

Gasifying agent

Figure 10.6 Typical gasifiers for sewage sludge: (a) updraft; (b) downdraft; (c) cross-draft; (d) bubbling fluidized bed; (e) circulating fluidized bed. (Source: Reproduced with permission from Sansaniwal et al. [24] of Elsevier Publishing.)

1000 ∘ C. According to the airflow pattern, the FXBG can be divided into updraft, downdraft, and cross-draft gasifier. Updraft gasifier has the simplest structure as shown in Figure 10.6a, in which sewage sludge is fed from the top of the reactor, then slowly flows down to pass through the reaction zones by gravity. Inversely, the gasifying agent is introduced from the bottom through a grate and the pyrolytic vapors are carried upward by the up-flowing hot product gas. After gasification, the residual ash is free falling and collected by discharge unit that is located at the bottom of the gasifier. Despite the high overall energy efficiency, unwanted tar can still be generated and interfused in gaseous product. Such tar is commonly in high concentration, which can restrict the application of gaseous product. In the downdraft gasifier, sewage sludge is also fed into the system from the top (Figure 10.6b), but the gasifying agent is introduced from the gasifier side to react with the pyrolysis products. Compared with the updraft gasifier, downward gasifier has a lower energy efficiency and lower tar content in the gaseous product, but with a high particulate content. Cross-draft gasifier (Figure 10.6c) has the same

10.4 Thermal-Chemical Treatment of Sewage Sludge

feeding method of sewage sludge and gasifying agent as downdraft gasifier except the gas product outlet, which is on the side opposite of the gasifying agent inlet. The high-velocity gasifying agent is introduced into the gasifier and burns with char, forming a high-temperature zone (>1500 ∘ C), which is conducive to the reduction of tar in product gas. Compared with the fixed bed, fluidized bed has the advantages of higher heat and mass transfer efficiency, and better solid-phase mixing. According to the fluidization velocity of medium, the fluidized bed can be divided into bubbling fluidized bed (BFB) and circulating fluidized bed (CFB). In BFB, as shown in Figure 10.6d, sewage sludge is fed from the side of reactor, and the gasifying agent enters from the bottom. In general, the movement of gasifying agent is controlled at its minimum velocity just above that required for fluidization of the bed material. Meanwhile, the produced gas is discharged from the top of the BFB gasifier and the ash is removed by a cyclone. CFB system (Figure 10.6e) normally consists of (i) fluidization unit, in which the fluidizing medium is driven by the gasifying agent in higher velocity than that in BFB, and (ii) a cyclone couple with a circulating system. So far, the CFB has been considered as the most promising technology for sewage sludge gasification with the advantages of the strong solid-phase mixing capacity, long retention time, etc. Especially, in CFB, catalysts can serve as part of the gasifier bed, helping to in situ reform the tar [25, 26], while the energy consumption of CFB is relatively high compared with BFB. 10.4.3.2 Performance–Cost–Benefit Analysis of Gasification Technology

The capital and operation costs of the gasification plants are mainly determined by the treatment capacity, type of gasifier, and potential of reusing end product. The capital cost mainly includes collection, transportation, and storage of sewage sludge, equipment, land, etc. According to the scale of gasification plant, the operation cost usually falls into the range of US$15/ton–US$ 230/ton [27], in which the auxiliary materials (i.e. chemicals, bed materials, catalysts) may contribute to about 3.5–5.7% of the total operation cost [27]. For large-scale sewage sludge gasification plants, the revenue can be derived from the government subsidies and end products. In general, the government subsidies to sewage sludge gasification ranges from US$42 to 166/ton [28]. However, the produced tar in sewage sludge gasification is difficult to be separated from gaseous products, leading to reduced commercial value of the end products. Apart from the economic benefits, gasification appears to be more environmentally friendly than incineration. One such example is that the PM2.5 emission in gasification is less than a third of that in incineration, while the gasification may also help to reduce the generation of carcinogens, e.g. the chloroethylene production in incineration was 54-fold higher than that in gasification [29].

10.4.4 Liquefaction 10.4.4.1 Typical Liquefaction Processes

Liquefaction, also known as hydrothermal liquefaction (HTL), is usually conducted in an autoclave stirred reactor made of stainless steel, which can work under high

257

258

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

pressure and medium temperature. The heat for liquefaction in autoclave stirred reactor can be provided by various means, e.g. molten tin bath [30], heating jacket [31], external electric heaters [32], and small GC-oven [33]. The sewage sludge, solvent, and catalyst are first mixed up and then injected into the autoclave stirred reactor purged with N2 or H2 gas to remove the residual air. Then, the reactor is pressurized with N2 or H2 gas to a desired pressure without heating, and the pressurized reactor is then heated to the set point temperature for the liquefaction of sewage sludge. It is worth noting that the pressure should be gradually increased with increase in temperature. The stirring is maintained at constant velocity during the whole heating period up to the desired temperature. The retention time of liquefaction of sewage sludge depends on the experimental conditions. After the desired reaction time, the stirring is off and the reactor is cooled to room temperature for harvesting end products. In the liquefaction process, the solid–liquid separation is a big challenge. In general, the liquid–solid mixture is filtered, and the filtrate with water and soluble organics and inorganics is collected. Then, acetone may be used to wash the reaction vessel for recovery of the residual deposits. The collected acetone solution mixed with the previously collected filtrate undergoes another round of filtration. After evaporation of newly collected filtrate, the extracted dark brown viscous material is known as biocrude. 10.4.4.2 Performance–Cost–Benefit Analysis of Liquefaction Technology

Biocrude produced from liquefaction of sewage sludge is considered as superior substitute of commercial fossil fuels. Different from pyrolysis, drying of sewage sludge is not necessary for liquefaction, indicating that sewage sludge can be directly liquefied without prior drying. This indeed helps to greatly reduce the operation cost of liquefaction compared with pyrolysis. As the solvents are employed as catalysts in the liquefaction process, high-quality end products can be produced with the lower water content and higher heating value. All these together clearly suggest that liquefaction can offer a feasible engineering solution for sewage sludge management; on the contrary, the high pressure required for liquefaction may pose a challenge to its economic viability.

10.5 Recovery of Energy and Resource from Sewage Sludge 10.5.1 Combustible Gas Combustible gas also known as synthesis gas (syngas) is the main product derived from sewage sludge gasification. Generally, the composition of combustible gas is complex. Basically, the combustible gas is a mixture of CO, H2 , CO2 , hydrocarbons including CH4 , C2 H6 , C2 H4 , C3 H8 , and small amount of other gases (e.g. NH3 , NOx , SOx ) and some contaminants (e.g. carbon particles, tar, and ash) [34]. The composition of combustible gas is related to gasification conditions (e.g. temperature, steam-to-SS ratio, gas residence time, addition of catalyst materials, etc.) [35].

10.5 Recovery of Energy and Resource from Sewage Sludge

The lower heating value (LHV) of the combustible gas usually falls in the range 4–20 MJ/Nm3 , depending on the type of gasifying agent. For example, when air or pure oxygen was used as gasifying agent, the LHV of the combustible gas was found to be in the range of 4–6 and 10–15 MJ/Nm3 , respectively. When steam was utilized as the only gasifying agent, the combustible gas was found to be rich in hydrogen, which contributed to higher LHV of 13–20 MJ/Nm3 [36]. Gasification is an endothermic reaction; temperature dominates the energy supply of gasification and has been considered as the most important operational parameter [17, 37]. Obviously, high temperature can help to reduce tar generation, while increasing the yield of combustible gas, leading to improved release of the volatile matter and increased tar cracking and char gasification rates. Consequently, combustible gas produced at high temperature appears to be rich in H2 , CO, and CH4 [38]. As discussed previously, high temperature, high heating rate, and long gas retention time all favor the production of hydrogen-rich combustible gas from the pyrolysis of wet sewage sludge [11]. This is likely because auto-generated hot steam atmosphere from wet sewage sludge gasification is helpful for a sequence of reactions to produce hydrogen [39]. In addition, as an energy-carrying by-product, the generation of tar not only reduces the energy recovery efficiency of the reaction system, but also poses technical hurdles to the commercial usage of produced combustible gases. Thus, it is reasonable to consider that in situ removal and energy recovery from tar are essential during the sewage sludge gasification, and help to avoid subsequent separation with reduced operation cost. For example, an integrated thermal cracking and catalytic reforming process had been developed for reducing tar, while increasing the gas yield during biomass gasification [37]. It was found that the higher gas yield of 1.36 Nm3 /kg with the LHV of 11.61 MJ/Nm3 was obtained. In addition, the H2 /CO ratio in the produced gases is also critical for subsequent industrial application. For example, the gasoline preparation may need an H2 /CO ratio of 0.5–1.0, while it may need about 2.0 for methanol production. Therefore, for production of target products, the H2 /CO ratio needs to be adjusted via different control strategies, such as reforming, shift reactions, etc. In general, the main problems of using combustible gas are related to the presence of particulate matter, tar, NOx , SOx , etc. The presence of these impurities in combustible gas is problematic because they can cause serious problems to the normal operation of gasification/pyrolysis reactors, such as equipment corrosion and catalyst poisoning. Therefore, the purification of combustible gas is a key step before its commercial applications.

10.5.2 Bio-oils Bio-oils derived from pyrolysis of sewage sludge possess different properties in terms of viscosity, pH, water content, and chemical compositions due to the heterogeneity in sewage sludge and operation conditions. In general, bio-oils obtained from sewage sludge pyrolysis are free-flowing liquids of dark brown color, mainly consisting of water (e.g. 23.0–69.7%) and various organic compounds, such as alkanes, aromatic compounds, fatty acids, ketones, fatty nitriles and amides [40, 41]. Inter-reactions

259

260

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

among these organic compounds may occur constantly in bio-oils. It was suggested that the properties of bio-oils respectively produced from pyrolysis of sewage sludge and lignocellulosic biomass were very similar, while there was a significant difference in pH, e.g. pH 8–10 for sewage sludge and 2–4 for lignocellulosic biomass [42]. The alkalinity of bio-oils derived from sewage sludge pyrolysis was attributed to the existing ammonia and other nitrogen-containing compounds, which were contributed by microorganisms in sewage sludge. In general, nitrogen-containing compounds contained in bio-oil can be easily converted into NOx during the combustion process, resulting in serious environmental pollution issues. Therefore, effectively controlling the nitrogen content in bio-oil when employing it as fuel is the key to realizing clean energy and sustainable resource recovery from sewage sludge. The water content in bio-oils is closely related to the original moisture of sewage sludge and the dehydration reactions during the homogeneous secondary reactions as well as the decomposition of oxygen-containing heavy organic compounds [43]. Obviously, a high water content can largely lower the heating value of bio-oils. In general, the higher heating value (HHV) of bio-oils derived from pyrolysis of sewage sludge is in the range of 10–32 MJ/kg, which is much lower than that of petroleum fuels (i.e. 42–45 MJ/kg) [41, 44]. Meanwhile, the high oxygen content in pyrolytic bio-oils can lead to high corrosiveness, instability, and low energy density of the produced bio-oils [45]. Thus, catalytic deoxygenation and hydrodeoxygenation should be employed for reducing the oxygen content, while increasing the HHV of bio-oils during the ex situ or in situ pyrolysis of sewage sludge. Bio-oils have clear economic value, with many potential applications, e.g. supply of heat and power to boilers, engines, and turbines. Moreover, bio-oils rich in valuable chemical compounds can also be used as feedstock for producing high value-added chemicals, such as preservatives, resin precursors, additives, flavoring agents, acetic acid, hydroxyacetaldehyde, levoglucosenone, and maltol [46–49].

10.5.3 Biochar Biochar is a low-cost carbon-rich material present in amorphous forms and with porous structure. It can be obtained from pyrolysis of various organic solid wastes, such as sewage sludge, lignocellulosic biomass, and food waste [50]. Biochar is the major product of slow pyrolysis of sewage sludge, which is generally operated at a long residence time and a low heating rate. Compared with slow pyrolysis, biochar produced by fast pyrolysis has a smaller surface area [51]. The properties of biochar are mainly affected by many parameters. Pyrolysis temperature is a crucial factor that significantly affects the physicochemical properties of biochar [52]. Studies have shown that the yield of biochar decreased with increase in pyrolysis temperature, with changes in biochar structure [53]. In addition, an increased pyrolysis temperature results in the reduction of total nitrogen content, water sorption capacity, and cation-exchange capacity (CEC) of produced biochar; on the other hand, pH, carbon content, available nutrients content, ash content, and heavy metal stability tend to increase accordingly [53].

10.5 Recovery of Energy and Resource from Sewage Sludge

The observed pH increase of biochar derived from higher temperature was attributed to the polymerization/condensation reactions of the aliphatic compounds and dehydration reactions of feedstock during pyrolysis of sewage sludge, which can cause the decrease of the acidic surface groups [54]. Moreover, the increase in alkali metal salts concentration of biochar with the increased temperature may also lead to pH increase [55]. As temperature increases, the increase in carbon content in biochar may be related to the loss of –OH surface functional groups [56, 57]. The lowered cation-exchange capacity of biochar at high temperature is likely due to the decrease in acidic functional groups [58]. It had been reported that the hydrophobicity of biochar was improved with gradual removal of the polar surface functional groups at high temperature, which indeed can prevent moisture from entering the porous structure of biochar [59]. Biochar is a good adsorbent for pollutant removal from wastewater. Thus, the conversion of sewage sludge into biochar based on pyrolysis/gasification processes can not only realize volume minimization of sewage sludge, but also product the adsorbents for pollutants removal towards a green strategy for removing pollutant by waste. It was reported that biochar derived from sewage sludge is rich in various surface functional groups and minerals, which can be used as a raw material for further synthesis of various high-value functionalized carbon materials [56]. Although functionalization of biochar is still at the infant stage, biochar-based functional materials have exhibited increasing applications in adsorption [50, 60], soil amendment [61], catalysis [62], and energy storage [63].

10.5.4 Ashes to Value-Added Materials Ash, a residue obtained from sewage sludge incineration, is a polyphasic material with mean particle diameters ranging from 8 to 263 μm [64]. In general, the major elements in ash are Si, Ca, P, Al and Fe. Crystalline forms of these elements are invariably quartz (SiO2 ),whitlockite (Ca3 (PO4 )2 ) and hematite (Fe2 O3 ) [8, 64]. Particularly, Al is present in the forms of feldspar and amorphous glassy phases according to XRD pattern [65]. The content of amorphous glassy phases may vary in different ashes, which is an important characteristic for ash to be used as a potential pozzolanic additive in blended cements. The extractable features of the major elements in ash highly relate to the nature of sewage sludge, treatment process, and dewatering process. In addition, a significant amount of heavy metals commonly exist in ash [66]. So far, landfill is still the main way for ash disposal, while the high heavy metal content in ashes poses a serious concern on such practices [66]. As such, there has been a growing interest in exploring the preparation of the value-added functional materials with incineration ashes as raw material as summarized in Figure 10.7.

10.5.5 Nutrient Recovery Undoubtedly, the sewage sludge is rich in nutrients with the nitrogen and phosphorus contents of about 2.4–5.0 and 0.5–0.7% in dry weight, respectively [67]. During the decomposition and dissolution of sewage sludge, released ammonia could react

261

262

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

(General ceramics, bricks, Tiles, glass-ceramics)

(Cement clinker production, light weigh aggregate, mortars, normal-weight concrete, blocks, aerated concrete, foamed concrete, Controlled low-strength materials

Concrete-related application

Ceramic application

Ash Road pavements application

Geotechnical application (unbound, hydraulically bound, bituminous bound) (As a stabilization additive for the treatment of soft soil, fill material)

Figure 10.7

Application of ashes from sewage sludge incineration.

with phosphates to generate magnesium ammonium phosphate (i.e. struvite) in the presence of magnesium [68]. The main form of phosphorus is the struvite, which has been known as a slow-releasing fertilizer that can be directly used as soil synergist for plant growth [69]. The global demand on fertilizers already stood at about 45 million tons of phosphates as P2 O5 and 116 million tons of nitrogen-N in 2016. Theoretically, the nutrients recovered from 30 million tons of sewage sludge produced globally could offset 5% of total phosphate consumption and 1.7% of total nitrogen consumption annually [9]. It has been predicted that phosphate rocks as a nonrenewable resource may be exhausted within one century [70]. Therefore, new strategies for phosphorus recovery from sewage sludge are urgently needed. Recently, extensive effort has been devoted to recover P from sewage sludge using chemical precipitation. Compared with Al- or Fe-based phosphates, Ca- or Mg-based phosphates have much higher commercial value as fertilizer [71]. It has been reported that the Ca-based phosphate (hydroxyapatites) has the same chemical composition as the mined phosphate ore, which is easily recycled in the phosphate industries, while Mg-based phosphate (struvite) can be directly used as fertilizer for plant growth. However, the cost of P recovery from sewage sludge is about 1.6- to 3.0-fold higher than that of the commercial P [72]. The investment cost for struvite production at the capacity of 231 kg/d was estimated to be about £50 000 per year, while the revenue generated was only about £16 000–20 000 per year [73]. Obviously, at the current technology status, P recovery from sewage sludge through chemical precipitation should not be economically viable. The P recovery from incinerated ash of sewage sludge has been actively explored as the P content in incineration ash reaches 5–11% by dry weight, which is much

10.5 Recovery of Energy and Resource from Sewage Sludge

higher than that in sewage sludge (e.g. 1–5% by dry weight) [72, 74]. The P recovery from incineration has been realized mainly through acid leaching and thermochemical treatment. Previous work by Oliver and CaretY [75] reported that about 76 and 61 wt% of total P on average could be recovered from eight ash samples leached with H2 SO4 and HCl at pH = 1.5. However, heavy metals in incineration ash could also be released simultaneously during acidic leaching. In order to solve this problem, multiple-step leaching process had been proposed [76], in which the H2 SO4 solution at pH = 2 was first added into the ashes, followed by solid–liquid separation. Then, the pH of separated liquid was adjusted to 4, followed by addition of a certain amount of Al2 (SO4 )3 . At this stage, the P in ash was precipitated out in the form of AlPO4 . Lastly, adding NaOH or Ca(OH)2 to the separated liquid for precipitating heavy metals at pH above 10, the end P product with abundance of 89–93% could be recovered. The thermochemical method has also been applied for P recovery from incineration ash. Different from acid leaching, the metals in ash could be volatilized and removed by thermochemical treatment at high temperature of 900–1100 ∘ C with addition of chlorine salts, such as NaCl, KCl, MgCl2, or CaCl2 [77]. Under appropriate conditions, the metal chlorides could be volatilized, while P is converted to fertilizer. However, about 30% of P in ash may be lost in the thermochemical treatment process. As such, some modified processes have been developed to reduce the P loss, e.g. use of granulated ash [78]. Currently, the high operation cost and equipment corrosion are still the major bottlenecks for process scale-up; thus, more cost-effective and environment-friendly P recovery technologies should be explored to meet the market needs.

10.5.6 Heavy Metals Removal and Recovery The level of heavy metals contained in sewage sludge may be as high as 1000 ppm, which may cause serious environmental risks [79]. It has been reported that about 50–80% of heavy metals in WWTPs was fixed into sewage sludge [80]. Therefore, the heavy metals in sewage sludge may be a problematic case for subsequent chemicals preparation and energy recovery. For this purpose, extraction and solidification have been employed for removing heavy metals from sewage sludge. In general, metal ion can stably bind to the cell surface under neutral condition. The addition of acidic or alkaline agents in sewage sludge can improve or decrease the heavy metals migration efficiencies. Gasco et al. [81] studied the pyrolytic behaviors of sewage sludge by acid treatment (HCl : HNO3 = 4 : 3, pH = 1, 2). The study indicated that acid treatment was beneficial to metal removal. Additionally, more than 70% of Cu and Zn was removed from sewage sludge by a Fenton reagent against about 40% removal of Cd and Ni by leaching [82]. In order to enhance the recovery of heavy metals, a two-stage acid leaching process coupled with ultrasonic-assistance with an efficiency of over 90% had been developed for the extraction and recovery of heavy metals from actual electroplating sludge [83]. However, this process suffered from the drawbacks of high operation cost and generation of waste acid solution. Recently, bioleaching has been considered as an environmentally friendly

263

264

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

approach for metals recovery from sewage sludge, leading to about 80% of saving in the operation cost compared with the traditional chemical approaches [84]. By bioleaching, the removal efficiencies for Cu and Zn from sewage sludge could reach more than 90% [85]. Currently, some integrated processing technologies have also been developed toward enhanced metals removal/recovery from sewage sludge, while lowering the operation cost. The bioleaching coupled with electrokinetic remediation technology has also been employed for the removal of heavy metals from sewage sludge with the removal efficiencies of 78.61 and 99.11% for Cu and Zn, respectively [85]. In addition, methods of bioleaching coupled with Fenton or Fenton-like reactions also display good removal efficiency of heavy metals [86].

10.6 Technology Limitations and Challenges 10.6.1 Deactivation of Catalyst In the thermal-chemical processes, catalyst plays a vital role in improving the reaction efficiency and product quality. Ideally, a catalyst should have high activity, stability, and durability as these will substantially reduce the frequency of catalyst regeneration. In practice, many factors may cause catalyst deactivation, such as fouling, coke/carbon formation, poisoning, sintering, and phase change [87–89]. For acid- or base-type catalysts, the deactivation of catalyst is mainly due to coking, whereas the deactivation would result from the formation of coke or the synergy between coking and metal sintering once supported metallic oxide catalysts is utilized. In some cases, the activity of catalysts may be decreased due to the chemical transformations or stream washing of the metals, leading to the loss of active species on the catalyst surfaces. Consequently, the deactivation mechanisms are catalystand process-specific. Lastly, it is worth noting that the recovery of precious metals from used catalysts should be considered to make thermal-chemical processes more cost-effective.

10.6.2 Tar Formation Tar is a by-product generated from sewage sludge pyrolysis or gasification, and it can cause blockage of downstream equipment, formation of tar aerosols, etc. [20]. Several approaches are currently available for tar minimization/removal, e.g. physical treatment, thermal cracking, use of nonthermal plasma, and catalytic cracking/reforming [62]. Physical treatment (e.g. scrubbing and electrostatic precipitation) can be carried out effectively at lower temperature; thus additional condensation is needed due to high tar temperature. Recently, rotating particle separator (RPS) was adopted for tar removal, but with low efficiency [90]. Thermal cracking of tar needs the higher temperature of 1100–1300 ∘ C, challenging its economic feasibility [91]. While plasma had also been applied for tar decomposition, this technology was too complex to be applied at large scale [92]. However, it should be realized that the activation energy of tar decomposition can be reduced

10.6 Technology Limitations and Challenges

in the presence of catalyst, suggesting that catalytic cracking/reforming may be practically feasible and economically viable for tar removal and minimization [20], but deactivation of catalysts is still a serious problem. Therefore, a highly active and durable catalyst is urgently needed for more cost-effective tar minimization and removal.

10.6.3 NOx and SOx Emission Sewage sludge generally contains over 9% N and 1% sulfur by dry weight, which may cause potential secondary pollution of photochemical smog and acid rain during the thermal-chemical treatment of sewage sludge [93, 94]. For example, nitrogen and sulfur in sewage sludge can be converted to NH3 , HCN, and H2 S gases under the reduction conditions, while NOx and SOx can be generated under the oxidation condition. In order to reduce the formation of N/S-containing gaseous products, many efforts have been made to study the species, structures, and distribution of nitrogen and sulfur generated during the thermal-chemical treatment of sewage sludge. Different nitrogen species were found during thermal-chemical treatment of sewage sludge and coal, i.e. nitrogen mainly existed as heterocyclic aromatic hydrocarbons for coal, while protein for sewage sludge [93, 95]. HCN and NH3 had been shown to be the major N-containing gases produced from the sewage sludge pyrolysis, which might be further converted to NOx . In order to minimize the NOx emission, it is essential to have a better understanding of the sewage sludge–nitrogen nexus during pyrolysis, which is still a main challenge. The mechanism of nitrogen migration and transformation indeed is very complex due to the heterogeneity of sewage sludge and different pyrolysis conditions. For example, Cao et al. [93] found that the high-concentration NH3 was observed in the sewage sludge pyrolysis process, while high-concentration HCN was detected in the study by Tian et al. [96]. Organic sulfur in sewage sludge should not be neglected, which includes sulfonic acid, sulfoxide, aromatic sulfur, and aliphatic sulfur [97]. During sewage sludge pyrolysis, the S-containing gases were mainly produced from the decomposition of aromatic sulfur and aliphatic sulfur. It should be pointed out that these gases can easily corrode metal surfaces, and can also be oxidized to SOx . Therefore, an engineering strategy for controlling the generation of NOx and SOx is urgently needed for future application of the thermal-chemical process.

10.6.4 High Moisture Content The moisture content (e.g. about 80%) in sewage sludge seriously compromises the energy recovery efficiency during the gasification and incineration of sewage sludge. During gasification, the moisture content at the appropriate level can promote the decomposition of tar and gasification of biochar to generate syngas; however, an excessive of moisture content should be needed more energy for evaporation. In addition, sewage sludge with high moisture content presents a pasty consistency, which poses great challenges to continuous feeding. It has been reported that the allowable maximum moisture content in sewage sludge is about 25% in downdraft

265

266

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

gasifier and 50% for updraft gasifier [98]. The gas yield was found to significantly decrease at a moisture content of over 30% in sewage sludge gasification. Therefore, it is necessary to reduce the moisture content to an acceptable level prior to the sewage sludge thermal-chemical treatment.

10.7 Conclusions and Perspectives The amount of sewage sludge has been continuously increasing with the expansion of global population and urbanization. Traditional treatment methods for sewage sludge, such as landfill and land-farming, are no longer acceptable for future sewage sludge management, which indeed are being banned in more and more countries. In such a situation, an environmentally friendly and economically viable technology is urgently needed for proper handling of sewage sludge. Currently, thermal-chemical technology has been considered as a promising approach for sewage sludge management, with the aims for concurrent volume reduction, pathogens destruction, and energy recovery. The end products from thermal-chemical processes have high commercial values, e.g. combustible gas, bio-oils, biochar, value-added materials, nutrients, and heavy metals. Despite the advantages of the thermal-chemical platform, the end products usually have an unsatisfactory quality due to the complex nature of sewage sludge. For example, more effort has been made to optimize the thermal-chemical process for production of high-quality combustible gas and bio-oils to meet the requirement for replacing nonrenewable petroleum fuels. However, in the thermal-chemical process, the nitrogen and sulfur originally present in sewage sludge are converted to various harmful gases, such as NH3 , HCN, and H2 S under the reduction conditions, and NOx and SOx under the oxidation condition. Therefore, the technical solution for reducing the generation of such hazardous gases is urgently needed for making thermal-chemical process more environmentally friendly. Polycyclic aromatic hydrocarbons (PAHs) in the bio-oils produced from sewage sludge at the pyrolysis temperature higher than 700 ∘ C will pose a serious limit on the commercial application of bio-oils. Moving forward, economical minimization of PAHs production or conversion of PAHs to nontoxic substances should be explored in a timely manner. Catalytic cracking/reforming has been considered as an economically viable and environmentally sustainable technology for tar minimization and removal. However, deactivation of catalysts is still a hurdle; thus the development of a highly active and durable catalyst is a promising research direction for tar minimization and removal. High moisture content of sewage sludge is another obstacle that seriously affects the energy recovery potential of thermal-chemical process. Although considerable progress has been made in thermal-chemical conversion of sewage sludge in the recent years, the following issues still need to be tackled in the future: i. The production efficiency of combustible gas with high calorific value (e.g. syngas, rich-H2 ) should be improved in consideration of integrating with in situ carbon dioxide capture.

References

ii. The bio-oils obtained from thermal-chemical methods are usually unstable, highly corrosive, with low calorific value due to their high oxygen content. In this regard, effective removal of oxygen should be addressed with the aim to improve the quality of bio-oil. Moreover, new catalysts are also needed for reducing the generation of tar, NOx , SOx , etc. Thus, highly active, durable, low-cost catalysts with deoxygenation function should be explored for future sewage sludge thermal-chemical treatment. iii. The transformation mechanisms of inorganic elements of sewage sludge into end products in the thermal-chemical processes should be examined. iv. Thermal-chemical conversion of sewage sludge offers a promising option for producing value-added functional materials, but some critical issues still remain unsolved, i.e. detailed transformation mechanism of sewage sludge in the thermal-chemical process and the relationship between the surface functionality and porosity of char materials derived from sewage sludge and catalytic activity, energy storage performance, and adsorption capacity.

References 1 Frišták, V., Pipíška, M., and Soja, G. (2018). Pyrolysis treatment of sewage sludge: a promising way to produce phosphorus fertilizer. Journal of Cleaner Production 172: 1772–1778. 2 Luo, S. and Feng, Y. (2016). The production of hydrogen-rich gas by wet sludge pyrolysis using waste heat from blast-furnace slag. Energy 113: 845–851. 3 Lee, C., Alvarez, P.J.J., Kim, H. et al. (2018). Phosphorous recovery from sewage sludge using calcium silicate hydrates. Chemosphere 193: 1087–1093. 4 Munir, M.T., Li, B., Boiarkina, I. et al. (2017). Phosphate recovery from hydrothermally treated sewage sludge using struvite precipitation. Bioresource Technology 239: 171–179. 5 Fijalkowski, K., Rorat, A., Grobelak, A. et al. (2017). The presence of contaminations in sewage sludge – the current situation. Journal of Environmental Management 203: 1126–1136. 6 Chan, W.P. and Wang, J. (2016). Comprehensive characterisation of sewage sludge for thermochemical conversion processes – based on Singapore survey. Waste Management 54: 131–142. 7 Jiang, L., Yuan, X., Li, H. et al. (2016). Co-pelletization of sewage sludge and biomass: thermogravimetric analysis and ash deposits. Fuel Processing Technology 145: 109–115. 8 Donatello, S. and Cheeseman, C.R. (2013). Recycling and recovery routes for incinerated sewage sludge ash (ISSA): a review. Waste Management 33 (11): 2328–2340. 9 Mulchandani, A. and Westerhoff, P. (2016). Recovery opportunities for metals and energy from sewage sludges. Bioresource Technology 215: 215–226. 10 Zhen, G., Lu, X., Kato, H. et al. (2017). Overview of pretreatment strategies for enhancing sewage sludge disintegration and subsequent anaerobic digestion:

267

268

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

11

12 13

14

15

16 17

18

19 20 21

22

23

24

25 26

current advances, full-scale application and future perspectives. Renewable and Sustainable Energy Reviews 69: 559–577. Manara, P. and Zabaniotou, A. (2012). Towards sewage sludge based biofuels via thermochemical conversion – a review. Renewable and Sustainable Energy Reviews 16: 2566–2582. Werther, J. and Ogada, T. (1999). Sewage sludge combustion. Progress in Energy and Combustion Science 25: 55–116. Xin-gang, Z., Gui-wu, J., Ang, L. et al. (2016). Technology, cost, a performance of waste-to-energy incineration industry in China. Renewable and Sustainable Energy Reviews 55: 115–130. Sharma, A., Pareek, V., and Zhang, D. (2015). Biomass pyrolysis – a review of modelling, process parameters and catalytic studies. Renewable and Sustainable Energy Reviews 50: 1081–1096. Bok, J.P., Choi, Y.S., Choi, S.K. et al. (2014). Fast pyrolysis of Douglas fir by using tilted-slide reactor and characteristics of biocrude-oil fractions. Renewable Energy 65: 7–13. Bridgwater, A.V. (2012). Review of fast pyrolysis of biomass and product upgrading. Biomass and Bioenergy 38: 68–94. Hu, M., Guo, D., Ma, C. et al. (2015). Hydrogen-rich gas production by the gasification of wet MSW (municipal solid waste) coupled with carbon dioxide capture. Energy 90: 857–863. Zhang, Y., Liu, S., Fan, L. et al. (2018). Oil production from microwave-assisted pyrolysis of a low rank American brown coal. Energy Conversion and Management 159: 76–84. Wilk, M. and Magdziarz, A. (2017). Hydrothermal carbonization, torrefaction and slow pyrolysis of Miscanthus giganteus. Energy 140: 1292–1304. Hu, M., Laghari, M., Cui, B. et al. (2018). Catalytic cracking of biomass tar over char supported nickel catalyst. Energy 145: 228–237. Sun, Y., Seetharaman, S., and Zhang, Z. (2018). Integrating biomass pyrolysis with waste heat recovery from hot slags via extending the C-loops: product yields and roles of slags. Energy 149: 792–803. Li, X., Hui, H., Li, S. et al. (2016). Integration of coal pyrolysis process with iron ore reduction: reduction behaviors of iron ore with benzene-containing coal pyrolysis gas as a reducing agent. Chinese Journal of Chemical Engineering 24: 811–817. Brown, T.R., Wright, M.M., and Brown, R.C. (2011). Estimating profitability of two biochar production scenarios: slow pyrolysis vs fast pyrolysis. Biofuels, Bioproducts and Biorefining 5: 54–68. Sansaniwal, S.K., Pal, K., Rosen, M.A. et al. (2017). Recent advances in the development of biomass gasification technology: a comprehensive review. Renewable and Sustainable Energy Reviews 72: 363–384. Gómez-Barea, A., Ollero, P., and Leckner, B. (2013). Optimization of char and tar conversion in fluidized bed biomass gasifiers. Fuel 103: 42–52. Udomsirichakorn, J., Basu, P., Salam, P.A. et al. (2013). Effect of CaO on tar reforming to hydrogen-enriched gas with in-process CO2 capture in a bubbling

References

27 28

29

30

31

32

33

34 35

36 37

38

39

40 41

fluidized bed biomass steam gasifier. International Journal of Hydrogen Energy 38: 14495–14504. Watson, J., Zhang, Y., Si, B. et al. (2018). Gasification of biowaste: a critical review and outlooks. Renewable and Sustainable Energy Reviews 83: 1–17. Yassin, L., Lettieri, P., Simons, S.J.R. et al. (2009). Techno-economic performance of energy-from-waste fluidized bed combustion and gasification processes in the UK context. Chemical Engineering Journal 146: 315–327. Arena, U., Ardolino, F., and Di Gregorio, F. (2015). A life cycle assessment of environmental performances of two combustion- and gasification-based waste-to-energy technologies. Waste Management 41: 60–74. Liu, D., Li, Q., Zhao, A. et al. (2015). Hydro-liquefaction of sawdust and its three components in supercritical ethanol with [BMIM]Cl/NiCl2 catalyst. Chemical Engineering Journal 279: 921–928. Trautmann, M., Lang, S., and Traa, Y. (2015). Direct liquefaction of lower-rank coals and biocoals with magnetically separable catalysts as a sustainable route to fuels. Fuel 151: 102–109. Toor, S.S., Reddy, H., Deng, S. et al. (2013). Hydrothermal liquefaction of Spirulina and Nannochloropsis salina under subcritical and supercritical water conditions. Bioresource Technology 131: 413–419. López Barreiro, D., Beck, M., Hornung, U. et al. (2015). Suitability of hydrothermal liquefaction as a conversion route to produce biofuels from macroalgae. Algal Research 11: 234–241. Couto, N., Rouboa, A., Silva, V. et al. (2013). Influence of the biomass gasification processes on the final composition of syngas. Energy Procedia 36: 596–606. Pereira, E.G., Da Silva, J.N., de Oliveira, J.L. et al. (2012). Sustainable energy: a review of gasification technologies. Renewable and Sustainable Energy Reviews 16: 4753–4762. McKendry, P. (2002). Energy production from biomass (part 2): conversion technologies. Bioresource Technology 83: 47–54. Hu, M., Guo, D., Ma, C. et al. (2016). A novel pilot-scale production of fuel gas by allothermal biomass gasification using biomass micron fuel (BMF) as external heat source. Clean Technologies and Environmental Policy 18: 743–751. Mun, T., Kang, B., and Kim, J. (2009). Production of a producer gas with high heating values and less tar from dried sewage sludge through air gasification using a two-stage gasifier and activated carbon. Energy & Fuels 23: 3268–3276. Hu, M., Gao, L., Chen, Z. et al. (2016). Syngas production by catalytic in-situ steam co-gasification of wet sewage sludge and pine sawdust. Energy Conversion and Management 111: 409–416. Fonts, I., Gea, G., Azuara, M. et al. (2012). Sewage sludge pyrolysis for liquid production: a review. Renewable and Sustainable Energy Reviews 16: 2781–2805. Syed-Hassan, S.S.A., Wang, Y., Hu, S. et al. (2017). Thermochemical processing of sewage sludge to energy and fuel: fundamentals, challenges and considerations. Renewable and Sustainable Energy Reviews 80: 888–913.

269

270

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

42 Kan, T., Strezov, V., and Evans, T.J. (2016). Lignocellulosic biomass pyrolysis: a review of product properties and effects of pyrolysis parameters. Renewable and Sustainable Energy Reviews 57: 1126–1140. 43 Alvarez, J., Amutio, M., Lopez, G. et al. (2015). Sewage sludge valorization by flash pyrolysis in a conical spouted bed reactor. Chemical Engineering Journal 273: 173–183. 44 Demirbas, A. (2007). The influence of temperature on the yields of compounds existing in bio-oils obtained from biomass samples via pyrolysis. Fuel Processing Technology 88: 591–597. 45 Chen, W., Li, K., Xia, M. et al. (2018). Catalytic deoxygenation co-pyrolysis of bamboo wastes and microalgae with biochar catalyst. Energy 157: 472–482. 46 Balat, M. (2011). An overview of the properties and applications of biomass pyrolysis oils. Energy Sources, Part A: Recovery, Utilization, and Environmental Effects 33: 674–689. 47 Czernik, S. and Bridgwater, A.V. (2004). Overview of applications of biomass fast pyrolysis oil. Energy & Fuels 18: 590–598. 48 Goyal, H.B., Seal, D., and Saxena, R.C. (2008). Bio-fuels from thermochemical conversion of renewable resources: a review. Renewable and Sustainable Energy Reviews 12: 504–517. 49 Mohan, D., Pittman, C.U., and Steele, P.H. (2006). Pyrolysis of wood/biomass for bio-oil: a critical review. Energy & Fuels 20: 848–889. 50 Ahmad, M., Rajapaksha, A.U., Lim, J.E. et al. (2014). Biochar as a sorbent for contaminant management in soil and water: a review. Chemosphere 99: 19–33. 51 Qian, K., Kumar, A., Zhang, H. et al. (2015). Recent advances in utilization of biochar. Renewable and Sustainable Energy Reviews 42: 1055–1064. 52 Antal, M.J. and Grønli, M. (2003). The art, science, and technology of charcoal production. Industrial & Engineering Chemistry Research 42: 1619–1640. 53 Agrafioti, E., Bouras, G., Kalderis, D. et al. (2013). Biochar production by sewage sludge pyrolysis. Journal of Analytical and Applied Pyrolysis 101: 72–78. 54 Gascó, G., Blanco, C.G., Guerrero, F. et al. (2005). The influence of organic matter on sewage sludge pyrolysis. Journal of Analytical and Applied Pyrolysis 74: 413–420. 55 Yuan, J., Xu, R., and Zhang, H. (2011). The forms of alkalis in the biochar produced from crop residues at different temperatures. Bioresource Technology 102: 3488–3497. ´ 56 Zielinska, A., Oleszczuk, P., Charmas, B. et al. (2015). Effect of sewage sludge properties on the biochar characteristic. Journal of Analytical and Applied Pyrolysis 112: 201–213. 57 Cao, X. and Harris, W. (2010). Properties of dairy-manure-derived biochar pertinent to its potential use in remediation. Bioresource Technology 101: 5222–5228. 58 Gaskin, J.W., Steiner, C., Harris, K. et al. (2008). Effect of low temperature pyrolysis conditions on biochar for agricultural use. Transactions of American Society of Agricultural and Biological Engineers 6: 2061–2069. 59 Gray, M., Johnson, M.G., Dragila, M.I. et al. (2014). Water uptake in biochars: the roles of porosity and hydrophobicity. Biomass and Bioenergy 61: 196–205.

References

60 Da˛browski, A., Podko´scielny, P., Hubicki, Z. et al. (2005). Adsorption of phenolic compounds by activated carbon – a critical review. Chemosphere 58: 1049–1070. 61 Saifullah, S., Dahlawi, S., Naeem, A. et al. (2018). Biochar application for the remediation of salt-affected soils: challenges and opportunities. Science of the Total Environment 625: 320–335. 62 Hu, M., Wang, X., Chen, J. et al. (2017). Kinetic study and syngas production from pyrolysis of forestry waste. Energy Conversion and Management 135: 453–462. 63 Deng, J., Li, M., and Wang, Y. (2016). Biomass-derived carbon: synthesis and applications in energy storage and conversion. Green Chemistry 18: 4824–4854. 64 Cyr, M., Coutand, M., and Clastres, P. (2007). Technological and environmental behavior of sewage sludge ash (SSA) in cement-based materials. Cement and Concrete Research 37: 1278–1289. 65 Mahieux, P.Y., Aubert, J.E., Cyr, M. et al. (2010). Quantitative mineralogical composition of complex mineral wastes – contribution of the Rietveld method. Waste Management 30: 378–388. 66 Donatello, S., Tong, D., and Cheeseman, C.R. (2010). Production of technical grade phosphoric acid from incinerator sewage sludge ash (ISSA). Waste Management 30: 1634–1642. 67 Tyagi, V.K. and Lo, S. (2013). Sludge: a waste or renewable source for energy and resources recovery? Renewable and Sustainable Energy Reviews 25: 708–728. 68 Gorazda, K., Tarko, B., Wzorek, Z. et al. (2017). Fertilisers production from ashes after sewage sludge combustion – a strategy towards sustainable development. Environmental Research 154: 171–180. 69 Wong, W.T., Chan, W.I., Liao, P.H. et al. (2006). Exploring the role of hydrogen peroxide in the microwave advanced oxidation process: solubilization of ammonia and phosphates. Journal of Environmental Engineering and Science 5: 459–465. 70 Havukainen, J., Nguyen, M.T., Hermann, L. et al. (2016). Potential of phosphorus recovery from sewage sludge and manure ash by thermochemical treatment. Waste Management 49: 221–229. 71 Rittmann, B.E., Mayer, B., Westerhoff, P. et al. (2011). Capturing the lost phosphorus. Chemosphere 84: 846–853. 72 Weigand, H., Bertau, M., Hübner, W. et al. (2013). RecoPhos: full-scale fertilizer production from sewage sludge ash. Waste Management 33: 540–544. 73 Cie´slik, B. and Konieczka, P. (2017). A review of phosphorus recovery methods at various steps of wastewater treatment and sewage sludge management. The concept of “no solid waste generation” and analytical methods. Journal of Cleaner Production 142: 1728–1740. 74 Ottosen, L.M., Kirkelund, G.M., and Jensen, P.E. (2013). Extracting phosphorous from incinerated sewage sludge ash rich in iron or aluminum. Chemosphere 91: 963–969. 75 Oliver, B.G. and Carey, J.H. (1976). Acid solubilization of sewage sludge and ash constituents for possible recovery. Water Research 10: 1077–1081.

271

272

10 Thermal-Chemical Treatment of Sewage Sludge Toward Enhanced Energy and Resource Recovery

76 Takahashi, M., Kato, S., Shima, H. et al. (2001). Technology for recovering phosphorus from incinerated wastewater treatment sludge. Chemosphere 44: 23–29. 77 Adam, C., Peplinski, B., Michaelis, M. et al. (2009). Thermochemical treatment of sewage sludge ashes for phosphorus recovery. Waste Management 29: 1122–1128. 78 Mattenberger, H., Fraissler, G., Jöller, M. et al. (2010). Sewage sludge ash to phosphorus fertiliser (II): influences of ash and granulate type on heavy metal removal. Waste Management 30: 1622–1633. 79 Tytła, M., Widziewicz, K., and Zielewicz, E. (2016). Heavy metals and its chemical speciation in sewage sludge at different stages of processing. Environmental Technology 37: 899–908. 80 Lester, J.N., Sterritt, R.M., and Kirk, P.W.W. (1983). Significance and behaviour of heavy metals in waste water treatment processes II. Sludge treatment and disposal. Science of the Total Environment 30: 45–83. 81 Gasco, G., Cueto, M.J., and Méndez, A. (2007). The effect of acid treatment on the pyrolysis behavior of sewage sludges. Journal of Analytical and Applied Pyrolysis 80: 496–501. 82 Andrews, J.P., Asaadi, M., Clarke, B. et al. (2006). Potentially toxic element release by Fenton oxidation of sewage sludge. Water Science and Technology 54: 197–205. 83 Li, C., Xie, F., Ma, Y. et al. (2010). Multiple heavy metals extraction and recovery from hazardous electroplating sludge waste via ultrasonically enhanced two-stage acid leaching. Journal of Hazardous Materials 178: 823–833. 84 Pathak, A., Dastidar, M.G., and Sreekrishnan, T.R. (2009). Bioleaching of heavy metals from sewage sludge by indigenous iron-oxidizing microorganisms using ammonium ferrous sulfate and ferrous sulfate as energy sources: a comparative study. Journal of Hazardous Materials 171: 273–278. 85 Peng, G., Tian, G., Liu, J. et al. (2011). Removal of heavy metals from sewage sludge with a combination of bioleaching and electrokinetic remediation technology. Desalination 271: 100–104. 86 Zhu, Y., Zeng, G., Zhang, P. et al. (2013). Feasibility of bioleaching combined with Fenton-like reaction to remove heavy metals from sewage sludge. Bioresource Technology 142: 530–534. 87 Cordero-Lanzac, T., Palos, R., Hita, I. et al. (2018). Revealing the pathways of catalyst deactivation by coke during the hydrodeoxygenation of raw bio-oil. Applied Catalysis B: Environmental 239: 513–524. 88 Okoye, P.U., Abdullah, A.Z., and Hameed, B.H. (2017). A review on recent developments and progress in the kinetics and deactivation of catalytic acetylation of glycerol – a byproduct of biodiesel. Renewable and Sustainable Energy Reviews 74: 387–401. 89 Rodríguez, E., Félix, G., Ancheyta, J. et al. (2018). Modeling of hydrotreating catalyst deactivation for heavy oil hydrocarbons. Fuel 225: 118–133. 90 Han, J. and Kim, H. (2008). The reduction and control technology of tar during biomass gasification/pyrolysis: an overview. Renewable and Sustainable Energy Reviews 12 (2): 397–416.

References

91 Woolcock, P.J. and Brown, R.C. (2013). A review of cleaning technologies for biomass-derived syngas. Biomass and Bioenergy 52: 54–84. 92 Torres, W., Pansare, S.S., and Goodwin, J.G. (2007). Hot gas removal of tars, ammonia, and hydrogen sulfide from biomass gasification gas. Catalysis Reviews 49: 407–456. 93 Cao, J., Li, L., Morishita, K. et al. (2013). Nitrogen transformations during fast pyrolysis of sewage sludge. Fuel 104: 1–6. 94 Zhang, J., Zuo, W., Tian, Y. et al. (2016). Sulfur transformation during microwave and conventional pyrolysis of sewage sludge. Environmental Science & Technology 51: 709–717. 95 Xu, M., Li, S., Wu, Y. et al. (2016). Effects of CO2 on the fuel nitrogen conversion during coal rapid pyrolysis. Fuel 184: 430–439. 96 Tian, F., Li, B., Chen, Y. et al. (2002). Formation of NOx precursors during the pyrolysis of coal and biomass. Part V. Pyrolysis of a sewage sludge. Fuel 81: 2203–2208. 97 Liu, S., Wei, M., Qiao, Y. et al. (2015). Release of organic sulfur as sulfur-containing gases during low temperature pyrolysis of sewage sludge. Proceedings of the Combustion Institute 35: 2767–2775. 98 Seggiani, M., Vitolo, S., Puccini, M. et al. (2012). Cogasification of sewage sludge in an updraft gasifier. Fuel 93: 486–491.

273

275

11 Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge Qilin Wang 1 , Jing Wei 2 , Huan Liu 1 , Dongbo Wang 2 , Long D. Nghiem 1 , and Zhiyao Wang 2 1 Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Ultimo, NSW, Australia 2 Advanced Water Management Centre, The University of Queensland, St. Lucia, QLD, Australia

11.1 Introduction Water utilities generate enormous amounts of sewage sludge globally each year. For example, 3 million wet tonnes of sewage sludge (327 000 dry tonnes) are produced annually in Australia [1] while 400 million wet tonnes (∼50 million dry tonnes) are produced globally [2]. With an ever-growing population and more people connected to the sewerage system, sewage sludge production is growing rapidly. The main methods for sewage sludge disposal have been and still are landfill, agricultural use, and incineration, all incurring very large costs (e.g. $30–65 per wet tonne in Australia and €30–100 per wet tonne in Europe) [3]. With the ever-stricter regulations in sludge disposal, reduced availability of suitable sites and its inherent linkage to transport, the costs for sewage sludge disposal can only rise. However, the sewage sludge itself is rich in organic carbon and represents a substantial energy source, equivalent to an energy content of 3.54 kWh/kg sludge [4]. Therefore, accompanying large sludge production is the loss/wastage of organic carbon that could otherwise be used for bioenergy recovery. Bioenergy recovery from sewage sludge can be achieved via anaerobic sludge digestion in the form of methane or hydrogen. This has been conducted in plenty of sewage treatment plants (STPs) worldwide. Sewage sludge is mainly classified into primary sludge and waste activated sludge. Primary sludge is generally readily biodegradable and therefore has a high bioenergy recovery efficiency during anaerobic digestion [5]. In contrast, waste activated sludge is poorly biodegraded. Therefore, lots of techniques have been developed to enhance bioenergy recovery from waste activated sludge in anaerobic digestion.

Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

276

11 Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge

In this chapter, the characteristics of sewage sludge are first described. Then, as the main bioenergy recovery technique, anaerobic digestion is introduced. Following that, the technologies for enhancing methane and hydrogen production are reviewed, respectively. These technologies are then evaluated and compared. Future outlook has also been put forward at the end of this chapter.

11.2 Characteristics of Sewage Sludge The activated sludge process is extensively used in domestic wastewater treatment. It generates a significant quantity of sewage sludge during primary settling (primary sludge), and chemical oxygen demand (COD) and nutrient removal (WAS, waste activated sludge). This section describes the typical characteristics of both primary sludge and WAS.

11.2.1 Primary Sludge Primary sludge is a result of the capture of suspended solids in the primary treatment process through gravitational sedimentation, typically by a primary clarifier. At this stage, no biological treatment is applied, so the primary sludge is mostly non-biotic organics, which are easily biodegradable as a result of absence of structural protection such as cell walls. When fed into anaerobic digester, primary sludge normally contains higher biogas production potential than WAS. Primary sludge usually contains 93–99.5% of water. Usually the water content of primary sludge could be easily reduced by thickening or dewatering. Fresh primary sludge is a gray or light brown suspension with solids of different sizes. Raw primary sludge particle size distribution is as follows: 5–20% greater than 7 mm; 9–33% 1–7 mm; and 50–88% smaller than 1 mm, of which about 45% is less than 0.2 mm. Table 11.1 lists some typical characteristics of primary sludge in America.

11.2.2 Waste Activated Sludge Waste activated sludge (WAS) comes from the biological treatment (e.g. activated sludge treatment) and secondary clarification in secondary wastewater treatment. It mainly consists of biomass as the product of bacterial growth and extracellular polymeric substances excreted by bacteria. In general, WAS had a lower biodegradability compared with primary sludge and therefore has a lower biogas production performance in the anaerobic digester. WAS generally contains 98–99.5% of water and has a lower dewaterability in comparison to the primary sludge. WAS is light gray or dark brown in color. The size distribution is as follows: 90% below 0.2 mm, 8% between 0.2 and 1 mm, 1.6% between 1 and 3 mm, and 0.4% over 3 mm.

Table 11.1

Summarized results of technologies for enhancing methane production from sludge.

Technology

Sludge

Pretreatment conditions

Digestion conditions

Scale

Results

References

Thermal hydrolysis

MSa)/WAS

121–180 ∘ C; 30–60 min

Semicontinuous, SRTb): 5–20 d; or batch, 7–28 d

Lab/pilot/full-scale

Enhanced CH4 production of 14–90%

[6–18]

Lysis-thickening centrifuge

MS/WAS

2250–3140 rpm

Semicontinuous, SRT: 35–40 d

Full-scale (70 000–150 000 PE)

Enhanced biogas production: 15–26%

[19, 20]

Stirred ball mill

MS/WAS

Ball velocity: 6–15 m/s; ball diameter: 0.25–0.35 mm

Semicontinuous, SRT: 7 d; or batch, 21 d

Lab/full-scale

Enhanced biogas production: 10–21%

[21–23]

High-pressure homogenizer

MS

150–600 bar

Semicontinuous, SRT: 20 d; or batch, 7 d

Lab/full-scale

Enhanced biogas production of 18–64%

[7, 24, 25]

Ultrasonic

MS/WAS

9–41 kHz; 1–150 min

Semi-continuous, SRT: 8–22 d; or batch, 11–100 d

Lab/pilot/full-scale

Enhanced CH4 /biogas production of 24–138%

[26–46]

Microwave

MS/WAS

2450 MHz; 700–1000 W; ∼10 min

Semicontinuous, SRT: 5–25 d; or batch, 18–33 d

Lab-scale

Enhanced CH4 production of 30–84%

[47–53]

Ozonation

MS/WAS

0.05–0.15 g O3 /g TS

Semicontinuous, SRT: 28 d; or batch, 18–35 d

Lab/full-scale

Enhanced CH4 production of 25–110%

[54–60]

2.0 g H2 O2 /g VS; 24 h

Semicontinuous, SRT: 30 d

Lab-scale

Enhanced VS removal: 15%

[61]

pH = 10, 120–130 ∘ C, 30–60 min; or pH = 10, 34–36 ∘ C, 8 d

Semicontinuous, SRT: 20 d; or batch, 7–9 d

Lab-scale

Enhanced CH4 production of 38–340%

[13, 18, 62]

H2 O2 Alkaline

MS WAS

(continued)

Table 11.1

(Continued)

Technology

Free nitrous acid

Sludge

Pretreatment conditions

Digestion conditions

Scale

Results

References

MS/WAS

1.0–2.5 mg HNO2 -N/L; 5–24 h

Batch, 40–44 d

Lab-scale

Enhanced CH4 production of 15–56%

[63–66]

Free ammonia

WAS

85–680 mg NH3 -N/L; 24 h

Batch, 50 d

Lab-scale

Increase CH4 production by 8–22%

[67, 68]

H2

WAS

0.3–0.5 bar

Semicontinuous; SRT = 24 d

Lab-scale

Increase CH4 production by 14–18%

[69]

Scrap iron

WAS

1–33 g/L

Semicontinuous; SRT = 24 d or batch, 20 d

Lab-scale

Increase CH4 production by 31–44%

[70, 71]

0.01 μM to 100 μM

Semicontinuous, SRT = 10 d

Lab-scale

0–3.8 times

[72, 73]

60–70 ∘ C; 9–48 h

Semicontinuous, 13–16 d; or batch, 10 d

Lab/pilot-scale

Enhanced biogas/CH4 production of 26–50%

[74]

Trace metals Biological

MS/WAS WAS

a) Mixed sludge (i.e. primary sludge + WAS). b) Sludge retention time.

11.3 Anaerobic Digestion for Bioenergy Recovery

11.3 Anaerobic Digestion for Bioenergy Recovery 11.3.1 Theory of Anaerobic Digestion Anaerobic digestion (AD) is a well-established process used for bioenergy recovery and for waste stabilization [75, 76]. The AD process also occurs in nature such as in lake sediments, animals’ stomachs, and marshes. In landfills, the organic waste is decomposed by microorganisms in the absence of air through this process. Biogas is given off as the result of AD. Biogas is a methane-rich gas that is produced in the anaerobic digester.” Biogas contains methane (40–75%), carbon dioxide (25–60%), and some other gases [77]. Anaerobic digestion generally comprises four successive stages: hydrolysis, acidogenesis, acetogenesis, and methanogenesis. During the hydrolysis phase, complex organics such as proteins, polysaccharides, and lipids are hydrolyzed to monomers (i.e. glucose, amino acids and fatty acids) by exoenzymes. Then acidogenic bacteria convert the monomers into volatile fatty acids (VFAs) in the second stage (i.e. acidogenesis), together with the evolution of H2 and CO2 . The typical VFAs here are acetic acid, propionic acid, and butyric acid. Only C1 and C2 organics can be used by methanogens; thus, in the next stage long-chain VFAs (C ≥ 3) are further degraded to acetate [78] (i.e. acetogenesis). H2 and CO2 can be converted to acetate as well in the presence of homoacetogens. In the last stage, i.e. methanogenesis, methane is produced in two pathways: about 30% of CH4 comes from H2 and CO2 in the process of hydrogentrophic methanogenesis and 70% of CH4 is derived from acetate in the heterotrophic way.

11.3.2 Bioenergy Recovery by Anaerobic Digestion Anaerobic digestion can produce bioenergy in the form of biogas with CH4 content normally ranging between 40 and 75%. Biogas is an excellent substitute to fossil fuels for producing electricity and heat. AD is usually followed by combined heat and power (CHP) generation to produce heat and electricity to compensate the energy consumption in the sewage treatment plants (STPs). The heat can be in situ used to heat up anaerobic digesters or warm the buildings in or outside the STPs in winter, while the electricity can be used as power supply for blowers in the aerobic treatment unit or other facilities. If the AD process is optimized to enhance methane production, it is totally promising to make an STP energy self-sufficient or even net energy positive [79]. Hydrogen is another potential bioenergy that could be recovered during anaerobic digestion of sewage sludge. H2 is generally viewed as the most promising substitute to fossil fuels due to its high calorific value (142.35 kJ/g) and low environmental impacts. Anaerobic digestion should be controlled to harvest H2 by inhibiting H2 -consuming microorganisms such as hydrogenotropic methanogens and homoacetogens. This could be achieved by inoculum pretreatment using heat, pH, or chemical inhibition [80]. Although practical application of hydrogen production from sewage sludge has not been achieved yet, quite a few bench-scale

279

280

11 Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge

tests have been done, and the feasibility of hydrogen production from sewage sludge has already been proven [81–85].

11.4 Technologies for Enhancing Methane Production from Sludge AD makes it promising to turn an STP from an energy consumer to an energy factory. However, the low energy conversion efficiency limits the wide application of AD. Consequently, currently attention is mainly paid to improving methane production using various technologies. These include physical, chemical, and biological technologies, as described in Sections 11.4.1–11.4.3 and summarized in Table 11.1.

11.4.1 Physical Pretreatment Disruption and hydrolysis of bacterial cells are bottlenecks limiting the efficiency of anaerobic digestion, but suitable pretreatment for sewage sludge could enhance the disruption of cell walls and thus greatly enhance the bioenergy recovery in sludge digestion. The physical pretreatment applies high pressure, high temperature, mechanical shear force or ultrasonic to break down of cell walls and makes the organics inside the cells more accessible to acidogens and methanogens for methane production. Physical pretreatment mainly consists of thermal hydrolysis, mechanical, ultrasonic, microwave, and focused pulsed pretreatment. 11.4.1.1 Thermal Hydrolysis Pretreatment

Thermal hydrolysis utilizes a high temperature (usually 165–180 ∘ C) for 30 minutes to break down the cell walls to release intracellular substrates and break down the complex molecules into simple ones. Using this treatment, the CH4 production was in the range of 14–90%, as summarized in Table 11.1. THP has a bunch of benefits on sludge treatment, including improvement of sludge biodegradability (both sludge degradation rate and sludge degradation extent), increase in biogas yield and methane content, and improvement in sludge dewaterability and hygienization. THP is the most widely used pretreatment and has been applied at full scale since 1995 [12]. HIAS STP (Hamar, Norway) is the first STP that put THP (named Cambi) into application. There are 77 operating THP plants worldwide as of 1 September 2018 (Table 11.2). These plants are sized from 850 dry tonnes (DT) per year to 91 000 dry tonnes per year. The key parameter findings for all THP plants with reported data are summarized in Table 11.3. The key benefits of THP with AD from full-scale plant data revealed that THP could increase feed solids contents to 2–3 times of conventional AD (5–12% TS), cause 41–64% VS reduction, improve dewaterability of solids after digestion to 25–34% TS, and produce higher quality cakes (lower odor, higher solids content, and no pathogen regrowth).

11.4 Technologies for Enhancing Methane Production from Sludge

Table 11.2

Thermal hydrolysis vendor system. Full-scale facilities built

Vendor

Capacities of installed base (DT/day)

Since

Cambi-THP

55

6–360

1995

Veolia-BioTHELYS

7

3–100

2004

Veolia-Exelys

3

10–66

2014

Sustec-TurboTec

2

20–35

2012

Haarslev

2

4–18

2014

CNP-Pondus

6

5–20

2005

Lysotherm

2

3–33

2016

Total

77

3–360

1995

Table 11.3

Key parameter findings for all THP plants with reported data.

Parameter

Range

Average

THP feed %TS

11–22

16.5

Digester feed %TS

5–12

9.3

Digester residence time (d)

17–35

22.3

25–34

29.5

0.745–1.44

1.01

Dewatered cake %TS Steam use (tonnes/t DS) 3

Biogas produced (Nm /t DS) VS reduction (%)

286–650

420

41–64

56.8

11.4.1.2 Mechanical Pretreatment

The mechanical pretreatment applies mechanical shear force to rupture cell walls and thus improve the accessibility of microorganisms to the intracellular substrates. The commonly used mechanical pretreatment includes stirred ball mills, high-pressure homogenization, and lysis-thickening centrifuge technique [22, 86, 87]. The stirred ball mill disintegrates bacterial cells by the shear and pressure forces between the grinding spheres inside the grinding chamber [23]. This technology was firstly applied to the STP in 2002, with the ball diameter and velocity being 0.25–0.35 mm and 6–15 m/s respectively. An increase of biogas yield by 10–21% was successfully achieved in this application (Table 11.1). High-pressure homogenization is also an established pretreatment technology for bioenergy production. Pressure as high as 150–600 bar is required in this technology and normally the biogas yield can be increased by 18–64% (Table 11.1). High-pressure homogenization has been applied at full scale and has been

281

282

11 Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge

patented by several companies, including MicroSludge (Pradigm Environmental Technology), Crown (Biogest), and Cellruptor (Eosolids). Lysis-thickening centrifuge pretreatment is another common mechanical pretreatment for higher biogas production. It is a thickening centrifuge with an additional rotating cutting tool for the rupture of cell structures, and thus the accessibility and degradability of sludge will be improved [88]. Full-scale application of lysis-thickening centrifuge applies the rotation speed between 2250 and 3140 rpm, and the biogas yield could be increased by 15–26% (Table 11.1). 11.4.1.3 Ultrasonic Pretreatment

Ultrasonic vibrations contain very high energy and penetrability and therefore can significantly modify the cell structures. The performance of ultrasonic pretreatment has close relationship with the ultrasonic intensity, frequency, duration, and sludge concentrations. Normally a higher ultrasonic intensity could cause a higher methane production, but the methane production will not increase anymore when the intensity is greater than 7000 kJ/kg TS (dry solid) according to the study of Bougrier et al. [43]. In general, ultrasonic pretreatment is conducted at 9–41 kHz for a few seconds to 150 minutes (usually less than 60 minutes), which could improve CH4 /biogas production by 24–138% (see Table 11.1). Ultrasonic pretreatment does not affect sludge degradation extent but can increase the sludge degradation rate [89]. The enhancement of ultrasonic pretreatment on bioenergy production has already been a proven technique after dozens of years’ study, and it has been put into full-scale application. For example, an increase of 50% of biogas yield was achieved when ultrasonic treatment was applied in a STP in America [90]. Several patented technologies with regard to ultrasonic pretreatment have been granted, such as Iwe.Tec (Germany), Sonix (Sonico, UK), smart DMS (Weber Ultrasonics), Hielscher (Germany), Biosonator (Ultrawaves, Germany), and Sonolyzer (Ovivo). 11.4.1.4 Microwave Pretreatment

Microwave pretreatment is capable of rapidly disintegrating biomass by microwave heating [49, 53]. It generally works between 700 and 1000 W at 2450 MHz for a couple of minutes (Table 11.1). Microwave pretreatment could increase methane yield by 30–84%. Studies on the performance of microwave pretreatment are all about bench-scale tests by now, with no application to full-scale WWTPs. 11.4.1.5 Focused Pulsed Pretreatment

Focused pulsed (FP) pretreatment disintegrates bacterial cells and extracellular polymeric substances (EPS) by applying a high voltage (20–30 kV) and rapid pulse [91, 92]. A significant increase in biogas production (2.5 times) was observed when FP was employed as the pretreatment technology in the study of Choi et al. [91]. Full-scale application also achieved an impressive TS reduction by 40% [92]. FP pretreatment has been patented by PowerMod and OpenCEL.

11.4 Technologies for Enhancing Methane Production from Sludge

11.4.2 Chemical Pretreatment or Dosage Ozonation and alkaline pretreatment are two conventional chemical pretreatment technologies, and there are many other new chemicals being proved effective day by day, including but not limited to free nitrous acid (FNA), free ammonia (FA), iron, H2, and trace metals. 11.4.2.1 Ozonation Pretreatment

Ozonation pretreatment is a widely used oxidation pretreatment for enhancing methane recovery [93]. The normally applied ozone concentration is at 0.05–0.15 g O3 /g TS, which could achieve an increase in methane production by 25–110% (Table 11.1). Full-scale application of ozonation pretreatment has already been set up. 11.4.2.2 Alkaline Pretreatment

The high pH in alkaline pretreatment is able to break down large molecules (carbohydrates, lipids and proteins) into smaller and soluble ones, and thus increase methane production [94]. The typical operation conditions for alkaline pretreatment involve controlling pH at about 10 at 120–130 ∘ C for a short contact time (30–60 minutes) or pH at 10 at 34–36 ∘ C for a couple of days. A large variation in methane production increase (38–340%) was observed due to different operation conditions and sludge characteristics (Table 11.1). Alkaline pretreatment has been tested a lot in lab-scale studies, but no full-scale application has been implemented so far. 11.4.2.3 Free Nitrous Acid Pretreatment

Free nitrous acid (FNA) pretreatment is a novel technology that was developed in 2013. FNA pretreatment was normally conducted at 1.0–2.5 mg HNO2 -N/L for 5–24 hours and the methane production was found to increase by 15–56% (Table 11.1). The improved methane production is due to the increased cell lysis and the break down of extracellular polymeric substances. Similar to THP, FNA pretreatment is reported to be able to improve both sludge degradation rate and sludge degradation extent [64]. As a novel technology, FNA pretreatment has been tested only at lab scale but pilot-scale tests are underway. Unlike other chemicals, FNA is a waste-derived, renewable chemical that can be produced directly from STPs as a by-product of sewage treatment [95]. 11.4.2.4 Free Ammonia Pretreatment

Free ammonia (FA) pretreatment is another innovative technology [67, 96]. It is usually conducted at 85–680 mg NH3 -N/L for 24 hours. Although FA has not been applied to full-scale STP yet, preliminary lab-scale trials have been conducted to prove its feasibility [68]. It was found that FA pretreatment enhances both methane potential and hydrolysis rate of secondary sludge. For instance, the biochemical methane potential increased by 22% (from 160 to 195 L CH4 /kg VS added) while the hydrolysis rate saw an increase of 140% (from 0.22 to 0.53/days) when the secondary sludge was pretreated at 85–680 mg NH3 -N/L for 24 hours. However, the

283

284

11 Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge

FA pretreatment technology is still in its infancy. Only lab-scale batch tests have been conducted. Technology optimization and the fundamental mechanisms still need to be investigated.

11.4.3 Biological Pretreatment Temperature-phased anaerobic digestion (TPAD) is the most widely used biological pretreatment. Based on temperature difference, TPAD is composed of two continuous phases: thermal anaerobic/microaerobic digester (TAD, 60–70 ∘ C, 9–48 hours) and a normal mesophilic anaerobic digester (MAD) phase. The TAD phase works as a pretreatment approach to allow better hydrolysis, acidogenesis, and acetogenesis phase performance. TAD pretreatment is reported to improve biogas production by 26–50% (Table 11.1). Studies indicated TPAD could improve sludge degradation rate but has no effect on the sludge degradation extent [97]. TPAD has already been successfully applied to the STPs.

11.5 Technologies for Enhancing Hydrogen Production from Sludge Exposing the sewage sludge to pretreatment can facilitate hydrogen production in two ways. On the one hand, it can enhance the hydrolysis rate, the rate-limiting step, because most organics in sewage sludge are enclosed by rigid microbial cell walls and not easily accessed by microbes [98, 99]. On the other hand, it can help with selection of hydrogen-producing bacteria. Some hydrogen-producing bacteria can form endospores and thus become resistant to harsh conditions, such as heat, acidity, and alkalinity [100, 101]. Consequently, pretreating sludge under undesirable conditions can probably wash out hydrogen-consuming microbes while retaining hydrogen-producing bacteria. Common pretreatment techniques can be generally divided into physical, chemical, and biological pretreatment (Table 11.4). All these pretreatment technologies are currently at the lab-scale phase.

11.5.1 Physical Pretreatment In general, physical pretreatment utilizes external forces, such as high temperature, high pressure, and shear forces, to disintegrate the cell structure and make the encapsulated organic substances available to hydrogen producers. Physical pretreatment technologies for enhancing hydrogen production include thermal, freezing/thawing, sterilization, microwave, ultrasonic, and Gamma irradiation pretreatment. 11.5.1.1 Thermal Pretreatment

Thermal pretreatment can enhance the hydrolysis step by destroying the sludge floc structure and attacking microbial cell walls, which will then increase the accessibility of intracellular and extracellular organic matter to microbes. Besides, it was

Table 11.4

Summarized results of technologies for enhancing hydrogen production from sludge.

Pretreatment

Physical pretreatment

Sludge

Pretreatment conditions

Digestion conditions

Scale

Results

References

Thermal pretreatment

DSa)/ASb)

65 ∘ C for 30 min

Batch, 200 h

Lab-scale

Enhanced the hydrogen yields to nearly 6 times (1.6–2.3 mol hydrogen/mol glucose)

[102–104]

Freezing/ thawing pretreatment

WASc)

/

Lab-scale

1.5–2.5 times higher hydrogen yield

[105]

Sterilization pretreatment

DS

−17 ∘ C for 24 h and then thawed for another 12 h in a water bath at 25 ∘ C 121 ∘ C for 20 min

Batch, 40 h

Lab-scale

1.5–2.5 times of increase in [106] specific hydrogen yield (30.38 mg H2 /g DS)

Microwave pretreatment

DS

2 min at a power of 560 w

Batch, 40 h

Lab-scale

Poultry slaughterhouse sludge

3 min at a power of 850 w

Serum bottle

Ultrasonic pretreatment

AS

10 s at a power of 100 w

Batch, 30 h; continuous EGSBd) reactor

Lab-scale

Increased 100–700% of the specific hydrogen yield; enhanced the hydrogen yield to 18.28 mg H2 /g DS; enhanced the hydrogen yield to 12.77 mL H2 /g tCOD Elevated the hydrogen yield by 20–32%

Gamma irradiation pretreatment

AS

pH = 12, 20 kGy (60 Co-source)

Batch, 48 h

Lab-scale

Increased the maximum hydrogen production rate and reduced the lag time. (Hydrogen yield: 10.5 ± 0.7 mL/g SCODconsumed )

[106, 107]

[108]

[109, 110]

(continued)

Table 11.4

(Continued)

Pretreatment

Chemical pretreatment

Sludge

Pretreatment conditions

Digestion conditions

Scale

Results

References

Acid pretreatment

Sewage sludge

pH = 3 for 12 h

Serum bottle

Lab-scale

Improved the hydrogen production by 400% (0.86 ± 0.07 mol H2 /mol glucose)

[111]

Alkaline pretreatment

DS

pH = 12 for 24 h

Serum bottle

Lab-scale

Increase the hydrogen production (8.1 mL/g of DS) at pH = 11

[112]

Free nitrous acid pretreatment

WAS

Nitrite 150 mg N/L, pH = 5.5, flushed with N2 for 5 min

Batch, 25 d

Lab-scale

Hydrogen yield was improved by 100% (hydrogen yield of 15.0 ± 0.8 mL/g VSS)

[96]

Free ammonia pretreatment

WAS

Ammonium of Serum bottle 266 mg/L, pH = 9.5, ∘ at 35 C for 24 h

Lab-scale

Produced two times of hydrogen (hydrogen yield of 15.6 ± 0.4 mL/g VSS)

[113]

Ozone pretreatment

WAS

0.07 g O3 /g TS

Lab-scale

Enhanced 17% of biogas production

[42]

Wet oxidation pretreatment

WAS

175 ∘ C for 30 min in Batch, 20 h an autoclave

Lab-scale

Increased the SCOD, [114] polysaccharide, and protein by 2, 2.2 and 102.5 times, promote the hydrogen production

Semi-continuous, 10 d

Biological pretreatment a) b) c) d) e) f)

Calcium peroxide pretreatment

WAS

from 0 to 0.25 g CaO2 /g VSS

Serum bottle

Lab-scale

Increased the hydrogen maximum yield by 13 times (from 0.8 to 10.9 mL/g VSS)

[115]

Triclocarban pretreatment

WAS

from 0 to 1403 ± 150 mg TCCe /kg TSS

Batch, 20 h; semi-continuous, 15 d

Lab-scale

Elevated 50% of the hydrogen yield

[116]

Selected enzymes

Sewage sludge

pH = 6, 8 h, 6.83 g S-TEf /L TSS

Serum bottle

Lab-scale

Improved the hydrogen yield (maximum for 42.90 mL H2 /g VSS)

[117]

Digest sludge. Activated sludge. Waste activated sludge. Expanded granular sludge bed. Triclocarban. Solubilization by thermophilic enzyme.

288

11 Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge

found that thermal pretreatment is an effective way to eliminate methanogens while retaining hydrogenogens [103, 104]. Baghchehsaraee et al. [102] found that heat pretreatment at 65 ∘ C for 30 minutes increased hydrogen yields by approximately five times compared with the untreated sludge. A 15% reduction in hydrogen yield was observed when the pretreatment temperature increased further to 95 ∘ C. 11.5.1.2 Freezing/Thawing Pretreatment

Sewage sludge after the freezing and thawing process was found to release considerable amounts of organic matter [118]. In the study of Hung et al. [118], it was found that hydrogen yield was increased by 1.5–2.5 times after the sludge was frozen at −17 ∘ C for 24 hours and then thawed for 12 hours at 25 ∘ C. 11.5.1.3 Sterilization Pretreatment

Sewage sludge after sterilization was found to release nutrients such as proteins and carbohydrate. Additionally, the sterilized sludge was also effective in adsorbing heavy mental, which was favorable for screening hydrogen-producing bacteria from sludge microflora [119, 120]. The sterilization was generally carried out at 121 ∘ C for 20 minutes, and 1.5–2.5 times of increase in specific hydrogen yield than that of raw sludge could be achieved [106]. 11.5.1.4 Microwave Pretreatment

Microwave radiation is also an effective way to destroy cell walls and release the intracellular substances [121]. The specific hydrogen yield of the pretreated sludge (at 560 W for 2 minutes) was found to increase by 100–700% compared with the untreated sludge [106, 107]. 11.5.1.5 Ultrasonic Pretreatment

Ultrasound can produce microbubbles when propagating in a liquid phase, which can further generate high localized pressure (180 MPa) and temperature (5000 K) as well as highly reactive radicals with the collapse of the bubbles [122]. When a huge number of microbubbles collapse simultaneously, the extreme shearing forces produced are able to rupture sludge cells, change the permeability of sludge, accelerate the release of intracellular substances, and thus enhance the hydrogen production process [123]. Low frequency of ultrasound was well proved to be an effective way of enhancing hydrogen yield from dark fermentation [124, 125]. Hydrogen yield was found to be significantly elevated by 20–32% when the sludge was pretreated by ultrasonic for 10 seconds at 100 W [108]. Meanwhile, the start-up time was also greatly shortened after the ultrasound pretreatment. 11.5.1.6 Gamma Irradiation Pretreatment

Gamma irradiation could destroy cellular structure by the active radicals produced during water radiolysis and thus increase the availability of organic matter encapsulated in the cells [126, 127]. An increase in maximum hydrogen production rate and reduction in lag time were observed in the sludge pretreated by 60 Co-source at varied dosages [109, 110].

11.5 Technologies for Enhancing Hydrogen Production from Sludge

11.5.2 Chemical Pretreatment Chemical pretreatment makes use of special chemicals to enhance cell solubilization and selectively inactivate hydrogen-consuming bacteria. Chemical methods are effective both in enhancing hydrogen production and inhibiting hydrogen consumption. Chemical pretreatment technologies for enhancing hydrogen production include acid, alkaline, free nitrous acid, free ammonia, ozone, wet oxidation, calcium peroxide, and triclocarban pretreatment. 11.5.2.1 Acid Pretreatment

The hydrogen-producing bacteria were proven to be less susceptible to acidic conditions than hydrogen-consuming bacteria [128]. Therefore, acid pretreatment could increase the survival ability of hydrogen-producing bacteria and contrarily suppress the hydrogen-consuming bacteria [129]. The sewage sludge subject to 0.5% (w/v) of hydrochloric acid pretreatment (pH at 3) for 12 hours showed obvious improvement in hydrogen production, with an increase in amplitude of nearly 400% [111]. 11.5.2.2 Alkaline Pretreatment

Alkaline pretreatment not only could attack cell walls and hydrolyze proteins, but also inhibit the growth of hydrogen-consuming anaerobes, thus maintaining stable and high levels of hydrogen production from sewage sludge [130]. Exposing sewage sludge to pH of 12 for 24 hours could enhance the hydrogen production yield by 50% as compared with untreated sludge [112]. 11.5.2.3 Free Ammonia and Free Nitrous Acid Pretreatment

Free ammonia plays roles in both facilitating sludge disintegration and inhibiting hydrogen consumption processes. Wang et al. [113] confirmed that pretreated sewage sludge under ammonium of 266 mg/L and pH 9.5 produced two times as much hydrogen as that of untreated sludge [113]. Similarly, hydrogen yield was improved by 100% under free nitrous acid pretreatment at 150 mg NO2 − -N/L and pH 5.5 [96]. Both free ammonia and free nitrous acid could be generated from the by-product of sewage treatment plants. Consequently, the two pretreatment methods represented a sustainable and close-loop option for hydrogen production enhancement. 11.5.2.4 Ozone Pretreatment

Ozone decomposes itself into radicals, which could react with the sludge, destroy the flocs, and disrupt the compact aggregates [93]. Fermentation of ozonized sludge (at 0.07 O3 /g TS) produced 17% more hydrogen compared with the untreated sludge [42]. Yang et al. [131] combined ozone pretreatment with ultrasound and obtained eight times higher hydrogen yield than control sludge [131]. 11.5.2.5 Wet Oxidation Pretreatment

Low pressure wet oxidation can be used to disintegrate and solubilize the sludge, as well as inactivate microbes consuming hydrogen. The wet oxidation was usually carried out in an autoclave heated to 175 ∘ C for 30 minutes. Increases in SCOD,

289

290

11 Improving Bioenergy Recovery from Anaerobic Digestion of Sewage Sludge

polysaccharide, and protein by 2, 2.2, and 102.5 times were obtained after wet oxidation pretreatment [114]. 11.5.2.6 Calcium Peroxide Pretreatment

Calcium peroxide (CaO2 ) was found to decompose into alkali, •O2 − and •OH radicals, which contributed to the breakage of cells and increased the biodegradability of released substances. In addition, hydrogen consumers were also proved to be susceptible to calcium peroxide pretreatment. Wang et al. [115] verified that the maximum yield was increased by 13 times when the pretreatment dosage increased from 0.05 to 0.25 g CaO2 /g VSS. 11.5.2.7 Triclocarban Pretreatment

Triclocarban (TCC) is a typical pharmaceutical and personal care product (PPCPs). It is to a certain extent toxic to many organisms. It was found that TCC facilitated the processes of solubilization and acidogenesis and severely inhibited the processes of methanogenesis and homoacetogenesis [116]. The varied responses were attributed to the distinct sensitivity of related enzymes to the toxicity of TCC. As a result, hydrogen yield could be elevated by 50%.

11.5.3 Biological Pretreatment Compared with physical and chemical pretreatment methods, biological pretreatment holds advantages of low operation costs and producing less secondary pollution, but the efficiency is generally lower than that of the other two. Biological pretreatment utilizes selected enzymes to help hydrolyze highmolecular weight polymers into low-molecular weight monomers and oligomers, for example the cellulase [132]. Guo et al. [117] also demonstrated that thermophilic enzyme secreted by thermophilic bacteria could stimulate the release of proteins, carbohydrates, and soluble chemical oxygen demand. The hydrogen yield was improved as a consequence.

11.6 Evaluation and Comparison of Technologies AD is a well-established technique to achieve bioenergy recovery and has been widely applied worldwide. Along with the application of AD, various pretreatment technologies are being developed to enhance bioenergy recovery from AD. As summarized in the previous sections, physical, chemical, and biological pretreatment technologies are all feasible strategies for increasing bioenergy recovery in the form of methane or hydrogen production. But each technology has its own advantages and disadvantages. Evaluation and comparison of each technique must be done carefully prior to making any decision. The advantages and disadvantages of pretreatment technologies for enhancing methane production are demonstrated in Table 11.5. In general, ozonation and ultrasonic pretreatment (>4400 kJ/kg TS) would deteriorate the sludge dewaterability

Table 11.5

Advantages and disadvantages of pretreatment technologies for enhancing methane production from sewage sludge.

Approaches

Physical pretreatment

Technology

Advantages

Disadvantages

Thermal pretreatment

Pathogen inactivation; improved sludge dewaterability; low operating cost; process applied at full scale

Odor formation; deterioration of equipment

Ultrasonic pretreatment

No odor formation; process applied at full scale

Erosion of sonotrodes; high operating costs; worsened sludge dewaterability (>4400 kJ/kg TS); no pathogen inactivation

Stirred ball mill

No odor production; short contact time; process applied at full scale

Deterioration of equipment; high investment and operating costs; no pathogen inactivation

High-pressure homogenization

No odor production; short contact time; process applied at full scale

Deterioration of equipment; high investment and operating costs; no pathogen inactivation

Lysis-thickening centrifuge

No odor production; short contact time; process applied at full scale

Deterioration of equipment; high investment and operating costs; no pathogen inactivation

Focused pulsed pretreatment

No odor formation; process applied at full scale

Erosion of electrode; high operating cost; no pathogen inactivation

Microwave pretreatment

Pathogen inactivation; improved sludge dewaterability; less reaction time; and lower thermal loss than thermal pretreatment

Odor formation; high investment and operating costs; only applied at lab scale (continued)

Table 11.5

(Continued)

Approaches

Chemical pretreatment

Biological pretreatment

Technology

Advantages

Disadvantages

Ozonation

No odor formation; process applied at full scale; pathogen inactivation

Worsened sludge dewaterability; high investment and operating costs

Alkaline pretreatment

Improved sludge dewaterability; low investment cost; pathogen inactivation

Corrosion of equipment; high operating cost; only applied at lab scale; potential adverse effect on the subsequent anaerobic digestion

Free nitrous acid

Low investment and operating costs; pathogen inactivation

Only applied at lab scale

Free ammonia

Low investment and operating costs

Only applied at lab scale; only preliminary batch tests were conducted

Temperature-phased anaerobic digestion

Improved sludge dewaterability; pathogen inactivation; low operating cost

High investment costs; odor formation

11.6 Evaluation and Comparison of Technologies

[55, 133]. In contrast, alkaline pretreatment and heat-associated pretreatment (i.e. microwave pretreatment, thermal pretreatment and temperature-phased anaerobic digestion) would improve sludge dewaterability [10, 53, 134]. In terms of pathogen destruction, it could be achieved by ozonation, alkaline, FNA, and heat-associated pretreatment (i.e. thermal pretreatment, microwave pretreatment, and temperature-phased anaerobic digestion) [97, 135, 136]. However, pathogen destruction was not affected by the other technologies. In terms of cost, it is advantageous to employ FNA, FA, thermal pretreatment, and temperature-phased anaerobic digestion, whereas high costs will be required for the other pretreatment technologies. In terms of bioenergy recovery performance, it seems ultrasonic, ozonation, and alkaline pretreatments are more effective in improving bioenergy production (>100% sometimes) compared with the other pretreatment technologies (always 95%, implying that the generation of VFAs from SCOD by acidogens during fermentation is not inhibited by the Al coagulant. The main inhibitory effect by the PACl coagulant on sludge fermentation should occur during the step of organic hydrolysis. Statistical two-factor analysis of variance indicated that the Al dosage had a significant influence on SCOD production (P < 0.05). A similar inhibitory effect was previously reported, where the concentration of VFAs produced by fermentation decreased by 32% for sludge produced by PACl dosed with 12 mg Al/Lsewage [34]. Figure 12.4b depicts the release of phosphate and Al after Al-sludge fermentation. It can be seen that for the control sludge with 0 mg Al/Lsewage , 34.7 mg/L of TP or 17.3 mg/L of PO4 -P was released from the sludge after fermentation. For the Al-sludge, the coagulant dosage of 8 mg Al/Lsewage did not influence the amount of P released, with 17.1 mg/L of PO4 -P being present in the supernatant. However, more P was retained in the solid phase of Al-sludge than in the Al-control sludge (112.1 vs. 6.0 mg/Lsludge ). With a further increase of PACl dosage to 24 mg Al/Lsewage , the TP

315

316

12 Recovery of Phosphorus from Wastewater and Sludge

in the supernatant decreased greatly to 0.15 mg/L after fermentation, showing that little P was released from the sludge under these conditions. The significantly lower P release from the Al-sludge indicates that there was much stronger bonding of P with the Al-based complex than with the Fe-based complex. Al was also released during sludge fermentation at a concentration of 25.5–36.9 mg/L, and the Al release ratio decreased from 16.4 to 3.9% when the Al dosage for CEPS increased from 8 to 24 mg/L. Overall, 8 mg Al/Lsewage was found to be the optimum dosage for organic carbon recovery via sludge fermentation, although further treatment may be needed to leave only ∼2 mg P/L in the primary effluent after CEPS treatment. 12.3.2.3 Recovery of Organic Carbon and P from the Semicontinuous Fermentation of CEPS Sludge

For more practical purposes, semicontinuous reactors were used for the fermentation of CEPS sludge. To achieve satisfactory pollutant removal, 20 mg Fe/L and 16 mg Al/L were selected for CEPS treatment. The sludge was fed daily into the fermenters, and the HRT and SRT of the fermenters were maintained at four days, together with low pH conditions, to prevent the growth of methanogens and thereby avoid the consumption of VFAs [49]. Over the 48 days of the fermentation, the system pH was stable at 5.2–5.3 and 5.7–5.8 for Fe-sludge and Al-sludge, respectively. The Fe-sludge gave an average VFA yield of 364.7 ± 26.5 mg COD/g VS, whereas the yield from the Al-sludge was much lower (204.5 ± 30.4 mg COD/g VS), due to the inhibitory effect of the Al coagulant on sludge hydrolysis within the limited fermentation period. Also, an increasing amount of PO4 -P was released from Fe-sludge, reaching 30 mg P/L in the supernatant, whereas approximately 0.1 mg/L of PO4 -P was detected in the supernatant of the fermented Al-sludge after a slight pH decrease. Greater than 95% of the SCOD content in the fermented CEPS sludge was VFAs, and this type of organic carbon source has been extensively studied for use in denitrification in BNR processes [50, 51] or biosynthesis of PHA [32]. However, the presence of a substantial concentration of PO4 -P in Fe-sludge supernatant would draw P back into the mainstream of BNR, which is undesirable for wastewater treatment. In contrast, the released P in the supernatant of Fe-sludge can be recovered as P fertilizer, which is valuable considering the increasingly depleted reserves of fossil P on Earth [52]. As previously mentioned, such P recovery can be effectively achieved via the simple re-precipitation of P in the form of vivianite [Fe3 (PO4 )2 ⋅8H2 O]. Accordingly, and without any other chemical addition, 99.5% of the released P in the fermented sludge supernatant could be precipitated with the released Fe by simple pH adjustment to 7.5. The X-ray powder diffraction (XRD) and scanning electron microscopy/energy-dispersive X-ray spectroscopy (SEM/EDS) results showed that the major component of the precipitated solids was vivianite, which would have a fertilizing effect similar to that of struvite [53]. Overall, up to 31% of P in the raw sewage was recoverable in the form of vivianite by the Fe-based CEPS and sludge fermentation process. At the same time, the P recovery had no influence on the VFA content in the sludge supernatant. Thus, the supernatant of this process after P recovery was suitable for PHA synthesis and/or BNR.

12.4 An Fe-dosing Membrane Bioreactor for Enhanced P Removal and Recovery

12.3.3 Summary The optimum coagulant dosage for pollutant removal from wastewater by the CEPS process has been found to be 16 mg Al/Lsewage of PACl or 20 mg Fe/Lsewage of FeCl3 . The suspended-pollutant removal efficiency of PACl was approximately 20% greater than that of FeCl3 , whereas FeCl3 was more effective in removing soluble phosphate. FeCl3 addition had little influence on the organic hydrolysis and acidogenesis of the sludge, whereas an inhibitory effect on the release of P was observed at a dosage of 30 mg Fe/Lsewage . Al addition had a significant inhibitory effect on hydrolysis of organics, as attested by its inhibition coefficient of 40.3% at 24 mg Al/Lsewage . Fe-sludge demonstrated a 160.2 mg COD/g VS higher VFA yield than Al-sludge, and substantial PO4 -P release was observed in the semicontinuous fermentation of CEPS sludge, which was attributed to the reduction of Fe3+ to Fe2+ and sludge disintegration under anaerobic conditions. Therefore, FeCl3 was determined to be better than PACl as a coagulant for CEPS-based wastewater treatment, for both P removal and resource recovery.

12.4 A Membrane Bioreactor with Fe Dosing and Sludge Fermentation for Enhanced P Removal and Recovery 12.4.1 Experimental Work An integrated system with an aerobic MBR and side-stream sludge fermentation was developed for enhanced wastewater treatment and P recovery (Figure 12.5). The MBR tank was made of Plexiglass with inner dimensions of 200 × 150 × 400 mm and a working volume of approximately 8 L [54]. A flat-plate ceramic membrane module with an average pore size of 100 nm (Meidensha Corporation, Japan) was vertically submerged in the membrane tank. The membrane plate had dimensions of 240 × 80 × 5.9 mm and an effective surface area of 0.0384 m2 . The MBR tank was filled with activated sludge for biological wastewater treatment with a mixed-liquor suspended solids (MLSS) concentration maintained within a range of 4.0–5.5 g/L. Fe coagulants Treated wastewater

Raw wastewater

Food waste P-rich sludge Feeding tank

Supernatant

Disposal sludge

Return sludge MBR

Precipitation

Fermenter

Figure 12.5 Schematic flow chart of the process comprising chemical dosing, aerobic MBR, and side-stream sludge fermentation for enhanced P removal and recovery.

317

318

12 Recovery of Phosphorus from Wastewater and Sludge

Prior to MBR treatment, the wastewater influent was pre-flocculated by FeCl3 at a dosage of 20 mg Fe/L, and the resulting unsedimented mixture was pumped into the MBR tank. Fe3+ -based chemical dosing was used to induce precipitation and flocculation for enhanced P removal and to achieve a low P concentration in the effluent [4]. Effluent from the MBR was withdrawn through the ceramic membrane by a suction pump (Master FLEX, Cole-Parmer), aiming at a constant flow rate of 16.6 L/day that corresponded to a hydraulic retention time (HRT) of 12 hours and a flux of 17.4 L/h-m2 (LHM). The effluent suction pump was controlled by a programmable logic controller for 5 hours 55 minutes on and 5 minutes off in each operational cycle (4 cycles/day). When the suction pump was off, the membrane module was backwashed with the treated effluent for five minutes at a flux of 46.9 LHM. The transmembrane pressure (TMP) was recorded during the MBR operation; when it exceeded 30 kPa, the membrane module was removed and washed with running tap water to remove the foulants from the membrane surface and restore its permeability. Periodically, e.g. once a month, the ceramic membrane was chemically washed; this involved backwashing of the membrane with 600 mL 0.1 M NaOH (20 mL/min), followed by 600 mL of 0.1 M HCl (20 mL/min), and finally 600 mL of tap water (20 mL/min). Anaerobic reactors were operated for sludge fermentation in a side-stream mode coupling with the MBR (Figure 12.5). Glass bottles (1.76 L) were used as the fermenters, and the activated sludge suspension from the MBR tank was circulated at a predetermined rate (e.g. 1.6 L/day) through the fermenters for P extraction and recovery from the sludge. During the anaerobic fermentation, the sludge underwent hydrolysis and acidogenesis, releasing PO4 -P into the supernatant. The resulting sludge, which had a reduced P concentration, was returned to the MBR. PO4 -P in the supernatant was precipitated with ferrous iron at pH 8, followed by sedimentation for recovery. To examine a possible means of enhancing the sludge acidogenesis and the resulting P release, cooked rice (a model food waste) was added to the sludge fermenter to form a co-fermentative mixture. The rice consisted of approximately 83.4% carbohydrate, 12.8% protein, and 3.8% other components. The cooked rice was pulverized in an electric blender and the resulting solids were mixed well with the sludge in the fermenters. Two fermenters were used to receive the MBR sludge on alternate days. Every day, 1.6 L of the MBR sludge, i.e. 20% of the sludge mixture in the MBR tank, was withdrawn and added to a fermenter to mix with 160 mL of the suspension that consisted of the residual seed sludge and the slurry of cooked rice with an organic content of 3.5 g COD, or 2.0 g COD/L. The sludge was mixed by magnetic stirring in the fermenters at room temperature. After 48 hours of fermentation, 160 mL of the sludge mixture was kept in the fermenter as seed, and the remaining sludge suspension was withdrawn. After 30 minutes of sedimentation, 1 L of supernatant was collected and processed for P recovery. Then, 500 mL of the remaining sludge sediment was returned to the MBR tank and 100 mL was discharged as waste sludge from the system.

12.4 An Fe-dosing Membrane Bioreactor for Enhanced P Removal and Recovery

To precipitate P in the supernatant, the pH of the solution was adjusted to 8 by addition of 2 M NaOH. After stirring for 30 minutes, precipitates were formed, and the mixture was centrifuged for 1 minute to collect these solids, which were then isolated, dried, and analyzed for P content and overall P recovery efficiency.

12.4.2 Results and Discussion 12.4.2.1 P Removal from Wastewater by Chemical Flocculation and MBR

The MBR performed as well as expected in terms of organic degradation and P removal from municipal wastewater (Figure 12.6). As can be seen, the COD removal efficiency was generally >95%, with an effluent COD 97% of PO4 3− -P removed;

Xie et al. [39]

> 90% of NH4 + -N removed Calcium phosphate

> 90% ofPO4 3− -P recovered;

Qiu et al. [37]

99% of NH4 + -N removed 90% of PO4 3− -P recovered

Ansari et al. [40]

> 95% of PO4 3− -P recovered;

Qiu and Ting [36]

98% of NH4 + -N removed

82% of NH4 + -N removed Struvite

> 99% of PO4 3− -P recovered;

Wu et al. [57]

> 93% of NH4 + -N removed Calcium phosphate and/or struvite

65% of PO4 3− -P removed;

Struvite and NH4 HCO3

> 79% of PO4 3− -P removed;

Hou et al. [58]

45% of NH4 + -N recovered Zou et al. [59]

> 99% of NH4 + -N recovered struvite

> 80% of PO4 3− -P removed; > 98% of NH4 + -N removed

Ye et al. [60]

14.3 FO Systems for Nutrient Recovery

FO membrane Wastewater

H2O Draw solution recovery

Mg2+, NH4+ Ca2+, PO43–

Draw solute Feed side

Draw side Nutrientsenriched solution

Recovered nutrients (e.g. struvite and calcium phosphate)

Alkaline chemicals

Figure 14.2

Schematic of FO-based system for nutrient recovery.

the draw solution is needed to ensure enough osmotic pressure between the feed side and draw side. Some researchers proposed pressure-driven membrane processes through the RO membrane or thermally driven membrane processes via the MD membrane to extract fresh water from the draw solute for re-enriching the draw solute [54, 61]. In this scenario, freshwater with high quality could also be produced [39, 62]. As an example, Nguyen et al. [63] used an FO–RO system to obtain 99.9% of phosphate and 92% of ammonium, which were rejected in the feed side of the FO process. Hancock et al. [64] examined this concept at pilot scale, in which >99% of phosphate was enriched in the feed solution and the quality of water permeated from the RO could reach the primary drinking water standards of the Environmental Protection Agency (EPA). It should be noted here that the FO–MD system could use solar energy for heating the MD process [65], which enhances the economic feasibility of the FO-based system in terms of recovering nutrients and clean water [39, 62]. Moreover, seawater can also be used as the draw solution, especially while conducting the FO process near a coastal area. There are two evident advantages: (i) seawater does not need to be regenerated while acting as the draw solute and the diluted seawater can be directly discharged to the sea; and (ii) using the seawater as the draw solute could reduce the costs in terms of raw materials and transport. Besides causing the loss of the draw solute, the reverse draw solute flux may also negatively influence the quality of feed solution [66, 67]. However, it will be beneficial for recovering nutrients if the Mg- and/or Ca-based salts are utilized as the draw solution, which could supplement Mg2+ /Ca2+ ions for the nutrient recovery via chemical precipitation [39]. Through the analysis of scanning electron

379

380

14 Forward Osmosis for Nutrients Recovery from Wastewater

microscopy (SEM) and energy-dispersive X-ray spectroscopy (EDS), however, Wu et al. [57] found that the struvite formed in the FO system was coated by MgCl2 while using the MgCl2 as the draw solute. As a result of this, the purity of struvite was deteriorated. In addition, some contaminants may migrate across the FO membrane to the draw solution, and then accumulate in the side in the closed-loop FO-based system [68]. As an example, Coday et al. [69] found that in an FO–RO system, organic foulants were enriched in the draw solution. Similar findings have also been reported in other FO-based systems [39, 62]. The increase in the cumulative permeate volume may result in higher micropollutant concentrations, which exerts detrimental effect on the FO-based system. Thus, it is of great significance to control the accumulation of contaminants to ensure system performance and reliability. Ansari et al. [54] compared the current wastewater treatment process with FO-based wastewater treatment technologies. They believed that the current conventional wastewater treatment process requires substantial energy supply with reference to the aeration and pressurized membrane systems (Figure 14.3a). Besides, water recovery is the main objective of such a process, but the recovery of chemical energy and nutrients is ignored and they are wasted [71]. Therefore, Ansari et al. [54] proposed an FO-based system for the recovery of nutrients, clean water, and chemical energy stored in the organics (Figure 14.3b). Specifically, the primarily treated effluent is treated by the FO–MD hybrid system to produce clean and freshwater while applying organic ionic solution as the draw solute. As a result of this, the effects of reverse draw solute flux could be reduced along with the probability of the methane inhibition during anaerobic digestion. Subsequently, the FO pre-concentrate is processed with anaerobic digestion for the biogas production, in which the biogas is used for heating the MD process and converting into electricity for the treatment operations. Apart from this, anaerobic effluents containing rich nutrients were treated by an FO–MO hybrid system. As discussed, the nutrient could be recovered by chemical precipitation in such a system [39] while clean water could be produced through the MD system. In addition, the FO technology was integrated with bioreactor in the current wastewater treatment plant. Qiu and Ting [36] proposed an OMBR-based system in the sewage treatment (Figure 14.4), in which >90% of nutrients could be directly recovered by chemical precipitation with additional alkaline chemicals. As discussed, ammonium and phosphate ions were enriched in the feed side of the bioreactor as well as mineral salts (i.e. magnesium and calcium ions). Consequently, it is evident that additional mineral salts may not be needed for recovering nutrients by chemical precipitation. Furthermore, the researchers utilized MgCl2 as the draw solute as the transport of Mg2+ ions caused by reverse draw solute flux could offer more Mg2+ ions for the struvite formation. The selection and enrichment of polyphosphate-accumulating organisms (PAOs) were not carried out in this study, which aims to prevent the phosphate from being converted and adsorbed by the activate sludge. This could both reduce the production of surplus sludge and loss of phosphate consumed by the biological process. Due to the removal of nutrients and mineral salts in the OMBR-based system, it is possible to achieve a moderate

14.3 FO Systems for Nutrient Recovery

Wastewater

(COD) N2, CO2 CO2

Primary treatment

Fe or Al salts Settler 2

MF

RO Disinfection

Anaerobic Anoxic Aerobic

Reuse water

Chemicals Energy

CH4, CO2

Anaerobic digestion

Energy

(a)

Wastewater

Energy Energy

CO2

Chemicals

Sludge dewatering

Biosolids

Energy

Primary treatment

FO

Energy recovery

Combined heat and power generation

MD

Disinfection

Reuse water

Chemicals Energy

FO

Energy Chemicals Sludge dewatering

CO2 CH4, CO2

Energy Anaerobic digestion

MD

(b)

Nutrient recovery

Energy

Energy recovery

Combined heat and power generation

Reuse water

Figure 14.3 Schematic of (a) current wastewater treatment technologies (Source: adapted from Verstraete and Vlaeminck [70]. © Taylor & Francis) and (b) FO-based wastewater treatment technologies (Source: adapted from Ansari et al. [54]). © Elsevier).

Wastewater

H2O

Mg2+, NH4+ Ca2+, PO43–

FO module

Draw solution recovery

H2O

Mg2+, NH4+ Ca2+, PO43–

Struvite calcium phosphate

NaOH

Figure 14.4 Schematic of OMBR-based system for nutrient recovery via chemical precipitation (Source: modified from Qiu and Ting [36]. © Elsevier).

381

382

14 Forward Osmosis for Nutrients Recovery from Wastewater

Wastewater

H2O

H2O Mg2+, Ca2+ PO43–

MF module

PO43

FO module

Draw solution recovery

H 2O

Ca2+ Ca2+ PO43–

Calcium phosphate

NaOH

Figure 14.5 Schematic of OMBR–MF hybrid system for nutrient recovery (Source: modified from Qiu et al. [37]. © American Chemical Society).

level of the system’s salinity, improving the membrane performance. Overall, most nutrients in the feed solution can be recovered by chemical precipitation despite partial nutrients consumed through biological uptake. To increase the purity of recovered nutrients, microfiltration (MF) membrane was installed into the feed side in the OMBR-based system [37] (Figure 14.5). In this case, the nutrients and mineral salts were retained by the FO membrane on the feed side, and then extracted by the MF membrane to the draw side. The MF extraction can not only recover phosphate, but also control the salt accumulation in the feed solution. Through the addition of sodium hydroxide, the calcium phosphate precipitates were formed, realizing the objective of nutrient recovery. In this study, Qiu et al. [37] found the phosphate concentration in the feed side could reach 70 mg/L through the FO retention effect, which results in a 98% phosphate recovery. However, it may be a challenge while applying the OMBR–MF system to wastewater containing high concentrations of nutrients and mineral salts. The possible reason for this is that the risk of spontaneous precipitation in form of struvite, magnesium phosphate precipices, calcium phosphate precipitates, or other precipitates still exists, which negatively affects the performance of both the FO membrane and MF membrane. Besides, this also exerts detrimental impacts on recovering nutrients due to the reduction in the concentration of nutrients and mineral salts. Qiu et al. [37] believed that a higher flow rate of the MF membrane may be an effective strategy. In contrast, Luo et al. [35] claimed that a low permeate flux or the periodic extraction mode of the MF membrane would be more effective for the system because this scenario could reduce the energy input and fouling potential of the MF membrane, and thus, increases this system’s economic feasibility. Based on the OMBR–MF system, Qiu et al. [55] developed a new OMBR system coupled with a fixed bed biofilm (OMBR-BF), in which the fixed bed biofilm replaced the MF membrane. Qiu et al. [55] indicated that there is no extraction of side stream for the solid/liquid separation under this mode, which means the

14.3 FO Systems for Nutrient Recovery

biomass from the FO membrane could be quarantined. Besides, the system could effectively and continuously remove the suspended growth in the absence of the MF membrane. Consequently, the membrane fouling of the FO membrane significantly decreased due to substantially reduced bacterial deposition and colonization. Similarly, Holloway et al. [56] utilized ultrafiltration (UF) membrane to replace the role of the MF membrane in the OMBR–MF system. Up to 50 mg/L of phosphate could be enriched in the UF permeate, which improves the economic feasibility of the phosphate recovery. Furthermore, an RO or MD membrane could be added in the OMBR-based system in order to reconcentrate draw solute of the FO process and recover clean water [35, 72]. For anaerobic FO-based systems, some researchers utilized the FO process to recover nutrients from the anaerobically digested wastewater [33, 73]. For example, Wu et al. [57] employed the FO process to recover nutrients in situ in the form of struvite from digested swine wastewater. They utilized the reverse solute flux, in which 0.5 M MgCl2 was used as a draw solution and the reverse-fluxed Mg2+ ions thereby improved the struvite precipitation [74]. As a result of this, >99% of phosphate and > 93% of ammonium could be recovered/removed. Simultaneously, 3.12 LMH of the water flux could be obtained in the FO system. A preliminary economic analysis was carried out in this study, demonstrating that the total value of recovered products (both struvite and water) was around $1.35/m3 . It should be noted here that the calculations of manpower costs and energy consumption of pump and magnetic stirrer were not included in the assessment. However, the relevant studies on the AnOMBR-based systems used for the nutrient recovery are rare. In 2017, Hou et al. [58] coupled AnOMBR with a microbial recovery cell (MRC) system, in which the current coming from the MRC drove nutrient and mineral salts from the AnOMBR and enriched them into a separate chamber to form nutrient-rich solution (Figure 14.6). The recovery efficiencies of phosphate and ammonium were 65 and 45%, respectively, with simultaneous production of 0.19 L CH4 /g COD. Although the aforementioned FO–MD/RO systems could effectively reconcentrate the draw solute, doubts about the economic analysis exist [57]. These aforementioned systems may consume energy and costs, and it is not sure whether the production of recycled water could offset the economic consumption associated with the draw solute recovery [75]. Besides, it was found that ammonium bicarbonate could be used as a promising draw solute since moderate heating could easily recover the chemical [76]. According to this finding, Qin and He [77] integrated a microbial fuel cell with the FO process (MEC–FO) and used the system for ammonium recovery and wastewater treatment (Figure 14.7). The wastewater was fed into the anode chamber of the MEC, after which the wastewater circulated between the anode chamber of the MEC and the feed side of the FO process. The current applied in the MEC could drive the ammonium transfer from the anode chamber to the cathode chamber, and then the ammonium would be stripped out of the cathode chamber due to the air supply as well as the hydroxyl ions resulted from the cathode reaction of the MEC. Besides, the stripped ammonia reacted with carbon dioxide derived from a power plant to form ammonium bicarbonate, which

383

384

14 Forward Osmosis for Nutrients Recovery from Wastewater

MRC Nutrient recovery solution

Dose solution

Pump

Draw solution

Pump

Pump Digital balance

DS balance

DS digital balance

Gas Gas collector Voltage PC

FO module

MRC/ MEC

T, pH, ORP, E.C

Air pump

Air diffuser

Reactor balance

Pump

Feed synthetic wastewater

Reactor digital balance

Figure 14.6 Schematic of MEC/MRC–AnOMBR system (Source: adapted from Hou et al. [58]. © Elsevier).

Power plant Water air (CO2)

PO43–

Power and resistor

Water heat

NH3 and CO2 NH4HCO3

Fresh water

FO membrane Wastewater NH4+

Anode

NH4+ and OH–

H2O H2O and NH4+

Feed side

Cathode

Draw side

Enriched PO43– Cation-exchange membrane

Air stone Mg 2+ Mg2+ + PO43– + NH4+

Figure 14.7 © Elsevier).

Struvite

MEC–FO system for nutrient recovery (Source: modified from Zou et al. [59].

14.4 Recommended Systems and Key Challenges

was used as the draw solute in the FO process. The waste power from the plant was utilized to heat the draw solution to recover ammonium bicarbonate and extract the freshwater, making the draw solute sustainable. Similarly, Qin et al. [78] also examined the MEC–FO system for ammonium and water recovery from landfill leachate. In this scenario, around 66% of ammonium in the feed solution could be recovered while 51% of water could be recycled from the MEC anode effluent. Zou et al. [59] developed an MEC–FO system for simultaneous recovery of ammonium and phosphate from high-strength side stream centrate. In this scenario, the phosphate was concentrated by the FO process while ammonium was enriched via the MEC. Subsequently, the additional magnesium ions reacted with the nutrients for the struvite precipitation, in which approximately 99.7% of ammonium nitrogen and 79.5% of phosphorus could be recovered. However, they also indicated the current challenges involved in this system, such as optimizing MEC–FO coordination toward making it energy-efficient and reducing nutrient loss.

14.4 Recommended Systems and Key Challenges 14.4.1 Recommended Systems Currently, nutrient recovery at a large scale is often conducted from specific types of wastewater sources containing high concentration of ammonium and phosphate [38]. Due to being effective in separation and concentration of nutrients from wastewater, membrane technologies have been widely utilized to integrate with current nutrient recovery approaches to improve the efficiency of such recovery system [79]. Obviously, the FO membrane process has advantages over other pressure-driven membrane technologies, including: (i) low overall costs due to absence of hydraulic pressure conditions; (ii) high water flux caused by great rejection rate for a wide range of contaminants and low scaling effect; and (iii) low membrane fouling potential compared to pressure-driven membrane processes [80, 81]. As a result of this, the FO-based systems have been widely explored for the nutrient recovery. The nutrient recovery via FO-based systems, in fact, could be divided into two processes: nutrient enrichment and nutrient recovery. In nutrient enrichment, the concentration of ammonium and phosphate is mainly affected by several factors including solution pH and membrane property. The rejection of ammonium and phosphate ions in the FO systems is highly influenced by the pH values of feed solution because the surface of FO membrane could be negatively charged in the alkaline environment (pH > 7) [82]. Therefore, at high pH, the electrostatic attractions between the ammonium ions and FO membrane surface could result in the enhanced adsorption of ammonium ions to the membrane surface, which contributes to the further transfer of ammonium ions to the draw side. This scenario, however, may decrease the number of ammonium ions being recovered in the feed side. On the other hand, the phosphate enrichment in the feed side is enhanced due to the electrostatic repulsion. Apart from this, the existing

385

386

14 Forward Osmosis for Nutrients Recovery from Wastewater

forms of ammonium and phosphate in the aquatic environment are also influenced by the solution pH as shown below [83]: NH4 + ⇌ H+ + NH3 (aq) (pKa = 9.3)

(14.2)

H2 PO4 − ⇌ H+ + HPO4 2− (pKa = 7.2)

(14.3)

Equation (14.2) shows that high pH (> 9.3) could cause the formation of a large proportion of NH3 (aq), which has higher transport rate than NH4 + ions in the FO process according to the Donnan exclusion [38]. It should be noted here that high pH of the feed solution may also convert the ammonium ions into volatile ammonia. Overall, alkaline conditions in the feed solution in FO systems lead to a reduction in the amount of ammonium retained in the feed side. On the other hand, the phosphate retention in the feed side of the FO system could be enhanced at higher pH due to the increased amount of negative charges that the phosphate ions have at the alkaline condition. For this reason, high pH is beneficial for the phosphate enrichment. However, the neutral pH may be more favorable for both phosphate and ammonium enrichment in the FO process. Xue et al. [38] explored the impacts of membrane properties on the rejection rate of nutrients in the FO-based system, in which commercial flat-sheet cellulose triacetate (CTA) membrane and thin-film composite (TFC) membrane were compared. Specifically, the TFC membrane is prepared using polyamide [84, 85] while CTA utilized cellulose triacetate as the active layers with asymmetric structure. In the study, they found that CTA membranes could obtain higher rejection rate for ammonia compared to the TFC membranes. There are two explanations for this: (i) the TFC membrane has higher ammonium permeability than the CTA membrane; and (ii) the TFC membrane has similarly high negative zeta potential to cation-exchange membranes [84, 85], so the transport of ammonium to the draw side across the membrane could be improved, which is attributed to the cation exchange-like mechanism. Compared to the ammonium ions, the phosphate ions are more easily retained in the feed side in FO systems [41, 63]. The possible reason for this is that the negatively charged surface of the FO membrane could retain a large amount of negatively charged phosphate ions due to electrostatic repulsion. Furthermore, phosphate ion has a large hydrated radius (0.339 nm) than that of the ammonium ion (0.104 nm) [86], and the associated sieving effect of the FO membrane could deter the transport of phosphate ions [38, 86]. In addition, the chemical recovery of nutrients enriched by the FO process is mainly influenced by pH values and chemical dose. Due to the pH effects on the species of nutrients and solubility of precipitates [87], the importance of pH value is highlighted, which could affect the quantity and quality of recovered products. High pH (more than 8 at least) is a prerequisite for the chemical precipitation. In contrast, increasing solution pH over 10 may negatively influence the chemical nutrient recovery. This is because ammonium can be converted into volatile ammonia at high pH, which means a decreased amount of ammonium being recovered by struvite precipitation. Apart from this, Mg- and Ca-based salts are always used to react with nutrients for realizing the chemical nutrient recovery, but the metal ions

14.4 Recommended Systems and Key Challenges

Table 14.2

Recommendations for optimizing the FO-based systems for nutrient recovery.

Nutrient enrichment

Nutrient recovery

Factors

Recommendations

pH

Neutral pH kept in the feed side of the FO-based system could facilitate the enrichment of both phosphate and ammonium

Membrane property

The FO membrane with small membrane thickness and high-density active layer can improve the nutrient retainment

Factors

Recommendations

pH

The solution pH in the range of 8–10 could be beneficial for the nutrient recovery via chemical precipitation

Chemical dose

The molar ratio of Mg: N : P for struvite precipitation and Ca: P for calcium phosphate precipitation should be more than 1 : 1 : 1 and 1.67 : 1, respectively

could be transformed into their hydroxides at alkaline conditions. As a result of this, nutrient recovery through chemical precipitation is detrimentally affected. For the struvite precipitation, the theoretical molar ratio of Mg: N : P is 1 : 1 : 1. However, the amount of magnesium used in practical application is larger than the theoretical value since partial Mg2+ ions may react with other ions existing in wastewater [88]. On the other hand, hydroxyapatite precipitation follows a theoretical molar ratio of Ca: P at 1.67 : 1 [89]. More importantly, if such molar ratio is less than 1.67, a decrease in the phosphate recovery efficiency will be detected. As discussed above, Table 14.2 summarized the recommendation of such factors.

14.4.2 Key Challenges Although the versatility and robustness of the FO-based systems for nutrient enrichment and further recovery have been highly accepted [90, 91], this technology is not without limitations. Firstly, one significant hindrance is the membrane fouling and scaling. In the FO-based systems, the nutrient recovery is obtained through chemical precipitation. In this scenario, an excessive amount of nutrients that are close to the surface of the FO membrane may result in the spontaneous precipitation with mineral salts, and the precipitates could be thereby formed on the membranes’ surface. Consequently, the membrane performance is negatively influenced along with the membrane life span. This is despite the fact that chemical cleaning, physical cleaning, and other methods could be employed for membrane scaling. Apart from this, few studies have reported such membrane scaling. The possible reason for this is that the current FO-based systems for the nutrient recovery were operated in a relatively short period. For this reason, the nutrient recovery via FO-based systems should be carried out at pilot scale or even plant scale so the risk of membrane scaling in the system could

387

388

14 Forward Osmosis for Nutrients Recovery from Wastewater

be evaluated, according to which we can settle on relevant strategies for addressing the membrane scaling. In addition, the cake layer of the FO membrane could be formed due to the existence of calcium and phosphate in the feed solution [92], but membrane flushing could be used to remove cake formation [39, 40]. Secondly, the membrane materials in the FO-based system should be given more attention [31] because the water permeability – solute selectivity trade-off is a key factor affecting the performance of the FO membrane [93, 94], especially the rejection rate of nutrients. Specifically, phosphate and ammonium ions could be highly concentrated in the feed solution for the production of nutrient-rich stream, while applying an FO membrane with high solute selectivity effectively. However, the reverse salt flux is negatively affected by this kind of membrane, which may reduce the concentrations of Mg2+ and Ca2+ coming from the draw solution. As a result of this, the nutrient recovery via chemical precipitation is impaired. In contrast, the FO membrane with high water permeability could result in higher reverse salt flux, which supplements Mg2+ and Ca2+ ions for the chemical precipitation. Nevertheless, phosphate and ammonium ions more easily transfer across the membrane to the draw side at the same time, which leads to a significant loss of nutrients. Hence, finding a balance between the water permeability and solute selectivity is critical for recovering nutrients in the FO-based systems.

14.5 Future Roadmap Since phosphate and ammonium can be biologically consumed for the bacterial growth and phosphate can be adsorbed to the sludge through the PAOs, the anaerobic FO-based system for nutrient recovery seems more effective. As discussed, the FO-based systems offer a unique opportunity to achieve nutrient recovery from wastewater. Although the aerobic FO-based systems have presented excellent potential for nutrient enrichment and further recovery [36, 56], the nutrients could be consumed or converted by activated sludge in the aerobic environment; simultaneously, ammonium may be converted into volatile ammonia under aeration supply. As a result of this, the amount of nutrients involved in the nutrient recovery via chemical precipitation could be decreased while the economic feasibility would be detrimentally affected. In contrast, anaerobic FO-based systems can maximize the content of phosphate and ammonium through biologically releasing nutrients and then converting them into soluble forms for further recovery by chemical precipitation. Furthermore, the amount of nutrients consumed by biological uptake is fewer under anaerobic conditions when compared to aerobic conditions. More importantly, the anaerobic FO-based process converts organic substances into methane-rich biogas, which could offset the energy consumption in the recovery system [95]. The easy integration of the anaerobic FO process with current wastewater treatment infrastructure also makes the system more accessible [39, 40]. In conclusion, anaerobic FO-based systems offer more advantages than the aerobic FO-based processes in wastewater treatment, including low energy consumption due to the absence of air supply, significant potential to achieve energy-neutral balance because

14.6 Conclusion

of biogas production, and the high content of nutrients due to anaerobically biological nature [96, 97]. Therefore, more studies are needed to focus on the anaerobic FO-based systems for recovering nutrients in wastewater treatment. As discussed above, the MEC–FO system presented excellent performance for simultaneously recovering ammonium and phosphate from wastewater. The use of microbial fuel cells (MFCs), which have similar functions to MECs, combined with the FO process should be highly encouraged for nutrient recovery. The possible reason for this is that: firstly, the MFC could generate electricity because of a series of bioelectrochemical activity of anaerobic bacteria, which shows less energy input than MECs; and secondly, the solution pH can be increased by the cathode reaction in MFCs without adding alkaline chemicals. Recently, Ye et al. [60] combined the FO membrane with the MFC to form an MFC–FO system, in which 98.81% of NH4 + -N and 83.18% of PO4 3− -P were removed/recovered from municipal wastewater. It was also reported that the current generated in MFCs could reduce the membrane fouling potential [98]. Thus, more investigations on MFC–FO systems would be beneficial for the nutrient recovery in wastewater treatment. In addition, the environmental value of nutrients has not been fully recognized by current economic analysis. More importantly, people still prefer to achieve the nutrients used for the fertilizer production through industrial activities such as mining and the Haber–Bosch process because of mineable phosphate-based rocks, low costs of natural gas, and electricity. In this scenario, the incentives to focus on the nutrient recovery in wastewater treatment are not adequate. Besides, the recovered nutrients (e.g. struvite) are not widely accepted worldwide despite some applications in certain areas of the world; thus, their market values still need more researching. It was reported that the purity and quality of recovered products determine the market value [54], but research conducted to date has not concentrated on the purity of recovered products achieved in the aforementioned FO-based nutrient recovery systems. Overall, the nutrient recovery from wastewater is essential for food security, especially given the rapid rise of global population, decrease of the phosphate-bearing rocks, and high costs for producing industrial ammonium. Therefore, publicity and governmental incentives may act as an effective strategy to enhance the research on nutrient recovery in FO-based systems.

14.6 Conclusion FO technology could highly retain nutrients within reactors with low energy input and membrane fouling potential from a wide range of wastewater sources. The strategic integration of the FO process with other technologies (e.g. chemical precipitation) can harvest valuable nutrients. Compared to aerobic FO-based technologies, anaerobic FO-based systems could better improve nutrient recovery efficiency and produce reusable biogas, which enhances the economic feasibility of the recovery systems. Nonetheless, the challenges such as anaerobic system integration, membrane fouling and scaling, and membrane materials still exist.

389

390

14 Forward Osmosis for Nutrients Recovery from Wastewater

Innovative approaches highlighted in this chapter may be effective to resolve these challenges. Besides, the performance of the FO-based systems for the nutrient recovery needs more investigations into the economic feasibility and applications at the pilot and plant scale to make these systems more viable.

References 1 West, P.C., Gerber, J.S., Engstrom, P.M. et al. (2014). Leverage points for improving global food security and the environment. Science 345: 325–328. 2 Elser, J. and Bennett, E. (2011). Phosphorus cycle: a broken biogeochemical cycle. Nature 478: 29–31. 3 Ashley, K., Cordell, D., and Mavinic, D. (2011). A brief history of phosphorus: from the philosopher’s stone to nutrient recovery and reuse. Chemosphere 84: 737–746. 4 Appl, M. (1999). Ammonia: Principles and Industrial Practice, 245–249. WILEY-VCH Verlag GmbH. 5 Ye, Y., Ngo, H.H., Guo, W. et al. (2018). A critical review on ammonium recovery from wastewater for sustainable wastewater management. Bioresource Technology 268: 749–758. 6 Smith, A. and Klosek, J. (2001). A review of air separation technologies and their integration with energy conversion processes. Fuel Processing Technology 70: 115–134. 7 Santos, F.M. and Pires, J.C. (2018). Nutrient recovery from wastewaters by microalgae and its potential application as bio-char. Bioresource Technology 267: 725–731. 8 Desloover, J., Vlaeminck, S.E., Clauwaert, P. et al. (2012). Strategies to mitigate N2 O emissions from biological nitrogen removal systems. Current Opinion in Biotechnology 23: 474–482. 9 Froehlich, P. (2013). A Sustainable Approach to the Supply of Nitrogen. Parker Balston Corporation. 10 Kool, A., Marinussen, M., and Blonk, H. (2012). LCI data for the calculation tool Feedprint for greenhouse gas emissions of feed production and utilization. GHG Emissions of N, P and K fertiliser production. http://www.blonkconsultants .nl/wp-content/uploads/2016/06/fertilizer_production-D03.pdf (accessed 17 December 2019). 11 Ye, Y., Ngo, H.H., Guo, W. et al. (2017). Insight into chemical phosphate recovery from municipal wastewater. Science of the Total Environment 576: 159–171. 12 Rahman, M.M., Salleh, M.A.M., Rashid, U. et al. (2014). Production of slow release crystal fertilizer from wastewaters through struvite crystallization – a review. Arabian Journal of Chemistry 7: 139–155. 13 Koh, K.Y., Zhang, S., and Chen, J.P. (2020). Hydrothermally synthesized lanthanum carbonate nanorod for adsorption of phosphorus: material synthesis and optimization, and demonstration of excellent performance. Chemical Engineering Journal 380: 122153.

References

14 Ye, Y., Ngo, H.H., Guo, W. et al. (2016). Insight into biological phosphate recovery from sewage. Bioresource Technology 218: 874–881. 15 Ye, Y., Hu, Y., Hussain, Z. et al. (2016). Simultaneous adsorptive removal of fluoride and phosphate by magnesia–pullulan composite from aqueous solution. RSC Advances 6: 35966–35976. 16 Lei, Y., Geraets, E., Saakes, M. et al. (2019). Electrochemical removal of phosphate in the presence of calcium at low current density: precipitation or adsorption? Water Research 169: 115207. 17 Xie, M., Shon, H.K., Gray, S.R. et al. (2016). Membrane-based processes for wastewater nutrient recovery: technology, challenges, and future direction. Water Research 89: 210–221. 18 Hall, M.R., West, J., Sherman, B. et al. (2011). Long-term trends and opportunities for managing regional water supply and wastewater greenhouse gas emissions. Environmental Science and Technology 45: 5434–5440. 19 Whitton, R., Ometto, F., Pidou, M. et al. (2015). Microalgae for municipal wastewater nutrient remediation: mechanisms, reactors and outlook for tertiary treatment. Environmental Technology Reviews 4: 133–148. 20 Kumar, R. and Pal, P. (2015). Assessing the feasibility of N and P recovery by struvite precipitation from nutrient-rich wastewater: a review. Environmental Science and Pollution Research 22: 17453–17464. 21 Li, K., Liu, Q., Fang, F. et al. (2019). Microalgae-based wastewater treatment for nutrients recovery: a review. Bioresource Technology 291: 121934. 22 Sun, Y., Chen, Z., Wu, G. et al. (2016). Characteristics of water quality of municipal wastewater treatment plants in China: implications for resources utilization and management. Journal of Cleaner Production 131: 1–9. 23 Jin, L., Zhang, G., and Tian, H. (2014). Current state of sewage treatment in China. Water Research 66: 85–98. 24 Koppelaar, R. and Weikard, H. (2013). Assessing phosphate rock depletion and phosphorus recycling options. Global Environmental Change 23: 1454–1466. 25 Grant, S.B., Saphores, J.-D., Feldman, D.L. et al. (2012). Taking the “waste” out of “wastewater” for human water security and ecosystem sustainability. Science 337: 681–686. 26 Desmidt, E., Ghyselbrecht, K., Yang, Z. et al. (2015). Global phosphorus scarcity and full-scale P-recovery techniques: a review. Critical Reviews in Environmental Science and Technology 45: 336–384. 27 Cullen, N., Baur, R., and Schauer, P. (2013). Three years of operation of North America’s first nutrient recovery facility. Water Science and Technology 68: 763–768. 28 Pizzol, M., Smart, J.C., and Thomsen, M. (2014). External costs of cadmium emissions to soil: a drawback of phosphorus fertilizers. Journal of Cleaner Production 84: 475–483. 29 Rouff, A.A. and Juarez, K.M. (2014). Zinc interaction with struvite during and after mineral formation. Environmental Science and Technology 48: 6342–6349.

391

392

14 Forward Osmosis for Nutrients Recovery from Wastewater

30 Gerardo, M.L., Aljohani, N.H., Oatley-Radcliffe, D.L. et al. (2015). Moving towards sustainable resources: recovery and fractionation of nutrients from dairy manure digestate using membranes. Water Research 80: 80–89. 31 Lutchmiah, K., Verliefde, A., Roest, K. et al. (2014). Forward osmosis for application in wastewater treatment: a review. Water Research 58: 179–197. 32 Shaffer, D.L., Werber, J.R., Jaramillo, H. et al. (2015). Forward osmosis: where are we now? Desalination 356: 271–284. 33 Soler-Cabezas, J.L., Mendoza-Roca, J.A., Vincent-Vela, M.C. et al. (2018). Simultaneous concentration of nutrients from anaerobically digested sludge centrate and pre-treatment of industrial effluents by forward osmosis. Separation and Purification Technology 193: 289–296. 34 Hafiz, M.A., Hawari, A.H., and Altaee, A. (2019). A hybrid forward osmosis/reverse osmosis process for the supply of fertilizing solution from treated wastewater. Journal of Water Process Engineering 32: 100975. 35 Luo, W., Hai, F.I., Price, W.E. et al. (2016). Phosphorus and water recovery by a novel osmotic membrane bioreactor–reverse osmosis system. Bioresource Technology 200: 297–304. 36 Qiu, G. and Ting, Y.P. (2014). Direct phosphorus recovery from municipal wastewater via osmotic membrane bioreactor (OMBR) for wastewater treatment. Bioresource Technology 170: 221–229. 37 Qiu, G., Law, Y.M., Das, S. et al. (2015). Direct and complete phosphorus recovery from municipal wastewater using a hybrid microfiltration-forward osmosis membrane bioreactor process with seawater brine as draw solution. Environmental Science and Technology 49: 6156–6163. 38 Xue, W., Tobino, T., Nakajima, F. et al. (2015). Seawater-driven forward osmosis for enriching nitrogen and phosphorous in treated municipal wastewater: effect of membrane properties and feed solution chemistry. Water Research 69: 120–130. 39 Xie, M., Nghiem, L.D., Price, W.E. et al. (2014). Toward resource recovery from wastewater: extraction of phosphorus from digested sludge using a hybrid forward osmosis–membrane distillation process. Environmental Science & Technology Letters 1: 191–195. 40 Ansari, A.J., Hai, F.I., Price, W.E. et al. (2016). Phosphorus recovery from digested sludge centrate using seawater-driven forward osmosis. Separation and Purification Technology 163: 1–7. 41 Zhang, J., She, Q., Chang, V.W. et al. (2014). Mining nutrients (N, K, P) from urban source-separated urine by forward osmosis dewatering. Environmental Science and Technology 48: 3386–3394. 42 Moody, C. and Kessler, J. (1976). Forward osmosis extractors. Desalination 18: 283–295. 43 Phuntsho, S., Hong, S., Elimelech, M. et al. (2014). Osmotic equilibrium in the forward osmosis process: modelling, experiments and implications for process performance. Journal of Membrane Science 453: 240–252.

References

44 Xu, W., Chen, Q., and Ge, Q. (2017). Recent advances in forward osmosis (FO) membrane: chemical modifications on membranes for FO processes. Desalination 419: 101–116. 45 Martinetti, C.R., Childress, A.E., and Cath, T.Y. (2009). High recovery of concentrated RO brines using forward osmosis and membrane distillation. Journal of Membrane Science 331: 31–39. 46 Van der Bruggen, B. and Luis, P. (2015). Forward osmosis: understanding the hype. Reviews in Chemical Engineering 31: 1–12. 47 Roy, D., Rahni, M., Pierre, P. et al. (2016). Forward osmosis for the concentration and reuse of process saline wastewater. Chemical Engineering Journal 287: 277–284. 48 Wang, X., Chang, V.W., and Tang, C.Y. (2016). Osmotic membrane bioreactor (OMBR) technology for wastewater treatment and reclamation: advances, challenges, and prospects for the future. Journal of Membrane Science 504: 113–132. 49 Bell, E.A., Holloway, R.W., and Cath, T.Y. (2016). Evaluation of forward osmosis membrane performance and fouling during long-term osmotic membrane bioreactor study. Journal of Membrane Science 517: 1–13. 50 Holloway, R.W., Achilli, A., and Cath, T.Y. (2015). The osmotic membrane bioreactor: a critical review. Environmental Science: Water Research & Technology 1: 581–605. 51 Gu, Y., Chen, L., Ng, J.W. et al. (2015). Development of anaerobic osmotic membrane bioreactor for low-strength wastewater treatment at mesophilic condition. Journal of Membrane Science 490: 197–208. 52 Chen, L., Gu, Y., Cao, C. et al. (2014). Performance of a submerged anaerobic membrane bioreactor with forward osmosis membrane for low-strength wastewater treatment. Water Research 50: 114–123. 53 Zou, S., Yuan, H., Childress, A. et al. (2016). Energy consumption by recirculation: a missing parameter when evaluating forward osmosis. Environmental Science & Technology 50: 6827–6829. 54 Ansari, A.J., Hai, F.I., Price, W.E. et al. (2017). Forward osmosis as a platform for resource recovery from municipal wastewater – a critical assessment of the literature. Journal of Membrane Science 529: 195–206. 55 Qiu, G., Zhang, S., Raghavan, D.S.S. et al. (2016). The potential of hybrid forward osmosis membrane bioreactor (FOMBR) processes in achieving high throughput treatment of municipal wastewater with enhanced phosphorus recovery. Water Research 105: 370–382. 56 Holloway, R.W., Wait, A.S., da Silva, A.F. et al. (2015). Long-term pilot scale investigation of novel hybrid ultrafiltration-osmotic membrane bioreactors. Desalination 363: 64–74. 57 Wu, Z., Zou, S., Zhang, B. et al. (2018). Forward osmosis promoted in-situ formation of struvite with simultaneous water recovery from digested swine wastewater. Chemical Engineering Journal 342: 274–280.

393

394

14 Forward Osmosis for Nutrients Recovery from Wastewater

58 Hou, D., Lu, L., Sun, D. et al. (2017). Microbial electrochemical nutrient recovery in anaerobic osmotic membrane bioreactors. Water Research 114: 181–188. 59 Zou, S., Qin, M., Moreau, Y. et al. (2017). Nutrient-energy-water recovery from synthetic sidestream centrate using a microbial electrolysis cell-forward osmosis hybrid system. Journal of Cleaner Production 154: 16–25. 60 Ye, Y., Ngo, H.H., Guo, W. et al. (2019). Feasibility study on a double chamber microbial fuel cell for nutrient recovery from municipal wastewater. Chemical Engineering Journal 358: 236–242. 61 Chekli, L., Phuntsho, S., Kim, J.E. et al. (2016). A comprehensive review of hybrid forward osmosis systems: performance, applications and future prospects. Journal of Membrane Science 497: 430–449. 62 Xie, M., Nghiem, L.D., Price, W.E. et al. (2013). A forward osmosis–membrane distillation hybrid process for direct sewer mining: system performance and limitations. Environmental Science and Technology 47: 13486–13493. 63 Nguyen, N.C., Chen, S.S., Yang, H.Y. et al. (2013). Application of forward osmosis on dewatering of high nutrient sludge. Bioresource Technology 132: 224–229. 64 Hancock, N.T., Xu, P., Roby, M.J. et al. (2013). Towards direct potable reuse with forward osmosis: technical assessment of long-term process performance at the pilot scale. Journal of Membrane Science 445: 34–46. 65 Qtaishat, M.R. and Banat, F. (2013). Desalination by solar powered membrane distillation systems. Desalination 308: 186–197. 66 Boo, C., Lee, S., Elimelech, M. et al. (2012). Colloidal fouling in forward osmosis: role of reverse salt diffusion. Journal of Membrane Science 390: 277–284. 67 Xie, M., Nghiem, L.D., Price, W.E. et al. (2014). Impact of organic and colloidal fouling on trace organic contaminant rejection by forward osmosis: role of initial permeate flux. Desalination 336: 146–152. 68 D’Haese, A., Le-Clech, P., Van Nevel, S. et al. (2013). Trace organic solutes in closed-loop forward osmosis applications: influence of membrane fouling and modeling of solute build-up. Water Research 47: 5232–5244. 69 Coday, B.D., Almaraz, N., and Cath, T.Y. (2015). Forward osmosis desalination of oil and gas wastewater: impacts of membrane selection and operating conditions on process performance. Journal of Membrane Science 488: 40–55. 70 Verstraete, W. and Vlaeminck, S.E. (2011). ZeroWasteWater: short-cycling of wastewater resources for sustainable cities of the future. International Journal of Sustainable Development and World Ecology 18: 253–264. 71 Wan, J., Gu, J., Zhao, Q. et al. (2016). COD capture: a feasible option towards energy self-sufficient domestic wastewater treatment. Scientific Reports 6: 25054. 72 Chang, H.M., Chen, S.S., Nguyen, N.C. et al. (2017). Osmosis membrane bioreactor–microfiltration with magnesium-based draw solute for salinity reduction and phosphorus recovery. International Biodeterioration and Biodegradation 124: 169–175. 73 Vu, M.T., Price, W.E., He, T. et al. (2019). Seawater-driven forward osmosis for pre-concentrating nutrients in digested sludge centrate. Journal of Environmental Management 247: 135–139.

References

74 Pramanik, B.K., Hai, F.I., Ansari, A.J. et al. (2019). Mining phosphorus from anaerobically treated dairy manure by forward osmosis membrane. Journal of Industrial and Engineering Chemistry 78: 425–432. 75 Achilli, A., Cath, T.Y., and Childress, A.E. (2010). Selection of inorganic-based draw solutions for forward osmosis applications. Journal of Membrane Science 364: 233–241. 76 McCutcheon, J.R., McGinnis, R.L., and Elimelech, M. (2005). A novel ammonia-carbon dioxide forward (direct) osmosis desalination process. Desalination 1740: 1–11. 77 Qin, M. and He, Z. (2014). Self-supplied ammonium bicarbonate draw solute for achieving wastewater treatment and recovery in a microbial electrolysis cell-forward osmosis-coupled system. Environmental Science & Technology Letters 1: 437–441. 78 Qin, M., Molitor, H., Brazil, B. et al. (2016). Recovery of nitrogen and water from landfill leachate by a microbial electrolysis cell-forward osmosis system. Bioresource Technology 200: 485–492. 79 Yan, T., Ye, Y., Ma, H. et al. (2018). A critical review on membrane hybrid system for nutrient recovery from wastewater. Chemical Engineering Journal 348: 143–156. 80 Lee, S., Boo, C., Elimelech, M. et al. (2010). Comparison of fouling behavior in forward osmosis (FO) and reverse osmosis (RO). Journal of Membrane Science 365: 34–39. 81 Mi, B. and Elimelech, M. (2010). Organic fouling of forward osmosis membranes: fouling reversibility and cleaning without chemical reagents. Journal of Membrane Science 348: 337–345. 82 Cartinella, J.L., Cath, T.Y., Flynn, M.T. et al. (2006). Removal of natural steroid hormones from wastewater using membrane contactor processes. Environmental Science and Technology 40 (23): 7381–7386. 83 Qin, J.J., Oo, M.H., Wai, M.N. et al. (2003). Effect of feed pH on an integrated membrane process for the reclamation of a combined rinse water from electroless nickel plating. Journal of Membrane Science 217: 261–268. 84 Yip, N.Y., Tiraferri, A., Phillip, W.A. et al. (2010). High performance thin-film composite forward osmosis membrane. Environmental Science and Technology 44: 3812–3818. 85 Xie, H., Saito, T., and Hickner, M.A. (2011). Zeta potential of ion-conductive membranes by streaming current measurements. Langmuir 27: 4721–4727. 86 Kiriukhin, M.Y. and Collins, K.D. (2002). Dynamic hydration numbers for biologically important ions. Biophysical Chemistry 99: 155–168. 87 Huang, H., Liu, J., and Ding, L. (2015). Recovery of phosphate and ammonia nitrogen from the anaerobic digestion supernatant of activated sludge by chemical precipitation. Journal of Cleaner Production 102: 437–446. 88 Tong, J. and Chen, Y. (2007). Enhanced biological phosphorus removal driven by short-chain fatty acids produced from waste activated sludge alkaline fermentation. Environmental Science and Technology 41: 7126–7130.

395

396

14 Forward Osmosis for Nutrients Recovery from Wastewater

89 Tran, A.T., Zhang, Y., De Corte, D. et al. (2014). P-recovery as calcium phosphate from wastewater using an integrated selectrodialysis/crystallization process. Journal of Cleaner Production 77: 140–151. 90 Ding, Y., Tian, Y., Li, Z. et al. (2014). Characterization of organic membrane foulants in a forward osmosis membrane bioreactor treating anaerobic membrane bioreactor effluent. Bioresource Technology 167: 137–143. 91 Ding, Y., Tian, Y., Liu, J. et al. (2016). Investigation of microbial structure and composition involved in membrane fouling in the forward osmosis membrane bioreactor treating anaerobic bioreactor effluent. Chemical Engineering Journal 286: 198–207. 92 Ning, R.Y. and Troyer, T.L. (2007). Colloidal fouling of RO membranes following MF/UF in the reclamation of municipal wastewater. Desalination 208: 232–237. 93 Yip, N.Y. and Elimelech, M. (2011). Performance limiting effects in power generation from salinity gradients by pressure retarded osmosis. Environmental Science and Technology 45: 10273–10282. 94 Geise, G.M., Park, H.B., Sagle, A.C. et al. (2011). Water permeability and water/salt selectivity tradeoff in polymers for desalination. Journal of Membrane Science 369: 130–138. 95 Smith, A.L., Stadler, L.B., Love, N.G. et al. (2012). Perspectives on anaerobic membrane bioreactor treatment of domestic wastewater: a critical review. Bioresource Technology 122: 149–159. 96 Sills, D.L., Wade, V.L., and DiStefano, T.D. (2016). Comparative life cycle and technoeconomic assessment for energy recovery from dilute wastewater. Environmental Engineering Science 33: 861–872. 97 Seib, M., Berg, K., and Zitomer, D. (2016). Low energy anaerobic membrane bioreactor for municipal wastewater treatment. Journal of Membrane Science 514: 450–457. 98 Wang, Y.K., Li, W.W., Sheng, G.P. et al. (2013). In-situ utilization of generated electricity in an electrochemical membrane bioreactor to mitigate membrane fouling. Water Research 47: 5794–5800.

397

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents from Single-Component and Multicomponent Adsorption Systems S. V. Manjunath and Mathava Kumar Environmental and Water Resources Engineering Division, Department of Civil Engineering, Indian Institute of Technology Madras, Chennai, Tamil Nadu, India

15.1 Introduction to Water Pollution The global water demand has been increasing at a rate of ∼1% annually, i.e. estimated ∼4600 km3 /year and projected to increase ∼6000 km3 /year by the year 2050, due to population intensification and economic development. For instance, in the year 2010 the global groundwater consumption was ∼800 km3 /year, out of which countries including United States, China, India, Pakistan, and Iran extract ∼67% of groundwater [1]. The scarcity of water is increasing globally because of escalation in water demand. It is worth noting that ∼27% of the global population is affected by water scarcity, and it is projected to increase ∼51% in the next three decades [1]. In general, it is estimated that ∼67% of the global population lives in areas that face water scarcity at least one month/year. Approximately 50% of people living in China and India face this level of water scarcity [2]. Subsequently, ∼56% of world’s water is discharged into the environment as sewage, agriculture runoff, and industrial effluents. However, ∼80% of wastewater across the world is discharged into the environment without appropriate treatment [2]. Globally, high-income countries treat ∼70% of wastewater generated, while upper middle-income countries treat ∼38%, lower middle-income countries treat ∼28%, and low-income countries treat only 8% of industrial and municipal wastewater [3]. In general, ∼245 000 km2 of marine ecosystem has been affected due to discharge of untreated wastewater into sea and oceans. Discharge of wastewater without proper treatment into the environment leads to surface and groundwater pollution. As a result, it affects water quality of available freshwater. According to World Health Organization (WHO), ∼842 000 deaths have been caused by consumption of contaminated drinking water and improper sanitation facilities in middle- and low-income countries in 2012. In order to address this problem, wastewater needs to be appropriately managed so as to reduce the water scarcity problem and to mitigate its effect on human health and on the environment.

Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

398

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

15.2 Nutrient Pollution in Aqueous Environment Nutrients are elements that are well known to be essential for biological life. The release of wastewater from various sources such as individual houses, apartments, industries, urban and agricultural runoff, landfill leachates, inappropriately discharged wastes from soak-pits and septic tank leakage are potential sources of nutrient pollution in surface and groundwater bodies. In general, human wastes such as urine contain ∼88% nitrogen and ∼66% phosphorus [4]. Discharge of nitrate (NO3 − ) and phosphate (PO4 3− ) to surface water bodies encourages growth of algal blooms and aquatic weeds, which worsens water quality and reduces dissolved oxygen content in water. As a result, unwarranted growth of algal blooms leads to unesthetic color development, unpleasant odor, increase in turbidity and loss of dissolved oxygen and leads to death of fish and other organism, and finally, eutrophication of water bodies occurs. The potential sources and pathways of nutrient pollution in the environment are shown in Figure 15.1.

15.2.1 Phosphate Pollution Phosphate is a crucial nutrient required for the growth of living organisms. Phosphate is naturally available as mineral deposits, i.e. in the form of phosphate rocks, which are nonrenewable sources. In earth’s crust, availability of phosphate rock is limited and disproportionately distributed. For instance, ∼65% of overall phosphate rocks are available in countries like United States, China, and Morocco [5]. Morocco has domination on Western Sahara’s phosphate rock reserves, and United States has phosphate rock reserve that can survive for three decades. Globally ∼148 million tons of phosphate rock per year are consumed [6]. On the other hand, China is significantly decreasing export of phosphate rock to secure domestic supply. Meanwhile, European countries and India are entirely dependent on imports [7]. As a result, phosphate rock reserves are in control of only a handful of countries, and thus subject to international political influence. Nevertheless, rest of the countries in the world have to import phosphate rocks from these countries. As a result, phosphate availability in a country is interrelated to economy and food security. In general, phosphate is applied to agricultural fields as chemical fertilizers. Due to excessive use of chemical fertilizers in modern agricultural practices, there is an increase in demand for phosphate. Nevertheless, it is estimated that by year 2030, phosphate demand will exceed supply, and phosphate rocks will be depleted within 5–10 decades due to excessive usage in an unsustainable manner [7]. Phosphate has gained worldwide attention mainly because of both scarcity and pollution. Globally, ∼1.5 × 106 tons of phosphate are discharged into freshwater aquatic systems every year from point and nonpoint sources, leading to ecosystem degradation [8]. It is estimated that global anthropogenic phosphorus load to freshwater from Asia is ∼52.8%, followed by ∼19.0% from Europe and ∼12.6% from Latin America and Caribbean. In general, domestic sewage, agricultural runoff, and industrial effluent contribute ∼54, ∼38, and ∼8% of phosphate load to aquatic system, respectively [8]. Phosphorus

15.2 Nutrient Pollution in Aqueous Environment

Figure 15.1

Potential sources and pathways of nutrients into the environment.

is considered as a limiting nutrient as it controls growth of living entities in aquatic systems. As a result, natural phosphorus cycle will be affected by discharge of phosphorus into surface waters, leading to eutrophication of water bodies. In this situation, wastewater is gradually reflected as a likely secondary source for phosphates. Therefore, elimination of phosphates from wastewater is vital to avoid eutrophication in receiving water bodies.

399

400

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

15.2.2 Nitrate Pollution Nitrate pollution sources can be divided into point and nonpoint sources. Excessive application of agricultural fertilizers is the major nonpoint source of nitrate pollution. Agricultural development has increased usage of chemicals globally to ∼2 million tons per year [1]. Organic nitrogen present in soil cannot be used by plants directly, whereas plants mineralize organic nitrogen into ammonium nitrate (NH4 + –N) and utilize it for their growth. Nitrification of NH4 + –N creates nitrate nitrogen (NO3 − –N), which leaches below the root zone and reaches groundwater table and contaminates it [9]. On the other hand, point sources of nitrate pollution include sewage and effluent discharge from industries, accidental spills of nitrogen-rich compounds, absence of slurry storage facilities and manure tanks, concentrated livestock confinement, etc. In addition, percolation of leachate from unscientifically constructed landfill, leakage from septic tanks and sewers, urban runoff, human and animal excreta, and plant wastes are other vital sources of nitrate pollution. In Europe, monitoring stations recorded that ∼30% of rivers and ∼40% of lakes were eutrophic or hypertrophic and ∼15% of groundwater monitoring stations recorded NO3 − concentration in drinking water was higher than WHO standard (45 mg/L) [1]. Consumption of water contaminated with nitrate affects oxygen-carrying capacity of blood from lungs to other body parts. The skin of babies suffering from this syndrome turns blue, leads to methemoglobinemia also called as blue baby syndrome in infants. Symptoms of this syndrome occur rapidly within few days with blueness of skin, difficulty and shortness of breath. Bacteria present in a baby’s mouth and stomach convert nitrate into toxic nitrite. Long-term consumption of water containing a higher concentration of NO3 − has a risk of formation of nitrosamines, which causes development of cancers related to gastric, colorectal, bladder, urothelial, and brain tumor [9]. High NO3 − concentration is harmful to reproductive and respiratory systems, kidney, and thyroid in children and adults. NO3 − poisoning in infants has been considered as a serious concern from the time when the problem was first testified in 1945 [9]. It was reported that consumption of nitrate-contaminated water during early pregnancy may increase risk of specific congenital disabilities. It was testified that NO3 − concentration in groundwater was around 50–100 mg/L in some Indian states, which exceeds the WHO standard for NO3 − in drinking water [9].

15.3 Treatment Technologies for Removal of NO3 − and PO4 3− Many conservative technologies are available for removal of PO4 3− removal, viz., biological methods [10], crystallization [11], electrocoagulation [12], ion exchange [13], membranes [14], precipitation [15], and adsorption. Similarly, for removal of NO3 − ,

15.3 Treatment Technologies for Removal of NO3 − and PO4 3−

universally used treatment methods include catalytic denitrification [16], biological techniques [17], electrocoagulation [18], electrodialysis [19], membranes [20], photocatalysis [21], zerovalent iron [22], and adsorption [23]. The technologies for removal of NO3 − and PO4 3− are relatively costly and cause process complexity during in situ application. Physical methods for removal of PO4 3− are regarded as either too expensive or too inefficient. Meanwhile, chemical precipitation is the most straightforward and flexible technique, which is easy to install and requires less space for removal of nutrients from wastewater at higher efficiency. Moreover, addition of chemicals increases the cost of treatment and generates sludge. Addition of chemicals for precipitation of nutrients requires pH adjustment, and it may not be adequate for wastewater with low NO3 − and PO4 3− concentrations. However, addition of chemicals leads to secondary pollution. On the other hand, end product obtained after crystallization technique can be directly used as a fertilizer for application to agricultural fields. Whereas, crystallization is a sophisticated technique requiring addition of chemicals as well as highly skilled operators. However, this process increases the salinity of effluent obtained after treatment. Removal of NO3 − and PO4 3− using zerovalent iron requires pH adjustment, which is considered to be difficult in real-time application. In general, removal of nutrients by biological techniques is one of the most widely used techniques across the globe. Unlike chemical precipitation and crystallization, biological removal of nutrient does not require addition of chemicals. Nevertheless, biological methods require many parameters for optimum operation; therefore, they are difficult to control. Furthermore, biological processes are difficult to apply to inorganic waste streams as they require addition of organic substrates as electron donors for treatment. Biological removal of NO3 − and PO4 3− is a complex process, which requires high capital and operational cost. Handling and management of sludge, aeration, and an addition of carbon sources incur additional cost and make treatment complex. Also, these systems are highly sensitive to change in nutrient load and change in temperature. The sovereign use of biological techniques cannot respond to existing strict effluent quality standards. Other treatment techniques such as ion exchange and membrane technique are highly efficient when compared to biological technique for nutrient removal and can adapt to different flow rates and concentration of nutrients. On the other hand, ion-exchange and membrane techniques are energy intensive and require high capital and operational cost. Membrane technologies do not resolve problems related to excess NO3 − and PO4 3− in the environment; besides, they generate NO3 − and PO4 3− concentrated waste streams high in saline content and that poses disposal problems. Compared to all the above techniques for removal of NO3 − and PO4 3− , adsorption is the simplest, most cost-effective and proven technology across the world. Low-cost adsorbents can be synthesized from waste sources such as industrial wastes, agricultural wastes, and renewable sources in a sustainable way. No sludge is generated in the process, and thus, handling and management of spent adsorbent used for recovery of NO3 − and PO4 3− can be used for application in agricultural fields as fertilizers.

401

402

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

15.4 Nitrate and Phosphate Recovery Using Low-Cost Adsorbents Activated carbon is utilized as an adsorbent for removal of various classes of pollutants from water and wastewater. For production of activated carbon at commercial scale, most commonly used feedstocks are peat, bituminous coal, lignite, and anthracite. Principally, commercial carbons prepared from coal are expensive when used at large scale. Therefore, there is a continuous search for inexpensive alternatives. This situation compelled researchers toward synthesis of low-cost adsorbents [24]. Figure 15.2 shows different low-cost adsorbents used in the past for removal of various pollutants. Low-cost adsorbents can be obtained from different sources such as industrial wastes, agricultural wastes, biosorbents, natural materials, etc. However, to minimize the overall cost of treatment, many diverse inexpensive adsorbents have been used, such as clay [25], minerals [26], solid waste [27], agricultural wastes [28], and industrial wastes [29]. Kang et al. [30] studied removal of PO4 3− using crushed concrete granules and thermally treated crushed concrete granules. Similarly, Karimaian et al. [31] studied adsorption of PO4 3− using natural pumice, and Kumar et al. [32] studied removal of PO4 3− using coir-pith activated carbon. Subsequently, Zhang et al. [33] studied the removal of NO3 − using MgO-modified peanut shell. Usage of invasive plant biomass for synthesizing activated carbon improves invasive plant management and can be

Figure 15.2

Low-cost adsorbents for removal of various pollutants.

15.4 Nitrate and Phosphate Recovery Using Low-Cost Adsorbents

economical [34]. On the other hand, very few of these adsorbents have been applied for removal of nutrients from multicomponent adsorption systems. For instance, Manjunath and Kumar [23] investigated the removal of NO3 − and PO4 3− using a low-cost adsorbent synthesized from Prosopis juliflora weed in single-component and multicomponent systems.

15.4.1 Factors Affecting Nutrient Adsorption The crucial parameters that affect removal efficiency and adsorption capacity of low-cost adsorbent for nutrient removal include initial adsorbate concentration, adsorbent dose, temperature, contact time, and solution pH. 15.4.1.1 Effect of Initial Concentration

In real-time wastewater treatment, concentration of pollutants present in wastewater will be fluctuating based on several operational and environmental factors. Although, most of adsorbents are effective in removal of pollutants at lower concentrations, adsorbents removal efficiency declines at higher pollutant concentrations. Therefore, it is important to investigate the effect of initial nutrient concentration on sorption process. Subsequently, Manjunath and Kumar [23] studied removal of PO4 3− and NO3 − using Prosopis juliflora activated carbon by increasing PO4 3− and NO3 − initial concentration from 1 to 100 mg/L. The researchers reported that adsorption capacity increased from 0.40 to 20.72 mg/g for PO4 3− and 0.90 to 23.58 mg/g for NO3 − . The increase in adsorbate concentration improves the possibility of collision between nutrient ions in adsorbate solution and adsorbent and that allows PO4 3− and NO3 − ions to overcome mass transfer resistance between liquid and solid phases, thereby improving adsorption capacity. Meanwhile, removal efficiency decreased from 82.7 to 20.7% for PO4 3− when the initial PO4 3− concentration increased from 1 to 100 mg/L. Similarly, removal efficiency decreased from 89.9 to 23.6% for NO3 − . The decline in removal efficiency with increase in initial adsorbate concentration is because of limited number of active adsorption sites on surface of adsorbent, which would have saturated above specific adsorbate concentration. 15.4.1.2 Effect of Adsorbent Dose

Typically, removal efficiency of PO4 3− and NO3 − increases with an increase in adsorbent dose to a specific level, then continues to exist as a constant or decreases marginally with a further rise in dose. The increase in removal efficiency of PO4 3− and NO3 − concerning adsorbent dose was described by large surface area and more number of binding sites available for adsorption of PO4 3− and NO3 − [35]. The decrease in removal of PO4 3− and NO3 − with an increase in adsorbent dose exceeding optimum dose may be due to agglomeration of adsorbent particles, mass transfer resistance, and repulsive forces between active sorption sites on adsorbent surface [36]. Furthermore, Riahi et al. [37] found that as dose of date palm fibers amplified from 2 to 6 g/L, adsorption capacity of PO4 3− improved to 4.69 from 3.75 mg/g. Meanwhile, further rise in adsorbent dose from 6 to 12 g/L decreased PO4 3− adsorption

403

404

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

capacity from 4.69 to 1.75 mg/g. Similarly, Cengeloglu et al. [38] observed that NO3 − adsorption capacity by red mud decreased from ∼80 to ∼20 mmol/g as adsorbent dosage increased from 1 to 8 g/L. This might be due to weak interaction between adsorbate and adsorbent corresponding to agglomeration of adsorbent particles at extreme dosage, thereby reducing surface area of adsorbent. 15.4.1.3 Effect of Temperature

Generally, removal of PO4 3− and NO3 − by low-cost adsorbents is temperature dependent. Some adsorption processes for removal of PO4 3− and NO3 − might be endothermic or exothermic in nature. Mor et al. [39] investigated the effect of temperature for PO4 3− removal using rice husk ash. It was observed that with rise in temper∘ ature from 20 to 50 C, a decline in adsorption of PO4 3− from 75.5 to 55.2% was ∘ ∘ observed. From negative ΔH and ΔG values, it was concluded that adsorption of 3− PO4 was exothermic and spontaneous. Similarly, adsorption of NO3 − using red ∘ ∘ mud was reduced with a rise in temperature from 20 to 50 C [40]. The negative ΔH ∘ (−19.56 kJ/mol) and negative ΔG (−5.17 to −6.51 kJ/mol) specify that the nature of adsorption was exothermic and spontaneous. These observations were in agreement with those testified by Wu et al. [41]. On the other hand, El Bouraie and Masoud [35] found that retention of PO4 3− by raw and Mg(OH)2 -modified bentonite was enhanced as temperature was increased ∘ from 25 to 50 C. It was reported that adsorption process was endothermic and ion-exchange mechanism favored PO4 3− adsorption at higher temperatures. These results were in agreement with the results presented by Wang et al. [42], Pan et al. [43], Kumar and Viswanathan [44] for removal PO4 3− . Similarly, Ohe et al. [45] reported that adsorption of NO3 − using bamboo charcoal decreased with an increase ∘ ∘ ∘ in temperature from 30 to 50 C. From negative ΔH and ΔG , it was concluded that nature of adsorption system was endothermic and spontaneous. These outcomes were in agreement with those testified by Bhatnagar et al. [46]. The rise in temperature leads to surface activation, adsorbent pore enlargement, and an increase in mobility of PO4 3− and NO3 − ions, thereby improving adsorption efficiency [23]. 15.4.1.4 Effect of Contact Time

The contact time is a vital parameter in evaluating performance of an adsorbent as it specifies speed of sorption process. The equilibrium time for PO4 3− adsorption was observed to be five minutes for Mg-treated wood waste biochar [47], 180 minutes for Citrus limetta fruit residue [48], 240 minutes for clays [49], 60 minutes for date palm waste [50], 120 minutes for pine saw dust [51], 24 hours for anaerobically digested sugar beet tailings [52], 24 hours for ferric sludge [53], 60 minutes for calcined eggshell [54], 180 minutes for coir pith activated carbon [32], 120 minutes for metal-loaded skin split waste [55], 10 minutes for wheat residue [56], 12 hours for crab shells [57], 120 minutes egg shell waste [36], 240 minutes for red mud [58], 10 hours for orange waste [59], 20 minutes for blast furnace slag [60], and 60 minutes for Prosopis juliflora activated carbon [23]. On the other hand, equilibrium time for NO3 − adsorption was found to be 120 minutes for modified granular activated carbon [61]; 24 hours for corn cob

15.4 Nitrate and Phosphate Recovery Using Low-Cost Adsorbents

and coconut copra [62]; 60 minutes for sugarcane bagasse [63]; 180 minutes for wheat-straw biochar [64]; 60 minutes for extracts of green tea and eucalyptus leaves [65]; 24 hours for corn stover, switchgrass, and Ponderosa pine wood chips [66]; 480 minutes for clay [67]; 90 minutes for modified rice husk [40]; 180 minutes for sugar beet pulp composite [68]; 10 minutes for wheat straw and mustard straw charcoal [69]; 45 minutes for activated carbon prepared from almond shells [70]; 150 minutes for wheat residue [71]; 60 minutes for red mud [38]; 20 hours for bamboo and coconut shell carbon [45]; and 60 minutes for Prosopis juliflora activated carbon [23]. In field application, shorter contact time is considered to be beneficial as it indicates that the adsorbents need not be placed in adsorption systems for a longer duration, thus minimizing space and cost of treatment system. 15.4.1.5 Effect of pH

The solution pH affects the degree of ionization and speciation of nutrients and surface charge of low-cost adsorbents. It also affects adsorption process through dissociation of functional groups on active sites on adsorbent surface. Subsequently, this leads to a shift in equilibrium characteristics and reaction kinetics of adsorption process. Therefore, a study on effect of solution pH is valuable for optimizing the process, explaining adsorption mechanisms and selecting suitable eluents for desorption [72]. Typically, low-cost adsorbents effectively adsorb PO4 3− and NO3 − in specific pH range. For instance, Hermassi et al. [73] found that red mud adsorbed maximum of 60 mg/g PO4 3− at optimum pH 8 while, increasing or decreasing pH accordingly decreased PO4 3− sorption capacity. Kang et al. [30] increased pH from 5 to 7 and found an increase in adsorption of PO4 3− on concrete granules. Meanwhile, PO4 3− adsorption was unaffected with further rise in pH from 7 to 11. The researchers conveyed that precipitation of PO4 3− on adsorbent surface occured at higher pH. Meanwhile, Mor et al. [39] reported that adsorption of PO4 3− declined with a rise in pH from 2 to 10. This decrease in PO4 3− adsorption at higher pH is probably because of presence of higher concentration of OH− ions in adsorbate solution that competes for adsorption sites with PO4 3− ions. Biswas et al. [59] explained that change in solution pH affects affinity of PO4 3− toward adsorption sites, dissociation constants of PO4 3− , and prevailing PO4 3− ions in medium. The dissociation constants of HPO4 2− , H2 PO4 − , and H3 PO4 are 12.67, 7.21, and 2.12, respectively. Subsequently, adsorption of governing P species will vary subjected to pH of adsorbate solution. On the other hand, Divband Hafshejani et al. [63] reported that adsorption of NO3 − on modified sugarcane bagasse reduced with a rise in solution pH from 2 to 11. This is owing to greater competition between NO3 − and OH− for active sites present on surface of adsorbent. Due to protonation reactions at low pH, modified sugarcane bagasse surface had a positive charge that enhanced electrostatic attraction between adsorbent surface and NO3 − ions. Meanwhile, Kalaruban et al. [62] observed no significant increase in NO3 − removal by coconut copra and corn cub because of net negative charges on surface of coconut copra and corn cub. Katal et al. [40] studied adsorption of NO3 − using modified rice husk in pH range of 3–11. It was observed that maximum adsorption happened at pH 7. At lower pH, the dissociation

405

406

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

of functional groups present on surface of modified rice husk reduced adsorption capacity. The rise in pH tends to intensify the electrostatic force of attraction between adsorbent and NO3 − ions, which in turn increases NO3 − adsorption capacity. At alkaline pH, reduction in adsorption capacity was attributed to antagonism between NO3 − and OH− ions in aqueous medium. At acidic pH, due to protonation, certain functional groups on adsorbent surface are positively charged and tend to bind electrostatically with negatively charged NO3 − . 15.4.1.6 Effect of Interfering Anions

In a real system, several other ions present in wastewater may be competing with PO4 3− and NO3 − . Therefore, it is essential to estimate the effect of other ions present in the system such as Cl− , CO3 2− , SO4 2− , NH4 + , F− ions on PO4 3− and NO3 − adsorption. Investigating coexisting ions’ effect on PO4 3− and NO3 − adsorption is necessary to improve efficiency of adsorption system in field application. Biswas et al. [59] investigated significance of Cl− , CO3 2− , and SO4 2− ions on adsorption of PO4 3− onto La(III)-loaded SOW gel and reported that adsorption of PO4 3− was not substantially influenced by the coexisting anions. Kose and Kivanc [54] studied adsorption of PO4 3− by calcined eggshell and reported that SO4 2− , NO3 − , and NH4 + ions at 10–50 mg/L had no major influence on PO4 3− sorption. Yao et al. [52] reported that Cl− and NO3 − had minimal effect on adsorption of PO4 3− (i.e. 4.3 and 11.7% decrease, respectively) onto anaerobically digested sugar beet tailings, suggesting low competitions between PO4 3− and these two ions. Meanwhile, existence of high concentrations of CO3 2− ions reduced PO4 3− adsorption by about 41.4%. This decrease in adsorption is attributed to competition for adsorption site between CO3 2− and PO4 3− and increase of solution pH. Subsequently, Divband Hafshejani et al. [63] studied adsorption of NO3 − by sugarcane bagasse and reported reduction in adsorption capacity with existence of SO4 2− , Cl− , CO3 2− , and PO4 3− ions. The Cl− ions exhibited maximum and CO3 2− ions showed minimum effect on adsorption of NO3 − . This is because of the fact that electrostatic interaction of coexisting anions with active sites on sugarcane bagasse was stronger when compared with NO3 − ions. Similarly, Kalaruban et al. [62] studied NO3 − adsorption using corn cob and coconut copra in presence of SO4 2− , Cl− , and H2 PO4 − ions. It was observed that NO3 − adsorption reduced drastically due to the existence of SO4 2− and moderately reduced due to the existence of H2 PO4 − and Cl− ions. The authors reported that adsorption occurred through electrostatic attraction between negative charges on anions by outer sphere surface complexation and positive surface charges on adsorbent, i.e. nonspecific adsorption. Meanwhile, Song et al. [74] investigated NO3 − adsorption using modified corn stalk in presence of Cl− , Cr2 O7 2− , SO4 2− , and PO4 3− . The effect of competing ions followed order of PO4 3− > SO4 2− > Cr2 O7 2− > Cl− . The electrostatic interaction of coexisting anions for active sorption sites on the surface of corn stalk was stronger in comparison with NO3 − ions. The researchers found out that lower hydration energy favored ion exchange. The hydration energy of NO3 − (−314 kJ/mol) was lower than SO4 2− (−1103 kJ/mol) and Cl− (−363 kJ/mol). The Jones–Dole coefficient of ions displayed strong correlations with hydration energy, representing that PO4 3−

15.4 Nitrate and Phosphate Recovery Using Low-Cost Adsorbents

hydration energy was greater than hydration energy of SO4 2− , Cr2 O7 2− , Cl− , and NO3 − . Therefore, once high concentration of PO4 3− ions was adsorbed on modified corn stalk, NO3 − ions are difficult to be removed.

15.4.2 Kinetic and Isotherm Modeling Kinetic models help in understanding the rate of reaction and mechanism of adsorption process. Table 15.1 shows expressions of most commonly used kinetic models used for modeling adsorption data. The various isotherm models used to analyze experimental adsorption data are shown in Figure 15.3. 15.4.2.1 Kinetic Models

Generally, pseudo first-order and pseudo second-order models depict rate and order of adsorption process [75]. These models also help to understand whether Table 15.1

Adsorption kinetic models.

Kinetic model

Expression

Plot

Pseudo first-order

ln(qe − qt ) = ln qe − K 1 t

ln(qe − qt )vs t

Pseudo second-order

t 1 t = + qt K2 qe 2 qe

t vs t qt

Intraparticle diffusion

qt = K ID t1/2 + C

qt vs t1/2

Liquid film diffusion

[ ] q ln 1 − e = −KFD t + A qt

[ ] q ln 1 − e vs t qt

Elovich

qt =

Bangham Avarami

Boyd

Vermeulen

1 1 ln 𝛼𝛽 + ln t 𝛽 𝛽

Co Ko m = log + 𝛼 log t Co − qt m 2.303V [ ] qe ln ln = n ln Ka + n ln t qe − qt [ 2 ] qt 𝜋 d 6 = 1 − 2 exp − 2 i t qe 𝜋 r ] [ [ 2 ] ( )2 𝜋 d qt = − 2 it ln 1 − qe r log log

qt vs ln t Co vs log t Co − qt m ] [ qe vs ln t ln ln qe − qt log

qt vs t qe [ ( )2 ] qt vs t ln 1 − qe

Kinetic constants: qe is the equilibrium adsorption capacity (mg/g); qt is the adsorption capacity at time “t” (mg/g); K 1 is the pseudo first-order rate constant (min−1 ), K 2 is the pseudo second-order rate constant (g/mg/min), K ID is the intraparticle diffusion rate constant (mg/g/min0.5 ), and C is the intercept; K FD is the liquid film diffusion rate constant (1/min), A is the liquid film constant, 𝛼 and 𝛽 are Elovich kinetic model constants, K o and 𝛼 are Bangham constants, m is the mass of adsorbent, and V is the volume of adsorbate sample; K a is the Avrami kinetic constant, n is the Avrami model constant; di and r denote intraparticle diffusivity (cm2 /min) and radial distance (cm) from center of particle, respectively.

407

408

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

Figure 15.3

Isotherm models for modeling the experimental data.

adsorption mechanism is physisorption or chemisorption [76]. Meanwhile, models such as intraparticle diffusion, Elovich, and Bangham models depict whether pore diffusion is a rate-limiting step [77]. On the other hand, if experimental data fit liquid film-diffusion model, it indicates that diffusion of adsorbate through liquid film is a rate-controlling step [78]. Tables 15.3 and 15.4 show literature review of various kinetic models for removal of PO4 3− and NO3 − , respectively. Babatunde and Zhao [79] reported that adsorptive removal of PO4 3− using waste alum obeyed intraparticle diffusion model. Meanwhile, Ismail [50] reported that adsorptive removal of PO4 3− using granular date stones and palm surface fiber obeyed pseudo first-order model. Mor et al. [39] investigated PO4 3− removal using rice husk ash and reported that pseudo second-order model described adsorption process. Similarly, Kalaruban et al. [62] investigated PO4 3− removal using amine-grafted corn cob and amine-grafted coconut copra and reported that sorption process was better exemplified by Elovich kinetic model. On the other hand, Hassan et al. [68] studied

15.4 Nitrate and Phosphate Recovery Using Low-Cost Adsorbents

removal of NO3 − using Zr(IV)-loaded sugar beet pulp composite and reported that pseudo first-order model described adsorption system. Meanwhile, Divband Hafshejani et al. [63] reported that adsorptive removal of NO3 − using chemically treated sugarcane bagasse biochar obeyed pseudo second-order model. Similarly, Kalaruban et al. [62] investigated removal of NO3 − using amine-grafted corn cob and amine-grafted coconut copra and reported that adsorption process was better exemplified by Elovich kinetic model. 15.4.2.2 Isotherm Models

Isotherm is the graphical depiction of relationship between amount of adsorbate remaining in a solution and amount of solute adsorbed by unit mass of adsorbent under equilibrium condition at constant temperature. The isotherm models can be categorized into two-parameter isotherms, three-parameter isotherms, four-parameter isotherms, and five-parameter isotherms [80]. Table 15.2 shows equations of several isotherms useful for modeling equilibrium data. Out of all the above categories, two-parameter isotherm models such as Langmuir, Freundlich, Temkin, Elovich, and Dubinin–Radushkevich models are most widely used. Kang et al. [30] studied removal of PO4 3− using crushed concrete granules and thermally treated crushed concrete granules and reported that sorption obeyed Langmuir model with sorption capacity of 21.52 and 8.93 mg/g, respectively. Similarly, Karimaian et al. [31] reported Freundlich isotherm for adsorption of PO4 3− using natural pumice with adsorption capacity of 6.16 mg/g. Meanwhile, Kumar et al. [32] studied removal of PO4 3− using coir pith activated carbon and reported that adsorption process was described by Temkin isotherm with maximum adsorption capacity of 7.74 mg/g. On the other hand, Zhang et al. [33] studied removal of NO3 − using MgO-modified peanut shell and reported that Langmuir model described adsorption process with maximum adsorption capacity of 94 mg/g. Similarly, Chatterjee and Woo [81] reported Sips isotherm represented NO3 − adsorption with adsorption capacity of 89.66 mg/g onto chitosan hydrogel beads. Meanwhile, Wang et al. [82] studied removal of NO3 − using MgO-modified wheat straw and reported that Freundlich isotherm predicted adsorption process better with an adsorption capacity of 2.69 mg/g. Tables 15.3 and 15.4 present literature review of various isotherms for removal of PO4 3− and NO3 − , respectively.

15.4.3 Mechanism of Nutrient Adsorption Adsorption is a surface phenomenon in which adsorbate molecules bind to an adsorbent surface. The transport of adsorbate molecule into adsorbent is illustrated in Figure 15.4. The comprehensive path of adsorption includes mass transfer and involves four steps: Step 1: Transport of adsorbate from bulk phase near adsorbent. Step 2: The diffusion of adsorbate from bulk phase to external film of an adsorbent surface, i.e. film diffusion. Step 3: The diffusion of adsorbate from external surface film into pores of adsorbent, i.e. pore diffusion.

409

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

Table 15.2

Two-parameter models

Type

Adsorption isotherm models. Model

Expression

Langmuir

1 1 1 = + qe qm qm KL Ce

Freundlich

ln qe = ln KF + 1∕n ln Ce

Temkin

qe = B ln AT + B ln Ce

Dubinin–Radushkevich

lnqe = ln(qm ) − (K DR 𝜀2 )

Elovich

ln

Flory–Huggins

θ = KFH (1 − 𝜃)nFH Co

Fowler–Guggenheim

KFG Ce =

qe Ce

= ln KE qm −

θFG

1 q qm e

(

exp

1 − θFG

2θFG W RT

K1H Ce =

Hill–de Boer

K1

Jovanovic

qe = qm (1 − e j e ) ) ( lnKH − ln Ce qe = exp nH )1∕2 ( AH qe = B2 − log Ce

Harkin–Jura

H

(1 − θK )(1 + Kn θK ) ) ( θH θH K θ Ce = exp − 2 H 1 − θH 1 − θH RT KC

Redlich–Peterson

qe =

Hill

qe =

Toth

qe =

Jossens

Ce =

Fritz–Schlunder

qe =

Sips

qe =

Kobel–Corrigan

qe =

Khan

qe =

Radke–Prausnitz

qe =

Frumkin Liu

)

θK

Kiselev

Halsey

Three-parameter models

410

KRP Ce 1 + αRP Ce g qS Ce nH H

KD + Ce nH KT Ce 1 [AT + Ce TT ] ∕TT

qe

p

exp(Fqe ) H qm KFS Ce FS

1 + qm Ce mFS qm KS Ce ms 1 + KS Ce ms αCe n 1 + bCe n qs bK Ce (1 + bK Ce )nK qm

RPI

KRPI Ce

(1 + KRPI Ce )mRPI ) ( −f 𝜃 KFF Ce = 𝜃exp 1−θ Qe =

Qmax .Kg Ce nL 1 + Kg Ce nL

15.4 Nitrate and Phosphate Recovery Using Low-Cost Adsorbents

(Continued) Model

Expression

Weber-van-Vliet

Ce = P1 qe (P2 qe

Fritz–Schlunder

qe =

Baudu

KL = b0 Ce x Q0 = qm0 Ce y

Fritz–Schlunder

qe =

Five-parameter models

Type Four-parameter models

Table 15.2

P3 +P ) 4

ACe ∝ 1 + BCe β

qm

FS5

K1 Ce m1

1 + K2 Ce m2

Isotherm constants: qe is the equilibrium adsorption capacity (mg/g); qm is the maximum adsorption capacity (mg/g); K L is the Langmuir constant (L/mg); Ce is the equilibrium concentration (mg/L), K F and n are Freundlich constants; AT is the Temkin equilibrium constant (L/mg) and B is the Temkin constant; 𝛥Q is the heat of adsorption; K DR is the mean free energy of adsorption (mol2 /kJ2 ); and 𝜀 is the Polyani potential; K E is the Elovich equilibrium constant (L/mg); K FG is the Fowler–Guggenheim equilibrium constant (L/mg); K 1K is the Kiselev equilibrium constant (L/mg); K 1H is the Hill–de Boer constant (L/mg); K J is the Jovanovic isotherm constant (L/mg); K H represents the Halsey isotherm constant; AH is the Harkins–Jura isotherm parameter; K RP is the Redlich–Peterson model isotherm constant (L/g); qsH is the Hill isotherm maximum uptake saturation (mg/L); H, F, and p represent parameters of Jossens equation; qmFS is the Fritz–Schlunder maximum adsorption capacity (mg/g); qmS is the Sips adsorption capacity (mg/g); a (Lnmg1−n /g), b (L/mg)n , and n are Koble–Corrigan parameters; bK is the Khan isotherm model constant; P1 , P2 , P3 , and P4 are isotherm parameters of Weber–van Vliet model; A and B are Fritz–Schlunder parameters; and a and b are Fritz–Schlunder equation exponents; qmB is the Baudu maximum adsorption capacity (mg/g).

Step 4: Entrapment/attachment of adsorbate to internal surface of adsorbent, i.e. physisorption or chemisorption. The first three steps are transportation steps while step 4 is a sorption step. The rate of adsorption is subjected to total resistance offered by all four steps in series, which is sum of three component resistances in series that is measured by fitting experimental data into kinetic and isotherm models. The decrease in resistance offered by any of steps leads to an increase in adsorption rate. Step 1 and step 4 are very rapid compared to step 2 and step 3. Therefore, resistance offered by these steps is negligible [112], whereas film diffusion (step 2) and pore diffusion (step 3) are considered to be rate-controlling steps in sorption process. In general, two types of interactions take place among adsorbent and adsorbate namely, physisorption and chemisorption. Physisorption is due to weak attractive forces, i.e. van der Walls force of attraction is responsible for physical adsorption between adsorbent and adsorbate molecules. The heat of adsorption ranging from 5 to 40 kJ/mol specifies that adsorption is dominated by physisorption process. Whereas, chemisorption is because of a stronger bond involving sharing or transfer of electrons between adsorbent and adsorbate. The heat of adsorption ranging from 40 to 125 kJ/mol designates that adsorption is dominated by chemisorption process.

411

Table 15.3

A literature review of various adsorbents for removal of PO4 3− .

Adsorbent

Kinetic model

Isotherm model

Adsorption capacity

Thermodynamic characteristic

References

Ferrihydrite

Pseudo second-order

Langmuir

66.6 mg/g

Endothermic, spontaneous

[83]

Goethite

Pseudo second-order

Freundlich

2.07 mg/g

Endothermic, spontaneous

Magnetite

Pseudo second-order

Freundlich

2.04 mg/g

Endothermic, spontaneous

Aluminum hydroxide-modified palygorskite nanocomposite

Pseudo second-order

Freundlich

2.09 mg/g

Endothermic, spontaneous Endothermic, spontaneous

Natural palygorskite nanocomposite

Pseudo second-order

Freundlich

0.21 mg/g

Crushed concrete granules

Pseudo second-order

Langmuir

21.52 mg/g

Thermally treated crushed concrete granules

Pseudo second-order

Langmuir

8.93 mg/g

Modified bentonite with Mg(OH)2

Langmuir

3.435 mg/m2

Endothermic, spontaneous

Raw bentonite

Langmuir

0.445 mg/m2

Endothermic, spontaneous

30.4 mg/g

[43]

[30]

[35]

Chinese reed

Pseudo second-order

Langmuir

Lanthanum loaded biochar

Pseudo second-order

Langmuir

46.37 mg/g

Endothermic, spontaneous

[42]

Modified chitosan beads

Pseudo second-order

Langmuir

60.6 mg/g

Endothermic, spontaneous

[84]

OH–Al pillared bentonite

Pseudo second-order

Langmuir

12.7 mg/g

Endothermic, spontaneous

[85]

OH–Fe pillared bentonite

Pseudo second-order

Langmuir

11.2 mg/g

Endothermic, spontaneous

OH–Al–Fe pillared bentonite

Pseudo second-order

Langmuir

10.5 mg/g

Endothermic, spontaneous

Natural pumice

Pseudo second-order

Freundlich

6.16 mg/g

Modified pumice

Pseudo second-order

Freundlich

9.53 mg/g

[31]

Apple peel

Pseudo second-order

Langmuir

20.35 mg/g

[86]

Modified sugarcane bagasse

Pseudo second-order

Langmuir

30.67 mg/g

[87]

Modified clinoptilolite

Elovich

Langmuir

1.972 mg/g

Endothermic, spontaneous

[88]

Boehmite

Pseudo second-order

Langmuir

11.47 mg/g

Endothermic, spontaneous

[89]

Pine sawdust

Pseudo second-order

Freundlich

12.86 mg/g

Endothermic, spontaneous

[51]

1.164 × 104 mg/kg

Anaerobically digested sugar beet tailings

Pseudo first-order

Freundlich

Ferric sludge

Pseudo second-order

Multi-Langmuir

36.67 mg/g

[53]

Calcined waste eggshell

Pseudo second-order

Freundlich

23.02 mg/g

[54] [90]

Magnetic biochar

[52]

Langmuir

1.24 mg/g

Waste alum sludge

Intraparticle diffusion

Langmuir

31.9 mg/g

Coir pith activated carbon

Intraparticle diffusion

Temkin

7.79 mg/g

Metal-loaded skin split waste

Pseudo second-order

Langmuir

72 mg/g

[55]

Scallop shell

Langmuir

23 mg/g

[91]

Crab shell

Langmuir

108.2 mg/g

[79] Exothermic, spontaneous

[32]

[57]

Iron hydroxide–eggshell waste

Pseudo second-order

Langmuir

14.49 mg/g

Endothermic, spontaneous

[36]

Aluminum pillared clay

Pseudo first-order

Freundlich

1.21 mg/g

Spontaneous, exothermic

[92]

Mixed La/Al pillared clay

Pseudo first-order

Spontaneous, exothermic

Freundlich

1.72 mg/g

Natural sand coated with Fe oxide

Langmuir

0.88 mg/g

Spontaneous, endothermic

Synthetic sand coated with Fe oxide

Langmuir

1.5 mg/g

Spontaneous, endothermic

Langmuir

1.8 mg/g

Spontaneous, endothermic

Langmuir

13.94 mg/g

[59]

Langmuir

155.2 mg/g

[58]

Crushed brick coated with Fe oxide Metal-loaded orange waste

Pseudo second-order

Heat-treated red mud Acid-treated red mud

[93]

Langmuir

202.9 mg/g

[58]

La-treated Juniperus monosperma bark

Pseudo second-order

Langmuir

0.30 mmol/g

[94]

Juniper fiber precipitated with acid mine drainage

Pseudo second-order

Freundlich

1.76 mg/g

[95]

Blast furnace slag

Freundlich

6.37 mg/g

[60]

Soybean hull resin

Langmuir

0.63 mmol/g

[96]

Langmuir

1.82 mg/g

Stem bark of Eucalyptus tereticornis Smith

Reversible first order

Spontaneous, endothermic

[97]

Table 15.4

A literature review of various adsorbents for removal of NO3 − .

Isotherm model

Adsorption capacity

Adsorbent

Kinetic model

Thermodynamic characteristic

Zr(IV)-loaded sugar beet pulp composite

Pseudo first-order

Langmuir

63 mg/g

[68]

Sodium hydroxide and cationic surfactant-modified activated carbon

Pseudo second-order

Langmuir

21.51 mg/g

[61]

Sugarcane bagasse (chemically modified)

Pseudo second-order

Langmuir

28.21 mg/g

Amine-grafted corn cub

Elovich

Freundlich

1.24 mg/g

Amine-grafted coconut copra

Elovich

Freundlich

2.07 mg/g

MgFe-layered double hydroxide-modified wheat-straw biochar

Pseudo second-order

Langmuir

24.8 mg/g

Endothermic, spontaneous

References

[63] [62] [64]

Phosphoric acid activated Pinus canariensis

Langmuir

0.16 mg/g

[98]

Urea post-treated Pinus canariensis

Langmuir

0.45 mg/g

[98]

Thermally post-treated Pinus canariensis

Langmuir

0.30 mg/g

Iron nanoparticles synthesized with Eucalyptus leaves

Pseudo second-order

9.69 mg/g

Iron nanoparticles synthesized with green tea extract

Pseudo second-order

13.06 mg/g

Corn stover biochar

Freundlich

384 mg/kg

Ponderosa pine wood residue biochar

Freundlich

158 mg/kg

Switchgrass biochar Graphene

Pseudo second-order

Organically functionalized silica material (MCM-48-NH3 -G)

Pseudo second-order

Synthetic organosilica (MCM-41)

Pseudo second-order

Freundlich

447 mg/kg

Langmuir

848 mg/g

Langmuir

[65]

[66]

Spontaneous, endothermic

[99]

19.9 mg/g

[100]

2.99 mg/g

[67]

Anionic rice husk

Weber–Morris

MgO-modified peanut shell

Freundlich

5.28 mg/g

Langmuir

94 mg/g

Spontaneous, exothermic

[40] [33]

Cationic polymer-modified GAC

Pseudo second-order

Langmuir

27.56 mg/g

Spontaneous, exothermic

Cetylpyridinium bromide-modified zeolite

Pseudo second-order

Freundlich

2.58 mg/g

Spontaneous, exothermic

[102]

Sugar beet bagasse activated carbon

Pseudo second-order

Langmuir

27.55 mg/g

Spontaneous, endothermic

[103]

Modified beet residue

Pseudo second-order

Langmuir

86.21 mg/g

Spontaneous, exothermic

Langmuir

104 mg/g

Sodium bisulfate-conditioned cross-linked chitosan beads

[104] [105]

Chitosan hydrogel beads

Pseudo second-order

Sips

89.66 mg/g

Spontaneous, exothermic

Mg/Al chloride hydrotalcite

Weber–Morris

Langmuir

41.77 mg/g

Spontaneous, endothermic

Langmuir

1.10 mg/g

Wheat straw charcoal

[101]

[81] [106] [69]

Mustard straw charcoal

Langmuir

1.30 mg/g

Sulfuric acid-treated carbon cloth

Langmuir

2.03 mmol/g

Spontaneous

[107]

Ammonium-functionalized mesoporous MCM-48 silica

Langmuir

51.8 mg/g

Spontaneous, exothermic

[108]

Modified wheat residue

Pseudo second-order

Langmuir

2.08 mmol/g

Chemically modified wheat straw

Pseudo first-order

Freundlich

2.69 mg/g

[82]

Langmuir

1.86 mmol/g

[38]

Red mud

[71]

Activated red mud

Langmuir

5.89 mmol/g

Chemically modified Chinese Reed

Langmuir

7.55 mg/g

[109] [110]

Natural Sepiolite

Pseudo second-order

Freundlich

0.386 mg/g

Surfactant-modified sepiolite

Pseudo second-order

Freundlich

3.097 mg/g

Sepiolite activated by HCl

Pseudo second-order

Freundlich

2.49 mg/g

Coconut shell activated carbon

Langmuir

0.26 mmol/g

Spontaneous, endothermic

[111]

Bamboo charcoal

Langmuir

0.10 mmol/g

Spontaneous, endothermic

[45]

416

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

Adsorbent Step 1

Film

Step 2 Adsorbate

Step 4

Step 3

Figure 15.4

Mechanism of adsorption.

Furthermore, physisorption and chemisorption respectively are multilayer and monolayer in nature. The adsorption of PO4 3− and NO3 − by low-cost adsorbents may be because of a mechanism such as surface precipitation, ion exchange, diffusion, ligand exchange, etc. Ion-exchange mechanism is deliberated as physisorption process that is described by weak and reversible adsorption process. For instance, Tshabalala et al. [113] reported that Lewis acid–base interaction and ion exchange are mechanisms accounting for adsorption of PO4 3− by cationized milled wood residues. Similarly, Namasivayam and Holl [109] stated that removal of NO3 − ions by quaternized biomass was due to ion-exchange mechanism. Meanwhile, ligand-exchange mechanism is considered as chemisorption described by fast, stable, and less reversible adsorption. Biswas et al. [59] studied PO4 3− adsorption using metal-loaded orange waste and reported that ligand-exchange mechanism between PO4 3− and OH− coordinated on metal ions infused on adsorbent surface. Similarly, Hassan et al. [68] investigated NO3 − removal by Zr(IV)/sugar beet pulp composite and reported that ligand-exchange mechanism was responsible for adsorption of NO3 − . Furthermore, surface precipitation is considered as chemisorption featured by fast and scarcely reversible adsorption. Shin et al. [94] reported that removal of PO4 3− is because of surface precipitation of PO4 3− ions binding onto Lanthanum-loaded bark fiber. The intraparticle diffusion process is considered as physisorption occurring inside pores of adsorbent. It is characterized by very slow and irreversible adsorption. If √ plot of adsorption capacity versus square root of adsorption time (i.e. qe vs t) is a straight line transiting through origin, √ it indicates that intraparticle diffusion is a significant mechanism. The qe vs t correlation obtained in study by Manjunath

15.5 Management of Spent Adsorbent

and Kumar [23] was nonlinear and passed with some intercepts, indicating that intraparticle diffusion was not a significant mechanism in sorption of PO4 3− and NO3 − onto Prosopis juliflora activated carbon.

15.5 Management of Spent Adsorbent Handling and management of spent adsorbent and recovery of nutrients are one of the essential stages aimed at sustainable treatment of wastewater using adsorption. In recent times, substantial developments have been made in pursuance of enhancing adsorption capacity of low-cost adsorbents for removal of various adsorbates including nutrients from wastewater. On the other hand, very few researchers have considered recycling spent adsorbent and recovering nutrients using various solvents using acids, alkalis, and other organic solvents through desorption. Only a few researchers have paid attention to recovery of nutrients from spent adsorbents and from regenerating solutions. The adsorbent used for removal of nutrients such as NO3 − and PO4 3− can be used as fertilizer for agricultural fields. This improves crop yield as well as provides essential nutrients required by soil from adsorbent. If toxic pollutants are present along with nutrients, adsorbent can be regenerated using various solvents for removal of these toxic pollutants and reused. Once the life of spent adsorbent is over, then it can be disposed of in a scientific manner such as landfilling, incineration, or used as alternate fuel to coal if calorific value of spent adsorbent is high. Figure 15.5 shows different strategies that can be adopted for management of spent adsorbent; however, the strategy depends on economic feasibility, stability, secondary pollution, and carbon sequestration [34]. In order to enhance the reusability of low-cost adsorbent after adsorption of PO4 3− and NO3 − , several researchers have carried out desorption studies using various desorption eluents such as distilled water, HCl, H2 SO4 , NaOH, NaCl, KCl, etc. For instance, Xu et al. [114] studied desorption of PO4 3− from surface of giant reed using 0.1 mol/L HCl, 0.1 mol/L NaOH, 0.1 mol/L NaCl, and distilled water. It was reported that desorption efficiency with HCl, NaOH, and NaCl was 100% meanwhile it was 3.5% with distilled water. Sreenivasulu et al. [97] studied desorption of PO4 3− from surface of stem bark of Eucalyptus tereticornis using distilled water and 0.5 M HCl and reported 50 and 96% desorption of PO4 3− with distilled water and 0.5 M HCl, respectively. It was also found out that adsorbent lasted three adsorption-desorption cycles. Meanwhile, Katal et al. [40] studied desorption of NO3 − from surface of modified rice husk using 0.1 mol/L NaOH and distilled water. It was reported that desorption efficiency of 0.1 mol/L NaOH was 85% while it was 4.2% with distilled water. Tables 15.5 and 15.6 depict desorption capability of different elution solutions for removal of PO4 3− and NO3 − , respectively. Chintala et al. [66] studied desorption of NO3 − using aqueous solution of pH and reported 85, 75, and 90% desorption of NO3 − from surface of corn stover, ponderosa pine wood residue, and switchgrass.

417

418

15 Removal and Recovery of Nutrients Using Low-Cost Adsorbents

Nutrient-loaded adsorbent

Figure 15.5

Pollutant-loaded adsorbent

Handling and management of spent adsorbent.

Table 15.5

Desorption of PO4 3− from various adsorbents.

Adsorbent

Eluent

Desorption (%)

Remark

References

Metal-loaded orange waste

0.4 M HCl

85

PO4 3− can be efficiently eluted with 0.4 M HCl

[59]

Ammonium-functionalized mesoporous silica

0.01 M NaOH

60.6

Five adsorption–desorption cycles

[108]

Giant reed

0.1 mol/L HCl 0.1 mol/L NaOH 0.1 mol/L NaCl Distilled water

100 100 100 3.5

Desorption mechanism is ion exchange

[114]

Ferrihydrite Magnetite

1 N NaOH

75 85 82

Adsorption–desorption cycle was repeated seven times

[83]

Composite metal oxides

0.1 M NaOH

93

Desorption increased with increase of alkalinity

[115]

Calcined waste eggshell

0.5 M NaOH 0.5 M NaCl

37.6 0.7

Addition of solid CaO to NaOH solution precipitated 37.72% of Ca3 (PO4 )2

[54]

Date palm fibers

Distilled water

13.33

Sorption of PO4 3− onto date palm fibers was not completely reversible

[37]

Stem bark of Eucalyptus tereticornis

Distilled water 0.5 M HCl

49.6 95.9

Three adsorption–desorption cycles

[97]

Modified sugarcane bagasse

0.05 M NaOH

95.6

Higher desorption results from OH− ions competing for adsorption sites

[87]

Zr-modified chitosan beads

0.5 M NaOH

92

Five adsorption–desorption cycles

[84]

Date stones (granular)

0.01 M KCl

11.2 13

PO4 3− were firmly attached and bonded to sorbents and adsorption was not completely reversible

[50]

Geothite

Palm fibers

Table 15.6

Desorption of NO3 − from various adsorbents.

Adsorbent

Eluent

Desorption (%)

Remarks

References

Modified corn stalk

HCl (0.1 and 0.5 mol/L) NaCl (0.1 and 0.5 mol/L)

100

Two adsorption–desorption cycles

[40]

Modified rice husk

Distilled water 0.1 mol/L NaOH

4.2 85

Seven adsorption–desorption cycles

[40]

Modified wheat straw

0.1 mol/L NaOH

90

Twelve adsorption–desorption cycles

[82]

Mg-Al-Cl hydrotalcite

Distilled water 1, 2 and 4 M NaCl

1.03 700 ∘ C, whereas wood and pine needle chars had ash contents of 2–4%. It has

441

442

16 Use and Development of Biochar-Based Materials for Effective Capture and Reuse of Phosphorus

been shown that biochar produced at 92% COD& Toxicity

[33]

Domestic wastewater

Real sewage

0.6 kGy dose for outlet wastewater

BOD5 52%, COD 34% for plant A

[34]

OE 2.3 μg/L, 2.7 kGy

BOD5 47%, COD 52% for plant B (continued)

Table 19.1

(Continued)

Pollutants

Matrix

Trichloroethylene and carbon tetrachloride

Simulated

Heavy metals

Aqueous solution

Conditions

Efficiency (%)

Reference

TCE 100 mg/L, 0.1 s residence time, 1.5 kGy

TCE 97% CCl4 95%

[35]

2 mg/L deaerated CdII , 5.10−3 mg/L, 3.5 kGy 5 mg/L deaerated PbII , 10−2 mg/L, 0.7 kGy

CdII < 0.1 mg/L

[36]

CCl4 10 mg/L, 0.1 s residence time, 1 kGy

PbII < 0.1 mg/L

Disperse red 60 (dyes)

Aqueous solution

pH 7, dose 10 kGy combined with coagulation and flocculation

>83% TOC

[37]

Drinking water

Water stream

pH 6–9, petroleum products 0.1 mg/L, 4 kGy

Petroleum products >98.9%

[38]

PCE and TCE

Groundwater

0.3 mg/L PCE, 6 mg/L ozone combined with EB, 80 Gy 0.3 mg/L TCE, 6 mg/L ozone combined with EB, 40 Gy

>90%

[39]

Molasses distillery slops

Simulated

Mixed with municipal wastewater with the ratio of 3 : 4, combined with coagulation and EB, 3.5 kGy

COD > 96.25% BOD20 > 98.5%

[40]

19.3 Textile Dyes in Textile Wastewater and Their Treatment Technologies

study by Strokin [38]; it is noted that the energy consumption of the water treatment using electron accelerator with beam power of 50–60 kW is 5 kWh/m3 and can provide quality water for 100 000 persons. An early research conducted by Rosocha et al. [35] confirmed that TCE and CCl4 can be treated using EB and its effectiveness is based on the irradiation dose and dose method; the research pointed out that using DC irradiation with higher dose, in this case, 1.5 kGy in 0.1 second for TCE with concentration of 100 mg/L and 1 kGy in 0.1 second for CCl4 with concentration of 10 mg/L, resulted in a better removal rate. To consolidate the research of Rosocha et al. [35], Gehringer and Eschweiler [39] applied EB process in combination with ozone to treat groundwater, focusing on two main products, TCE and PCE. It was found that with the addition of 6 mg/L ozone, both PCE and TCE need a significantly lower irradiation dose to decompose 90% of the compounds; PCE went from requiring approximately 1000 Gy of irradiation dose to only 80–110 Gy while TCE went from requiring 320 Gy to 40–60 Gy. A series of studies by Pikaev [36, 37, 40] combining EB with other processes (coagulation and flocculation, biological, advanced oxidation, etc.) has proven that the applications of EB are vast and high efficiency can be achieved for heavy metals (PbII , CdII ), dyes (disperse red 60), surfactants (nekal), molasses distillery slops, etc. Likewise, a study by Deogaonkar et al. [41] emphasized the necessity of combining biological treatment with EB by dosing 1–2 kGy as post-treatment to increase the BOD/COD ratios; this has proven that higher biodegradability is possible and therefore EB is a highly reliable technique in treating nonbiodegradable compounds. In brief, EB radiation can transform many refractory organic compounds into easily biodegradable products, thus improving efficiency and reducing costs for subsequent treatment steps. The mentioned studies indicated that EB could successfully destroy persistent and complex organic compounds even under high salinity or temperature.

19.3 Textile Dyes in Textile Wastewater and Their Treatment Technologies Wastewater from the textile and dyeing industry represents a significant environmental problem due to the presence of synthetic dyes (especially reactive dyes), for several reasons [42–44]. First, government legislation on the standard of water is becoming increasingly stringent. Second, reactive dyes are entirely soluble in water; their complex chemical structures contain low-biodegradable groups, so it is not possible to remove them by the conventional or biological treatment process alone. Third, some of the reactive dyes might cause cancer and/or mutation. Bearing in mind that thousands of tons of reactive dyes are manufactured and used annually by humans, extensive amounts of these compounds can reach the aquatic environment, mainly through wastewater from treatment plants because of their inadequate removal [42, 45]. In order to solve this threat, modern techniques like ozonation, adsorption, membrane, sonolysis, etc. and their combinations are employed to treat textile effluents [43, 46–49]. Most of the techniques do not reduce contaminants but they transfer

533

534

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

pollutants from one stage to another stage. Some methods are selective but have slow to moderate degradation rate or produce more harmful intermediates. Hence, they cannot be employed to treat real textile wastewater. To fully understand this kind of wastewater and the application treatment techniques, this section provides both the dye principle such as dye’s characteristics, dye classification, dyeing process, dyeing wastewater, etc. and current methods for textile treatment.

19.3.1 Textile Dye 19.3.1.1 Dye Classification

There are two practical ways to classify dyes according to chemical structure or the mode of application (usage). The first is of more interest to chemists, who desire to know what molecule makes the color, such as azo, anthraquinone, and phthalocyanine dyes. The second is of relevance to the users, the dye technologists, who need to know which dye is appropriate to the material they want to dye, such as acid dyes for wool and reactive dyes for cotton, and the resultant color. Dyes can be classified according to the nature of their chromophore, an extensive conjugated double bond system containing unsaturated groups. According to the color index [50], there are 25 structural classes based on the chemical type. Among them, the essential dyes are anthraquinone, azo, and phthalocyanine dyes. The dye

Azo dyes Procion orange MX-2R

Phthalocyanine dyes Aizen primula turquose blue GL

Anthraquinone dyes Procion blue MX-R

Figure 19.3

Chemical structure formula of some main dye groups.

19.3 Textile Dyes in Textile Wastewater and Their Treatment Technologies

structure also has electron-withdrawing or electron-donating substituents, called auxochromes, that cause or intensify the color of the chromophores (Figure 19.3). This classification has many benefits. First, it quickly determines dyes as belonging to a group that has specific properties, i.e. anthraquinone dyes (weak and expensive) and azo dyes (robust and cost-effective). Second, there are several chemical groups (approximately 25). Most importantly, the classification is generally accepted by both the dye technologists and the synthetic dye chemists. Therefore, this can be used to quickly identify dyes with terms such as an anthraquinone blue, an azo black, and phthalocyanine green. Application classification: application characteristics and color identify each dye. It is beneficial to consider the classification because of dye nomenclature arising from this system (approximately 28 000 commercial dye names, representing ∼10 500 different dyes). The classification is also of interest to the dye technologist. Using the color index (CI), 15 kinds of dyes are listed in Figure 19.4 [50]. It shows the class, major substrates, characteristics, and the world market share percentage. However, both terminologies are used; for example, an anthraquinone disperses dye for polyester and an azo-reactive dye for cellulose. 19.3.1.2 Nomenclature of Dyes

Dyes are named by their commercial trade name (from the manufacturer) or by their CI, devised by the Society of Dyers and Colorists in the year of 1924 [50]. The trade names of of dyes are usually made up of three parts. The first is a trademark used by the particular manufacturer to designate both the producer and the class of dye. The second is the color, and the third is a series of letters and numbers used as a code by the manufacturer to define more precisely the hue that is normally not standardized (depending on the manufacturer). According to CI, each dye also has a generic name and constitution number that is determined by its application characteristics. However, each manufacturer has a specific name for their dye product. So, a dye may have a lot of trade names but only one CI name. Figure 19.5 describes an example of the names and chemical structure of a dye with azo structure. 19.3.1.3 Dyeing Processes

The dyeing process is fundamentally a process of transferring dye species onto the fiber through an aqueous medium (dye exhaustion) followed by dye penetration into the fibers (dye diffusion). This process is one of the primary factors in the successful trading of textile products. Basically, fabrics are cleaned by water, then dye, and auxiliary chemicals are applied and fixed to these fabrics [3, 4]. The dyeing process involves many steps, which are illustrated in Figure 19.6. It is noticeable that during the dyeing process, the consumption of water and chemical is a significant problem. Most of the dye after the dyeing process is in the wastewater effluent. 19.3.1.4 Reactive Dyeing Mechanism

During the dyeing process, usually, reactive dyes react through nucleophilic substitution or nucleophilic addition reaction with fiber substrate. In the nucleophilic

535

Figure 19.4

Dye classification, substrates, % shared market, and loss to effluent.

Figure 19.5

Name and chemical structure of an example dye.

538

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

Figure 19.6

Flow diagram for textile processing.

substitution reaction, a halogen atom on the reactive dye molecule is replaced by an oxygen cellulose ester covalent bond. In the nucleophilic addition reaction, an alkaline medium could transform the reactive dye to an active species by converting the sulfatoethyl-sulfone to the vinyl sulfone, which reacts with cellulose to form an ether bond. This mechanism of reactive dye and cotton is illustrated in Figure 19.7. At the end of the dyeing process, the dye is found in two forms: linked to the fiber and hydrolyzed (10–50%); the latter form is toxic, shows low biodegradability, and is readily washed off during the dyeing process. Consequently, the residue dye (as

19.3 Textile Dyes in Textile Wastewater and Their Treatment Technologies Nucleophilic substitution reaction-triazine dyes With cellulose (Fixation)

Step 1 Cellulose

Cellulose

H2O Cellulose

Alkalinity

Step 2 Cellulose

Cl

Fixed dye (75%) Hydrolysis reaction

Cl

Hydrolyzed dye (25%)

Figure 19.7

Reaction of triazine and vinyl sulfone dyes with cellulose in alkaline medium.

much as 50% of the initial dye load) consistently appears in the dye bath effluent, which explains why a high load of color and pollutants usually exists in textile wastewater.

19.3.2 Textile Wastewater Characteristic Industrial textile processes generally need vast amounts of clean water, auxiliaries, chemicals, and energy, in which dyeing and finishing stages use approximately 72.3% of water and chemicals. Most of this water and the chemicals are discharged as wastewater, which is described by extreme fluctuations in a lot of parameters such as BOD5 , COD, pH, color, SO4 2− , and temperature as shown in Table 19.2. Another considerable aspect is the degree of fixation on the fiber of reactive dye. Although it highly reacts with fiber, the residue presented in the dye bath effluent usually is 10–50% of the initial dye load. The wastewater composition depends on the different organic-based compounds, chemicals, and dyes used in the industrial dry and wet processing steps [52]. These wastewaters are difficult to degrade by aerobic digestion or steady in oxidizing agents and typically discharged into the aquatic environment without adequate treatment. Additionally, the wastewater from the dyeing processes is at a high temperature and different pH, allowing it to hold a

539

540

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

Table 19.2

Typical characteristics of textile wastewater in Vietnam.

Parameters

Unit

Typical values

Vietnamese technical regulation on the effluent of textile industry (Column A QCVN 13 : 2015/BTNMT)

Temperature

∘C

60–80

40

pH

−

8–13

6–9

BOD5

mg/L

30–5000

30

COD

mg/L

200–11 000

50

Color

Pt−Co

400–5000

20

SO4 2−

mg/L

50–1000

400

TSS

mg/L

0–200

50

Source: Modified from Thanh etal. [51]. © John Wiley & Sons.

high amount of the color elements. Therefore, the pollutant has to be strictly controlled or requires more effective treatment methods before being discharged into the environment [53].

19.3.3 Practical Methods for Treating Textile Wastewater There are lots textile wastewater treatment techniques, which can be classified into three kinds: physicochemical, advanced oxidation, and biological. In general, each method has its drawbacks, and each process alone may not be sufficient to achieve complete decolonization. Some details are provided in Sections 19.3.3.1–19.3.3.3 and Table 19.3. 19.3.3.1 Physicochemical

Physicochemical treatment is a popular method to remove color from dye-containing wastewaters. Physicochemical techniques involve coagulation/flocculation, adsorption, ion-exchange, membrane filtration, electrocoagulation, etc. [5–9]. Among the methods, coagulation is widely employed. Coagulation or chemical precipitation is the most popular and practical method of removing color from reactive textile wastewater [52]. This technique is a process of destabilizing colloids, aggregating them, and joining them together for ease of sedimentation by using coagulants. The widely used coagulants are inorganic coagulants [45] such as ferrous sulfate (FeSO4 ⋅7H2 O), ferric chloride (FeCl3 ), aluminum sulfate (Al2 (SO4 )3 ⋅18H2 O), and polyaluminum chloride (Al2 (OH)n Cl6-n ); organic coagulants [60] such as polyacrylamide, poly(diallyldimethylammonium chloride), and polyacrylic acid coacrylamide; and natural coagulants [61–63] such as chitosan, xanthan gum, and Moringa gum. These coagulants could be categorized into the two parts, as shown in Figure 19.8. The removal mechanisms between coagulant and dye particles are described in Figure 19.9.

Table 19.3 Technology

Physical– chemical

Advanced oxidation process

Summarized evaluation of various treatment technologies of reactive dyeing wastewater. Process

Advantages

Disadvantages

Reference

Adsorption (activated carbon)

The most effective adsorbent, high capacity, proven as one of the most effective treated effluents

High disposal or regeneration of exhausted adsorbent results in a reduction of adsorption capacity

[54]

Ion exchange

The resin contains a high concentration of anion sites, which makes them an effective catalyst for the facile reaction.

High operating cost

[49]

Membrane filtration

Reclamation of textile effluent with good treated water for reuse

Controlling concentrated stream, flux decline, high pressure, expensive, impossible of treating large volumes

[48]

Sodium hypochlorite

Rapid and efficient

Requires the use of chlorine and produces more harmful intermediates

[55]

Coagulation and Simple, economically feasible, excellent flocculation removal of colorants and dissolved organic compounds

Sensitivity to variable water input, high requirement of acid or alkaline for pH adjustments, and high sludge production

[56]

Ozonation

Able to convert resistant dyes into biodegradable species, no alteration of volume

Short half-life (20 min)

[47]

Sonochemical

Simple

Extreme condition needed

[43]

Photocatalysis

No sludge production, can be powered by sunlight, efficient for recalcitrant dye

Useful for a small number of colorants

[57]

Fenton

Fast reaction with strong oxidation agent that could degrade both soluble and insoluble dye

High cost and sludge production

[43, 46]

Aerobic

Destructs azo molecules, environmentally friendly

Need strict operational conditions like pH, temperature, oxygen, and adequate nutrients; produces lots of sludge in which bulking is often observed, not economically feasible on a commercial scale

[58]

Anaerobic

Reduces azo bond, cheaper than other methods, and produces biogas

Needs strict operational condition, longer time for stable condition, the dye cannot be totally mineralized

[58]

Fungi

Degrades many kinds of structures among reactive dyes, even in a complex mixture dye

Long operation time and not economical in reality

[59]

Biological

542

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater Nucleophilic addition reaction-vinylsulfone dyes With cellulose (Fixation)

Step 1 Cellulose

OH

Cellulose

H2O

Alkalinity

OH

OH

Step 2

H2O

Cellulose

se

ulo

ll Ce

OH

H2O

Fixed dye (50%) Hydrolysis reaction

OH

H2O

Hydrolyzed dye (50%)

Figure 19.8

Classification of coagulants according to their origins.

The type of coagulant used influences the degree of destabilization. The higher charge valence the coagulant ion carries, the more destabilization it exerts, and the lesser the dosage required for coagulation. If the pH of solution or wastewater is lower than the isoelectric point of metal hydroxide (during precipitation of colloids by different positively charged coagulants), the metal hydroxide can destabilize negatively charged colloids by charge neutralization. Above the isoelectric point, anionic polymers could dominate where particle destabilization may take place through adsorption and bridge formation. At the overdose of metal coagulants, a sufficient degree of oversaturation appears to generate quick precipitation of a large quantity of metal hydroxide, catching the colloidal particles, which are called sweep floc [64]. For example, when Fe(III) salts are used as coagulants, monomeric and polymeric ferric species are formed; this formation is highly pH-dependent [65]. Some of the reported inorganic and synthetic organic coagulation technologies are summarized below.

19.3 Textile Dyes in Textile Wastewater and Their Treatment Technologies Coagulant

Inorganic

Organic

Aluminum sulfate Ferrous sulfate Synthetic

Ferric chloride Polyaluminum chloride

Natural

Polyacrylamide Poly-diallyl-dimethyl-ammonium chloride Polyacrylic-acid-co-acrylamide Microorganism based Xanthan gum

Animal based Chitosan

Plant based Seed gum extract from – Moringa oleifera – Ipomoea dasysperma – Plantago psyllium – Cyamopsis tetragonoloba Tannin Potato starch

Figure 19.9

Mechanisms of the coagulation process.

Assadi et al. [56] examined the elimination of reactive blue 19 in synthetic wastewater using three coagulants (alum, PAC, and ferric chloride). The impact of coagulant dose, pH, and initial dye concentrations on the coagulation process was investigated in this study. Results showed that the removal capacity of the coagulants increases with a decrease in initial dye concentration. The best dye removal of 91, 92, and 81% was achieved by using 200 mg/L PAC, 300 mg/L alum, and 400 mg/L ferric chloride, respectively, at neutral pH. Golob et al. [45] compared the coagulation capacities of alum, ferrous, ferric sulfate, and commercial cationic coagulant (Colfloc RD) individually and in combination with synthetic reactive black 5 dyeing wastewater. While a poor removal result was obtained by using alum, ferrous, and ferric coagulants alone, the combination of organic coagulant and alum achieved the best results (over 96%) with the addition of an antifoaming agent at neutral pH values. Joo et al. [65] evaluated the removal of reactive black 5, reactive blue 2, reactive red 2, and reactive yellow 2 from aqueous solution and real wastewater by using organic coagulants (synthesized from cyanoguanidine and formaldehyde) and inorganic coagulants (alum or ferric salt) alone or in combination. In this study, the influence of pH, coagulant dose, and the addition of the polymer was determined. For the aqueous solution (synthetic wastewater), the individual use of inorganic coagulant (1 g/L) only reached a maximum color removal of 20%; however, the decolorization efficiency increased to almost 100% on adding an amount of the organic coagulant. Color removal capacity increased as the organic coagulant dosage was incremented, and the efficiency was significantly dependent on the kinds of coagulants used and solution pH. Similarly, the individual use of inorganic coagulants did not significantly mitigate the color from the real wastewater. However, the combination of ferric/organic coagulant and alum salt/ organic coagulant increased decolorization efficiencies up to 40 and 60%, respectively.

543

544

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

Unfortunately, there are some drawbacks to the application of inorganic coagulants and synthesis of organic coagulants in primary treatment as the treatment processes are highly sensitive to wastewater pH and need to dispose of a lot of sludge, which is difficult or nonbiodegradable and may cause secondary pollution. The occurrence of sludge toxicity (e.g. the residual aluminum ions in water relating to Alzheimer’s disease) depends mainly on the coagulant used and wastewater composition [45]. Moreover, the high cost associated with using coagulants for large-scale municipal wastewater treatment also makes it less attractive. 19.3.3.2 Advanced Oxidation Processes (AOPs)

Advanced oxidation can be defined as oxidation by compounds with an oxidation potential (Eo ) higher than that of oxygen (1.23 V) such as hydrogen peroxide (Eo = 1.78 V), ozone (Eo = 2.07 V), and the hydroxyl radical (Eo = 2.28 V). Therefore, AOPs are mostly based on the production of highly reactive radical species (especially the hydroxyl radical HO• ), which can readily degrade recalcitrant organic pollutants and remove certain inorganic pollutants in water and wastewater [66]. AOPs such as ozonation, sonolysis, Fenton’s reagent (Fe2+ /H2 O2 ), and photocatalytic oxidation (UV/TiO2 ) have been widely employed to eradicate dyes compounds. A lot of studies reported that some AOPs seem to be sustainable and successful technologies for destruction of dyes from textile wastewaters [43, 46, 47, 57]. However, most of these techniques are limited by costs and difficulties in industrial operation [43, 46]. Hence, they cannot be applied for large-scale real dyeing wastewater treatment. 19.3.3.3 Biological Methods

Biological treatment of wastewater seems like the most cost-effective method as compared to other treatment techniques [67]. Some biological methods such as fungal or microbial degradation (aerobic or anaerobic) are widely used to mitigate industrial pollutants because a lot of microorganisms, i.e. bacteria, fungi, can degrade various pollutants. However, in practice, compared with other methods, biological treatment is impossible to reduce dyes readily, the degradation rate is usually slow due to the toxicity of commercial dyes. Moreover, the technique requires more treatment time [42, 58, 59].

19.4 Electron Beam Processes for Textile Wastewater Treatment 19.4.1 Lab-Scale Tests Many researchers have investigated the treatment capacity of EB for eliminating real textile wastewater in the lab scale. Some experiments of these EB treatments of real textile and dyeing wastewaters are highlighted in Table 19.4. This section summarizes some of the typical studies [69, 72].

19.4 Electron Beam Processes for Textile Wastewater Treatment

Table 19.4

Some experiments of the EB treating real textile and dyeing wastewaters.

EB systems (employed Investigated absorbed doses) parameters

Changes in parameters

Remarks

Reference

Aerated solutions with coagulations

[68]

EB with coagulation (3 kGy)

BOD5 , COD, TOC, Color index

Irradiation with 3 kGy BOD5 : 1620–700 mg/L

EB (up to 20 kGy) with biodegradation

BOD5 , COD, TOC

2 kGy dose prior to biological treatment reduce time by up to 50% with similar efficiency

Industrial scale with 1000 m3 /day output

[16]

EB (1 kGy)

BOD5 , COD

As using EB, BOD5 /COD ratio was improved from 0.68 to 0.79

EB had converted refractory organic substances to readily biodegradable intermediates

[69]

EB (up to 20 kGy)

BOD5 , COD, TOC

Increase of removal efficiencies up to 30–40%

Irradiation combined [15, 70] with biological treatment

EB (up to 20 kGy)

Color Toxicity

55–96% removal 33–55% reduction

Data for three distinct [71] effluents from the textile industry

TOC: 1000–305 mg/L

Kim et al. [69] conducted an experiment using an EB accelerator (1.0 MeV, 40 kW) on real wastewater collected from Textile Industry Complex at Daegu, Korea. The raw wastewater showed a characteristic pH of 11, temperature of 42 ∘ C, total phosphorus of 2.5 mg/L, total nitrogen of 38 mg/L, BOD5 of 1184 mg/L and COD of 1805 mg/L. It can be seen that the total nitrogen, BOD5, and COD could be categorized as high-strength contaminants. The experiment also investigated the coupling of EB pretreatment and aeration biological treatment for the treatment of textile wastewater. The textile wastewater was treated by an individual aeration tank and EB combined with aeration tank. In the aeration part, dissolved oxygen (DO) values were maintained higher 4 mg/L using a diffuser, while the temperature was kept at room temperature and the mixed liquor suspended solids (MLSS) concentration were controlled in a range of 2000 to –2700 mg/L by return sludge and sludge wasting (1:1 vol./vol.), which is described in Figure 19.10. The results show that with a low dose (1.0 kGy), EB radiation could accelerate the aeration efficiency. For example, BOD5 /COD ratio was incremented after irradiation. The results proved that EB could turn recalcitrant organic compounds into readily biodegradable compounds, therefore, enhancing the efficiency and decreasing the cost of a further biological treatment step. Khomsaton et al. [72] conducted similar experiments by using the EPS 3000 EB machine to treat combined textile (900–3000 mg/L COD, 100–150 mg/L BOD5 and

545

546

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

Figure 19.10

Schematic diagram of biological reactor. Source: Kim et al. [69]. © Elsevier.

above 1000 ADMI unit color) and food processing wastewater (530–1000 mg/L COD, 200–400 mg/L BOD5 and 50–100 ADMI unit color). The current and energy of EBs were 30 mA and 1.0 MeV, respectively, with the dose of 100 kGy irradiation at room temperature. The biological experiments after irradiation tests were conducted using an activated sludge process (DO > 4 mg/L). DO, pH, and MLSS were tested daily for ensuring the system work. The first system was treated for the mixture of non-irradiated textiles and food industry wastewater, and the second for the mix of irradiated textiles and food processing wastewater. Results show that COD removal efficiency and color unit after irradiation reached 62.4% and 379.3 ADMI as compared 29.4% and 899.5 ADMI with single biological treatment. A new membrane bioreactor (MBR) combined with EB (10 MeV, 100 kW; Rhodotron TT 200, IBA, Belgium) was explored by Sun et al. [73] on real textile wastewater containing polyvinyl alcohol (PVA) at Suzhou, China. The influent contained 90–100 mg/L of PVA, 211.7 ± 26.3 mg/L of COD, and pH of 8.1 ± 0.2 with around 0.1‰ nutrient solution. The influence factors of EB system such as current intensity, scan length, and flow velocity of the wastewater were controlled from range of 0 to10 mA, 0 to 100 cm and 0 to 12 m/min, respectively. The coupling

19.4 Electron Beam Processes for Textile Wastewater Treatment

Figure 19.11 Effect of the combination of EB and MBR process on COD removal. Source: Sun et al. [73]. © Elsevier.

process included a batch EB pretreatment and continuous MBR treatment, i.e. the irradiated wastewater from the EB was continuously transferred to the MBR. As a control experiment, the raw, non-radiated wastewater was directly degraded using another MBR. The two MBRs were operated at similar conditions. The maximum flux of the MBRs with hollow fiber membrane (PVDF, 0.02 μm) was 12.5 L/(m2 h). MLSS and MLVSS concentrations are 1.29 g/L and 0.93 g/L, respectively. The sample was obtained from a real aeration tank in a textile wastewater treatment plant (Suzhou, China). The HRT was maintained for 24 hours using a peristaltic pump. The variation of COD in the wastewater using individual biological, individual EB, and EB combined biological treatment are displayed in Figure 19.11. At the stable operation stage (36–86 days), 12 kGy EB process militates approximately 14% of COD as compared to raw wastewater. The hybrid of EB and activated sludge even shown a better result, the COD removal capacities could reach 45% in 86 operating days. The effect of irradiation dose and dye concentration was evaluated by Ting [74] with the textile wastewater from Rawang integrated industrial park (RIIP), Malaysia. The raw dyeing wastewaters were a mixture of reactive dyes (80–90%) and disperse dyes (10–20%). The pH of collected raw dyeing wastewater at RIIP was alkaline (ranged from 10.35 to 10.45), while COD, color, and TSS were 515 mg/L, 1900 color unit (CU), and 250 mg/L, respectively. Before irradiation, the raw wastewater was diluted by distilled water to the desired concentrations of color at 255 CU (abs = 0.072), 520 CU (abs = 0.147), 990 CU (abs = 0.280), and 1900 CU (abs = 0.539). The EB radiation used a 3.0 MeV and 30 mA electron

547

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

accelerator (Nissin, Japan). The accelerator was controlled by an automatic control system that could alter the current intensity (5–20 mA), the flow velocity of the conveyor (0–12 m/min), and the sample thickness is 3.0 mm. Irradiation desired dose was adjusted following EB’s current (mA) and the speed of the conveyor, and the detailed adjustment could be found elsewhere [74]. The absorbed dose was quantified using a film dosimeter CTA (FTR-125, Fuji Photo Film Co.). The experiment was implemented on the Petri dishes in a batch system at the following doses: 0.5, 2.4, 8, 18, 41, 53, 108, and 215 kGy at room temperature. The results indicated that after 108 kGy, the decolorization efficiencies of all concentrations reached over 80%.

19.4.2 Industrial Applications Based on the obtained results for eliminating pollutants from textile wastewater on the lab scale, EB seems like a promising green technology as it not only significantly reduces pollutants for wastewater treatment but also involves no sludge production. The application of EB at the chemical treatment stage or immediately before biological treatment can help significantly decline the consumption of chemical reagents, reduce processing time. Following here is some practical of EB in pilot scale and industry. A 1000 m3 /day hybrid of EB (1.0 MeV, 40 kW) and biological treatment had set up in Taegu, Korea, since October 1998 (Figure 19.12). The wastewater is injected under

250

200

COD (mg/L)

548

150

100

Raw water After IR treatment After IR-MBR treatment

50

0 33

38

43

48

53

58

63

68

73

78

83

88

Days

Figure 19.12 Schematic diagram of a demonstration plant in Korea. Source: Han et al. [15]. © Elsevier.

19.5 Economic Feasibility and Limitations of EB Processes

the EB irradiation area through the injector to reach enough penetration depth. The speed of injection could be altered by the dose and dose rate. After irradiation, the wastewater directly moved into the biological treatment unit and with a low absorbed dose of 1–2 kGy; hydraulic retention time of the biological treatment was decreased by two times at the same removal degree [15, 75]. Han et al. [76] also presented another similar demonstration of EB treatment plant (1.0 MeV, 400 kW accelerator combined with an existing biological treatment facility) in Daegu Dyeing Industry (capacity 10 000 m3 /day) with mixture of raw wastewater from a dyeing process and wastewater from polyester fiber production enriched with Terephthalic acid (TPA) and ethylene glycol (EG). Obtained results indicate that EB treatment could reduce the biological processing time twice at the same level of removal, with an HRT of 17 hours in bio-treatment without preliminary irradiation and of 8 hours in bio-treatment with preliminary EB treatment at the absorbed dose 1–2 kGy. The obtained results show that EB treatment accelerates the process, resulting in a more significant reduction in TOC, CODCr , and BOD5 . Moreover, as compared with an earlier study [15], this EB process not only improves treatment capacity to 30% but also reduces 50% of used chemical reagents.

19.5 Economic Feasibility and Limitations of EB Processes 19.5.1 Economic Feasibility The EB treatment cost relies on a lot of factors, i.e. electron accelerator cost, the dose required for decomposition, wastewater flow rate, the size of the treatment facility, the time utilization of the facility, and capital recovery. Elimination of pollutants in industrial wastewater system usually requires electron accelerators reaching energies more than 1.0 MeV [15, 74]; however, to ensure EC can penetrate wastewater (excluding the hydrodynamic regimes of wastewater flow), some of the electron accelerators (direct current – DC or ultra-high frequency – UHF) used are capable of reaching energies above 4.0 MeV. Some selected accelerators used in wastewater treatment are summarized in Table 19.5. Table 19.5 shows that the cost of the accelerator is dependent on their power. Surprisingly, higher power leads to the lower cost ($/W), therefore making it suitable for wastewater treatment. Han et al. [70] presented that a high-capital cost requirement for 10 000 m3 /d of textile and dyeing wastewater with a high-power electron accelerator (1.0 MeV, 400 kW) is linked with the cost of the accelerator, auxiliary equipment, transport, construction, tax, and installation, which was around $4.0 million. The total operation cost of EB was below $1.0 million/year and approximately US$0.3/m3 of wastewater. With the same electron accelerator (1.0 MeV, 400 kW) and water flow capacity (10 000 m3 /d) of textile and dyeing wastewater from Daegu Dyeing Industrial Complex, similar results on total investment and operation cost were

549

550

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

Table 19.5

Some accelerators for wastewater treatment.

Manufactures (country)

Accelerator

Current (mA)

Budker Institute of Nuclear Physics (Russia)

DC

400

Ion Beam Application (Belgium)

UHF

Radiation Dynamics (USA)

Energy (MeV)

Power (kW)

Price (M$)

Cost ($/W)

1

400

2.0

5.0

15

10

150

6.1

40.7

DC

50

5

250

4.9

19.6

Nissin High Voltage (Japan)

DC

30

5

150

5.0

33.3

Budker Institute of Nuclear Physics (Russia)

UHF

10

5

50

1.2

24.0

also recorded by Han et al. [15]; these authors emphasized the operation cost to be lower than the cost of existing process (around 1.1–1.2$/m3 ). Some construction and operation costs are detailed in Table 19.6. Emami et al. [77] stated that the cost of an EB process is cheaper than activated sludge processes; however, this process is more expensive than chlorination

Table 19.6 Typical cost of the dyeing industrial wastewater plant using EB in Daegu, Korea since 2005; the plant was developed and built by EB-Tech Co. Ltd. Construction cost

Items

Cost (M$)

Accelerator – 1 MeV, 400 kW 3 windows

2.0

Shield room and construction Installation cost-welding/piping/ inspection, transportation, etc.

1.5

Tax, documentation, insurance, others

0.5

Total

4.0

Note

Cost for land, R&D, and approval form authorities are not included

Operating cost Interest

0.24

6%

Depreciation

0.20

20 years

Electricity

0.32

800 kW

Labor

0.10

3 shifts

Maintenance

0.08

2%

Total

0.94

∼ $1.0 million/year

19.5 Economic Feasibility and Limitations of EB Processes

1,000 m3/day Influent 80,000 m3/day

1,000 m3/day

EB irradiation

Figure 19.13

Aeration tanks

Effluent

Reservoir

Treatment cost of different processes [77].

(Figure 19.13). Similar results were also achieved by Maruthi et al. [34], who pointed out that the cost of using EB for disinfecting 200 000 m3 /day water approximated US$0.041/m3 , which is higher than using chlorine (US$0.013/m3 ) but is less expensive as compared to using ozone (US$0.053/m3 ) and UV (US$0.047/m3 ). Therefore, EB can be recommended only if conventional treatments are insufficient or chlorination is not allowed for health reasons.

19.5.2 The Limitations of EB Technology for Wastewater Treatment Even though there are a lot of benefits to using EB in water treatment applications, only some companies have been applying these techniques in reality. The application of this technique in practice has been limited, presumably due to some of the following drawbacks of these techniques [15, 21, 77–79]: ●

● ●

●

High investment and the abundant energy consumed by this process increase doubts about its cost-effectiveness. Presence of suspended solids in wastewater reduces EB treatment efficiency. The EB unit operation is limited by the minimum and maximum flow rates at which a single unit can be operated. The dose cannot be further increased to enhance system performance if treatment goals are not met at the minimum flow rate and the maximum beam current. In such a case, the EB system has to operate in series to obtain a larger EB unit or has to add pretreatment or post-treatment processes, any of which would increase space requirements and costs. If the influent contains volatile organic compounds (VOCs), the EB treatment may result in toxic by-products, which requires additional collection/treatment processes to reduce the VOC by-products to acceptable levels.

EB is only utilized in case the influent wastewater is recalcitrant and these pollutants cannot efficiently be removed by conventional techniques. There are two ways to apply EB economically in real wastewater treatment. In the first case, EB is put prior to other units (usually biological treatment) at high doses because this process could enhance the biodegradation of high load persistent organic pollutants in the wastewater which mentioned in several studies [15, 70, 75]. In the other case, EB is used as a low dose in the last or tertiary stage to eliminate the low concentration

551

552

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

of persistent organic compounds or long-lived pathogens in wastewater before discharging to the environment [34]. While the volume of wastewater in the first case is usually small with high concentrations of pollutants, the wastewater in the last case has a large volume with a low concentration of pollutants.

19.6 Conclusion Textile and dyeing wastewater could contain bio-recalcitrant pollutants, and the inability of conventional wastewater treatment methods to effectively eradicate these bio-recalcitrant contaminants has led to exploring new alternatives. The use of EB technology is almost relatively new in the field of wastewater treatment, but the developments in EB technology have so far helped achieve excellent performance; some EB systems show the possibility of complete mineralization of refractory matter including dyes, antibiotics, pesticides, etc. The hybrid EB-coupled ozone or biological treatment unit has been demonstrated as an eco-friendly and economical alternative technology to mitigate individual textile and dyeing wastewater or a mixture of this wastewater sources with other pollutants in lab-scale and full-scale operation. Based on the obtained results, the EB unit could be economically utilized in two positions in the wastewater treatment plant: firstly, before the biological treatment unit to destruct or improve the biodegradable capacity of the persistent organic compounds at high dose rate; and secondly, as the tertiary process instead of chlorination at low dose rate. Even, some findings indicated that EB could be a promising and green alternative for water and wastewater treatment (with advantage such as no additional chemicals, no sludge production, ease of control, etc.). This technique has some drawbacks especially in terms of investment cost and further investigation is needed to assess the removal performance of pollutants in various textile and dyeing wastewaters and/or other similar wastewaters containing nonbiodegradable matter.

References 1 Bisschops, I. and Spanjers, H. (2003). Literature review on textile wastewater characterisation. Environmental technology. 24: 1399–1411. 2 Kant, R. (2012). Textile dyeing industry an environmental hazard. Natural science. 4 (1): 22–26. 3 Pirkarami, A. and Olya, M.E. (2017). Removal of dye from industrial wastewater with an emphasis on improving economic efficiency and degradation mechanism. Journal of Saudi Chemical Society. 21: S179–S186. 4 Mani, S., Chowdhary, P., and Bharagava, R.N. (2019). Textile wastewater dyes: toxicity profile and treatment approaches. In: Emerging and Eco−Friendly Approaches for Waste Management, 1e (eds. R.N. Bharagava and P. Chowdhary), 219–244. Singapore: Springer Nature.

References

5 Chávez, A.M., Gimeno, O., Rey, A. et al. (2019). Treatment of highly polluted industrial wastewater by means of sequential aerobic biological oxidation−ozone based AOPs. Chemical Engineering Journal. 361: 89–98. 6 Ha, B.M., Yuan, P.S., and Huong, D.T.G. (2016). The use of artificial neural network for modelling coagulation of reactive dyeing wastewater using Cassia Fistula gum. Journal of Environmental Science and Management. 19 (2): 1–10. 7 Ha, B.M. (2016). Modeling the removal of Sunfix Red S3B from aqueous solution by electrocoagulation process using artificial neural network. Journal of the Serbian Chemical Society. 81 (8): 959–970. 8 Ha, B.M., Huong, D.T.G., and Cuong, N.D. (2016). Applying an artificial neural network to predict coagulation capacity of reactive dyeing wastewater by chitosan. Polish Journal of Environmental Studies. 25 (2): 545–555. 9 Salleh, M.A.M., Mahmoud, D.K., Karim, W.A.W.A. et al. (2011). Cationic and anionic dye adsorption by agricultural solid wastes: a comprehensive review. Desalination. 280: 1–13. 10 Yuan, P.S. and Ha, B.M. (2015). The feasibility of cassia fistula gum with polyaluminium chloride for decolorization of reactive dyeing wastewater. Journal of the Serbian Chemical Society. 80 (1): 115–125. 11 Ha, B.M. (2018). Decolorization of Suncion Red P−2B solution by gamma Co−60 irradiation in the presence of hydrogen peroxide. Journal of Advanced Oxidation Technologies. 21 (1): 118–126. 12 Chen, C.M., Shih, M.L., Lee, S.Z. et al. (2001). Increased toxicity of textile effluents by a chlorination process using sodium hypochlorite. Water Science and Technology. 43 (2): 1–8. 13 Muñoz, I., Malato, S., Rodriguez, A. et al. (2008). Integration of environmental and economic performance of processes. Case study on advanced oxidation processes for wastewater treatment. Journal of Advanced Oxidation Technologies. 11 (2): 270–275. ´ 14 Bilinska, L., Gmurek, M., and Ledakowicz, S. (2016). Comparison between industrial and simulated textile wastewater treatment by AOPs – biodegradability, toxicity and cost assessment. Chemical Engineering Journal. 306: 550–559. 15 Han, B., Kim, J.K., Kim, Y. et al. (2012). Operation of industrial−scale electron beam wastewater treatment plant. Radiation Physics and Chemistry. 81 (9): 1475–1478. 16 Han, B., Kim, J., Kim, Y. et al. (2012). Industrial application of e-beam accelerators in Korea. Journal of the Korean Physical Society. 61 (2): 180–184. 17 Hayashi, T., Takahashi, Y., and Todoriki, S. (1998). Sterilization of foods with low−energy electrons (“soft−electrons”). Radiation Physics and Chemistry. 52 (1–6): 73–−76. 18 Borrely, S.I., Del Mastro, N.L., and Sampa, M.H.O. (1998). Improvement of municipal wastewaters by electron beam accelerator in Brazil. Radiation Physics and Chemistry. 52 (1): 333–337. 19 Wojnárovits, L., Takács, E., and Szabó, L. (2018). Gamma−ray and electron beam based AOPs. In: Advanced Oxidation Processes for Wastewater Treatment, 1e (ed. M. Stefan), 241–296. London: IWA publishing.

553

554

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

20 Aal, S.E.A., Dessouki, A.M., and Sokker, H.H. (2001). Degradation of some pesticides in aqueous solutions by electron beam and gamma−radiation. Journal of Radioanalytical and Nuclear Chemistry. 250 (2): 329–334. 21 Kurucz, C.N., Waite, T.D., and Cooper, W.J. (1995). The Miami electron beam research facility: a large scale wastewater treatment application. Radiation Physics and Chemistry. 45 (2): 299–308. 22 Kim, T.H., Nam, Y.K., and Lee, M.J. (2009). Electron−beam irradiation of livestock wastewater in the presence of natural zeolite. Journal of the Korean Physical Society. 54 (5): 2109–2114. 23 Cho, J.Y. (2010). Evaluation of degradation of antibiotic tetracycline in pig manure by electron beam irradiation. Bulletin of Environmental Contamination and Toxicology. 84 (4): 450–453. 24 McColl, N., Auvinen, A., Kesminiene, A. et al. (2015). European code against cancer 4th edition: ionising and non−ionising radiation and cancer. Cancer Epidemiology. 39: S93–S100. 25 Buxton, G.V. (2008). An overview of the radiation chemistry of liquids. In: Radiation Chemistry: From Basics to Applications in Material and Life Sciences, 1e (eds. S.M. Melanie, M. Mehran, D. Thierry and B. Jacqueline), 3–16. Orsay: Édition Diffusion Presse Sciences. 26 Gehringer, P. and Fiedler, H. (1998). Design of a combined ozone/electron beam process for waste water and economic feasibility of the process. Radiation Physics and Chemistry. 52 (1–6): 345–349. 27 Trojanowicz, M., Bobrowski, K., Szreder, T. et al. (2018). Gamma−ray, X−ray and electron beam based processes. In: Advanced Oxidation Processes for Waste Water Treatment: Emerging Green Chemical Technology, 1e (eds. S.C. Ameta and R. Ameta), 257–331. Chenai, India: Academic Press. 28 Jung, I.H., Lee, M.J., and Mah, Y.J. (2012). Decomposition of PCBs in transformer oil using an electron beam accelerator. Radiation Physics and Chemistry. 81 (7): 899–905. 29 Duarte, C.L., Sampa, M.H.O., Rela, P.R. et al. (2002). Advanced oxidation process by electron−beam−irradiation−induced decomposition of pollutants in industrial effluents. Radiation Physics and Chemistry. 63 (3–6): 647–651. 30 Petrovic, M., Gehringer, P., Eschweiler, H. et al. (2007). Radiolytic decomposition of multi−class surfactants and their biotransformation products in sewage treatment plant effluents. Chemosphere. 66 (1): 114–122. 31 Shin, H.S., Kim, Y.R., Han, B. et al. (2002). Application of electron beam to treatment of wastewater from papermill. Radiation Physics and Chemistry. 65 (4): 539–547. 32 Bojanowska-Czajka, A., Nichipor, H., Drzewicz, P. et al. (2011). Radiolytic decomposition of pesticide carbendazim in waters and wastes for environmental protection. Journal of Radioanalytical and Nuclear Chemistry. 289 (2): 303–314. 33 Thill, P.G., Ager, D.K., Vojnovic, B. et al. (2016). Hybrid biological, electron beam and zero−valent nano iron treatment of recalcitrant metalworking fluids. Water Research. 93: 214–221.

References

34 Maruthi, Y.A., Das, N.L., Hossain, K. et al. (2011). Disinfection and reduction of organic load of sewage water by electron beam radiation. Applied Water Science. 1 (1): 49–56. 35 Rosocha, L.A., Secker, D.A., and Smith, J.D. (1994). Kinetic modeling of trichloroethylene and carbon tetrachloride removal from water by electron−beam irradiation. In: Emerging Technologies in Hazardous Waste Management, 4e (eds. T.D. William and P.G. Frederick), 184–196. Washington, DC: American Chemical Society. 36 Pikaev, A.K., Kartasheva, L.I., Zhestkova, T.P. et al. (1997). Removal of heavy metals from water by electron−beam treatment in the presence of an hydroxyl radical scavenger. Mendeleev Communications. 7 (2): 52–53. 37 Pikaev, A.K., Makarov, I.E., Ponomarev, A.V. et al. (1997). A combined electron−beam and coagulation method of purification of water from dyes. Mendeleev Communications. 7 (5): 176–178. 38 Strokin, N.A. (2007). Electron−beam treatment of drinking water in an industrial system. High Energy Chemistry. 41 (1): 1–4. 39 Gehringer, P. and Eschweiler, H. (1999). Ozone/electron beam process for water treatment: design, limitations and economic considerations. Ozone: science & engineering. 21 (5): 523–538. 40 Pikaev, A.K., Ponomarev, A.V., Bludenko, A.V. et al. (2001). Combined electron−beam and coagulation purification of molasses distillery slops. Features of the method, technical and economic evaluation of large−scale facility. Radiation Physics and Chemistry. 61 (1): 81–87. 41 Deogaonkar, S.C., Wakode, P., and Rawat, K.P. (2019). Electron beam irradiation post treatment for degradation of non biodegradable contaminants in textile wastewater. Radiation Physics and Chemistry. 165: 108377. 42 Gottlieb, A., Shaw, C., Smith, A. et al. (2003). The toxicity of textile reactive azo dyes after hydrolysis and decolourisation. Journal of Biotechnology. 110 (1): 49–56. 43 Ince, N.H. and Tezcanl𝚤´ , G. (2001). Reactive dyestuff degradation by combined sonolysis and ozonation. Dyes and Pigments. 49 (3): 145–153. 44 Ha, B.M. (2014). Treatment of wastewater containing reactive dyes by coagulation and electrocoagulation processes. Doctoral thesis. Dayeh University. 45 Golob, V., Vinder, A., and Simonic, M. (2005). Efficiency of the coagulation/flocculation method for the treatment of dyebath effluents. Dyes and Pigments. 67 (2): 93–97. 46 Chung, J. and Kim, J.O. (2011). Application of advanced oxidation processes to remove refractory compounds from dye wastewater. Desalination and Water Treatment. 25 (1–3): 233–−240. 47 Zou, L. and Zhu, B. (2008). The synergistic effect of ozonation and photocatalysis on color removal from reused water. Journal of Photochemistry and Photobiology A: Chemistry. 196 (1): 24–32. 48 Zhu, X., Zheng, Y., Chen, Z. et al. (2013). Removal of reactive dye from textile effluent through submerged filtration using hollow fiber composite nanofiltlration membrane. Desalination and Water Treatment. 51 (31–33): 6101–6109.

555

556

19 Trends in Using Electron Beam for Treating Textile and Dyeing Wastewater

49 Karcher, S., Kornmüller, A., and Jekel, M. (2002). Anion exchange resins for removal of reactive dyes from textile wastewaters. Water Research. 36 (19): 4717–4724. 50 Krause, J. (2002). Color Index: Over 1,000 Color Combinations, CMYK and RGB Formulas. Oxford: F & W Publications. 51 Thanh, B.X., Tin, N.T., and Dan, N.P. (2013). Influence of recirculation rate on performance of membrane bioreactor coupling with ozonation treating dyeing and textile wastewater. Journal of Water Sustainability. 3 (2): 71–78. 52 Verma, A.K., Dash, R.R., and Bhunia, P. (2012). A review on chemical coagulation/flocculation technologies for removal of colour from textile wastewaters. Journal of Environmental Management. 93 (1): 154–168. 53 Hossain, L., Sarker, S.K., and Khan, M.S. (2018). Evaluation of present and future wastewater impacts of textile dyeing industries in Bangladesh. Environmental Development. 26: 23–33. 54 Yang, X. and Al-Duri, B. (2005). Kinetic modeling of liquid−phase adsorption of reactive dyes on activated carbon. Journal of Colloid and Interface Science 287 (1): 25–34. 55 Ho, C.H., Chen, L., Ho, Y.P. et al. (2010). Oxidative decomposition of reactive blue CI 19 with sodium hypochlorite. Environmental Engineering Science. 27 (1): 103–109. 56 Assadi, A., Amin, M., and Nateghi, R. (2013). Application of coagulation process reactive blue 19 dye removal from textile industry wastewater. International Journal of Environmental Health Engineering. 2 (1): 5–5. 57 Ganesan, R. and Thanasekaran, K. (2011). Decolourisation of textile dyeing wastewater by modified solar photo−Fenton oxidation. International Journal of Environmental Science. 1: 1168–1176. 58 Naimabadi, A., Attar, H.M., and Shahsavani, A. (2009). Decolorization and biological degradation of azo dye reactive Red 2 by anaerobic/aerobic sequential process. Iranian Journal of Environmental Health Science & Engineering. 6 (2): 67–72. 59 Machado, K.M.G., Compart, L.C.A., Morais, R.O. et al. (2006). Biodegradation of reactive textile dyes by basidiomycetous fungi from brazilian ecosystems. Brazilian Journal of Microbiology. 37: 481–487. 60 Sanghi, R., Bhattacharya, B., and Singh, V. (2006). Use of Cassia javahikai seed gum and gum−g−polyacrylamide as coagulant aid for the decolorization of textile dye solutions. Bioresource Technology. 97 (10): 1259–1264. 61 Wen, Y.Z., Liu, W.Q., Fang, Z.H. et al. (2005). Effects of adsorption interferents on removal of Reactive Red 195 dye in wastewater by chitosan. Journal of Environmental Sciences. 17 (5): 766–769. 62 Juan, A.L.Á., Juan, J.R., Guillermo, M.D. et al. (2009). Study of sorption equilibrium of biopolymers alginic acid and xanthan with C.I. disperse yellow 54. Journal of the Mexican Chemical Society. 53: 59–−70. 63 Ndabigengesere, A. and Subba, N.K. (1998). Quality of water treated by coagulation using Moringa oleifera seeds. Water Research. 32 (3): 781–791.

References

64 Ghernaout, D. and Ghernaout, B. (2012). Sweep flocculation as a second form of charge neutralisation − a review. Desalination and Water Treatment. 44 (1–3): 15–28. 65 Joo, D.J., Shin, W.S., Choi, J.H. et al. (2007). Decolorization of reactive dyes using inorganic coagulants and synthetic polymer. Dyes and Pigments. 73 (1): 59–64. 66 Glaze, W.H. (1987). Drinking−water treatment with ozone. Environmental Science & Technology. 21 (3): 224–230. 67 Josmaria, L.D.M. and Patricio, P.Z. (2005). Use of advanced oxidation processes to improve the biodegradability of mature landfill leachates. Journal of Hazardous Materials. 123 (1–3): 181–186. 68 Ponomarev, A.V., Makarov, I.E., Bludenko, A.V. et al. (1999). Combined electron−beam and coagulation treatment of industrial wastewater with high concentrations of organic substances. High Energy Chemistry. 33 (3): 145–−149. 69 Kim, T.H., Lee, J.K., and Lee, M.J. (2007). Biodegradability enhancement of textile wastewater by electron beam irradiation. Radiation Physics and Chemistry. 76 (6): 1037–1041. 70 Han, S.K., Myun, S.K., Gu, W.K. et al. (2011). High−power accelerator for environmental applications. Journal of Korean Physical Society. 59: 3485. 71 Borrely, S.I., Morais, A.V., Rosa, J.M. et al. (2016). Decoloration and detoxification of effluents by ionizing radiation. Radiation Physics and Chemistry. 124: 198–202. 72 Khomsaton, A.B., Selambakkanu, S., and Jamaliah, S. (2011). E-beam irradiation and activated sludge system for treatment of textiles and food base industrial waste water. Jurnal Sains Nuklear Malaysia. 27 (1): 1–6. 73 Sun, W., Chen, J., Chen, L. et al. (2016). Coupled electron beam radiation and MBR treatment of textile wastewater containing polyvinyl alcohol. Chemosphere. 155: 57–61. 74 Ting, T.M. (2008). Decolorization and decomposition of organic pollutants for reactive and disperse dyes using electron beam technology: effect of the concentrations of pollutants and irradiation dose. Chemosphere. 73 (1): 76–80. 75 Han, B., Ko, J., Kim, J. et al. (2002). Combined electron−beam and biological treatment of dyeing complex wastewater. Pilot plant experiments. Radiation Physics and Chemistry. 64 (1): 53–59. 76 Han, B., Kim, J., Kim, Y. et al. (2005). Electron beam treatment of textile dyeing wastewater: operation of pilot plant and industrial plant construction. Water Science and Technology. 52 (10–11): 317–324. 77 Emami, M.M., Parsaeian, M.R., Amraei, R. et al. (2016). An experimental investigation of wastewater treatment using electron beam irradiation. Radiation Physics and Chemistry. 125: 82–87. 78 Hossain, K., Maruthi, Y.A., Das, N.L. et al. (2018). Irradiation of wastewater with electron beam is a key to sustainable smart/green cities: a review. Applied Water Science. 8 (1): 6. 79 Kurilova, A.A., Poloskov, A.V., Chubik, M.V. et al. (2015). Application of electron beam for wastewater disinfection. Procedia Chemistry. 15: 187–192.

557

559

20 Approaches Toward Resource Recovery from Breeding Wastewater Huu Hao Ngo 1 , Dongle Cheng 2 , and Wenshan Guo 2 1 Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, Australia 2 Joint Research Centre for Protective Infrastructure Technology and Environmental Green Bioprocess, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW 2007, Australia and Department of Environmental and Municipal Engineering, Tianjin Chengjian University, Tianjin 300384, China

20.1 Introduction Due to important changes of human dietary patterns, the food demands of animal origin, including meat, milk, and dairy products, aquatic products, and poultry eggs and meats, have experienced high rates of growth. Since 1961, global per capita meat, food fish, and milk consumption has increased approximately 20, 11, and 448 kg, respectively, which double the level of consumption in the 1960s [1]. Driven by the rapid growth of global population, the production of cattle, sheep, goats, pigs, and chickens increased 258, 90.5, 637.2, 456.6, and 189%, respectively, during the period of 1890 and 2014 [1]. The development of aquaculture as a source of affordable animal protein is also essential to human existence, which increased 46.1 times from the year 1960 to 2014, mainly driven by the increase of demands and reduction of wild fish catch [2]. To increase animal densities and production efficiency, small-scale family farms have changed to more specialized and much larger concentrated animal feeding operations (CAFOs) since the late 1970s [3]. However, high density of animals in CAFOs, results in the generation of large amounts of wastes with high concentrations of nitrogen, phosphorus, organic matter, and fecal microbes [4]. As reported, global manure production from all livestock increased 66% during 1961–2016 [5]. Wastes from CAFOs are usually applied to agriculture for irrigation and as a fertilizer in crop production. However, large amounts of wastes produced from CAFOs are beyond the crop nutrient needs and absorption capacity of the nearby land. The application of excessive amounts of livestock and poultry wastes to farmland can lead to the accumulation of nutrients in soils with potential surface water and groundwater pollution, such as surface water eutrophication and groundwater nitrate enrichment [6, 7]. Additionally, the rapid expansion of aquaculture has also contributed to the excessive increase of nutrients in aquatic ecosystems Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

560

20 Approaches Toward Resource Recovery from Breeding Wastewater

[8]. Therefore, effective management of breeding wastes is significant before the discharge of their effluents to the environment. The high carbon, nitrogen, and phosphorus contents in breeding wastewater make it highly suitable for targeted resource recovery [9]. Animal wastes were regarded as the greatest potential feedstock for energy recovery (547.1 trillion British thermal units) in the United States in 2017 [10]. Compared with traditional wastewater treatment strategies, resource recovery from wastewater has the benefit of (i) reducing the wastewater treatment costs by the production of valuable energy and products, (ii) reducing production costs of these products by recovering the raw materials from wastewater, and (iii) eliminating environmental pollutants [11]. The increasing demands of renewable energy and resources are becoming strong drivers for the development of technologies for resource recovery [12]. Technologies for resource recovery from breeding wastewater mainly include biological, physicochemical, and plant-based processes. The biological conversion of breeding wastes can be done via anaerobic digestion and dark fermentation, where methane or hydrogen is produced from organic matter by microorganisms under anaerobic conditions. Approaches such as ammonium stripping, ion-exchange, chemical precipitation, and membrane filtration represent the main physicochemical technologies for recovering nutrients and water resources from breeding wastewater. Nutrients in breeding wastewater also can be recovered by cultivating aquatic plants in the wastewater through the production of biomass. Moreover, the harvested biomass can be further converted to valuable bioenergy and products. All of these processes are affected by multiple factors and require some specific conditions such as wastewater compositions, temperature, pH, etc. [13]. This chapter aims to outline (i) the characteristics of breeding wastewater; (ii) major resources in breeding wastewater; (iii) approaches for resource recovery; and (iv) the current application of resource recovery from breeding wastewater.

20.2 Characteristics of Breeding Wastewater 20.2.1 Livestock Wastewater In livestock farms, increasing quantities of livestock wastewater is produced with the growth of meat consumption worldwide, which poses a serious threat to the environment and a challenge for the appropriate treatment [14]. Specifically, the global consumption of pork and beef was up to 120.71 and 71.72 million tons in 2018, respectively. China and European Union have been reported as the largest producers of swine, while India and Brazil lead in cattle production [15]. According to the report of [16], the swine production in China and European Union was around 4.3325 × 108 heads and 1.5 × 108 heads in 2018, respectively. The cattle production in India and Brazil in 2018 was up to 3.05 × 108 heads and 2.3235 × 108 heads, respectively. Take Brazil as an example, the volume of produced swine wastewater is of approximately 7–15 L/swine per day [17]. Theoretically, the production of livestock wastewater is positive with the livestock production and meat consumption. Thus,

20.2 Characteristics of Breeding Wastewater

Table 20.1

Characteristics of livestock, poultry and aquaculture wastewater.

Characteristics

Swine wastewatera)

Cattle wastewatera)

Poultry wastewaterb)

Aquaculture wastewatera)

pH

7.0–8.5

6.0–7.9

8.23–8.8

6.5–8.2

SS (mg/L)

45–10 770

938

–

0.42–50

COD (mg/L)

2 388–65 200

890–41 000

4 500–7 690

80–1 201

TN (mg/L)

378–5 034

–

1 570–2 473

34–133

NH4–N (mg/L)

80–4 800

70–2 400

279–1 787

4–101

TP (mg/L)

30–2 800

10–1 200

200–248

0.6–25

Ca (mg/L)

39–255

348

35.5–308

–

Mg (mg/L)

6.7–131

156

13.7–146

8.45

K (mg/L)

301–3 800

400–5 200

691–2 585

5.9

a) Ref. [18]. b) Refs. [16, 19]. c) Refs. [20–22].

the larger the livestock production and meat consumption, the more wastewater would be generated. Livestock wastewater mainly consists of feces, urine, and water from discharge facilities. Factors including the environment, livestock species, and their diet, their ability to digest certain foods, age of the animal, and their stage of life can affect the properties of livestock wastewater [13]. In general, wastewater generated from low food digestibility (98

(NH4 )2 SO4

High water and ammonia recovery

Membrane fouling, pH, and heating source

[87]

52–78

Concentrated N Nutrient concentration Membrane fouling and P

pH > 9.7; ammonia stripping solution: H2 SO4 ED

Electrical pressure: 98 < 60 mA/cm2

Disadvantages

References

[100]

[189]

20.4 Approaches for Resource Recovery

10.5–11. At high temperature (80 ∘ C), the ammonia removal efficiency of 65, 69, and 98.8% was recorded from fresh pig slurry when the stripping was performed at non-modified pH, initial pH = 9.5, and initial pH = 11.5, respectively [89]. Liao et al. [90] also indicated that the best ammonia removal (90%) from swine wastewater achieved with seven hours of treatment by air-stripping method was at a high pH of 11.5. By increasing the initial pH from 7.5 to 11.5, the time needed for complete ammonia removal reduced from five to three hours [89]. Therefore, higher pH promotes energy saving and reduces time for treatment. Moreover, the pretreatment of raw breeding wastewater through air stripping at an appropriate pH range has been considered as a viable option for reducing the inhibition of ammonia to the anaerobic digestion processes. Zhang et al. [91] demonstrated that the methane productivity increased more than three times compared to the control when ammonia in swine wastewater was air-stripped at pH 9.5. Nevertheless, high doses of lime and sodium hydroxide are usually added to the wastewater for the adjustment of initial pH, leading to an increase of the operational costs [92]. Moreover, the introduction of chemicals also increases the sludge precipitation, causes serious scaling problems, as well as inhibits microorganisms. Moreover, microorganisms in anaerobic digestion processes could be inhibited by the toxicity of residual cations [91]. To address the above challenges of ammonia stripping from wastewater, several innovative technologies have been developed [92]. For example, an electrolyzed water system with an electro-diaphragm between the electrodes was reported to produce acidic water (pH 2–3) at the anode and alkaline water (pH 11–12) at the cathode, leading to increasing pH for ammonia stripping [93]. Ukwuani and Tao [94] developed an innovative vacuum thermal stripping coupled with acid absorption process to reduce the cost of heating. Due to vacuum, ammonia is stripped out of wastewater at a temperature lower than the normal boiling point and the high-purity ammonium sulfate was obtained from acid solution [94]. Further researches are required to limit the drawback of the ammonia stripping approach and realize the full-scale application on recovering NH4 + from breeding wastewater. 20.4.2.2 Ion-Exchange and Adsorption

Another approach for recovering nutrients from breeding wastewater is ion-exchange and adsorption, which are not highly influenced by temperature variations with relatively simple operation [95]. Ion-exchange and adsorption utilize sorbents to extract the target compounds (e.g. PO4 3− or NH4 + ) from the feed solution through intermolecular forces and ionic forces, respectively. These two processes may occur simultaneously in one media [87]. Normally, the ion-exchange consists of contact between a solid ion exchanger and the medium in an aqueous solution. An example reaction between the exchanger and NH4 + is as follows: M− A+ + NH4 + ↔ M− NH4 + + A+ [96]. There are different types of sorbents that have been used for nutrient removal from wastewater, such as activated carbon, biomaterials, clay minerals, natural and modified zeolites, and some industrial solid wastes. Zeolites are the first and most frequently studied materials in last decades. Natural zeolites contain massive

573

574

20 Approaches Toward Resource Recovery from Breeding Wastewater

cavities free NH3 and exchangeable cations for NH4 + and molecular adsorption is the main mechanism [97]. After land application, zeolites can potentially release the adsorbed NH4 + to soil, reducing overall fertilizer requirements and decreasing eutrophication by increasing the soil ion-exchange capacity [98]. Likewise, phosphorus can be adsorbed by zeolite, especially with the coexistence of NH4 + [99]. Therefore, ion-exchange using zeolites can be a potentially cost-effective method for recovering nutrients from breeding wastewater [100]. In general, zeolites should be pretreated to improve their purity and NH4 + exchange capacity for removing NH4 + from breeding wastewater or the digested effluent, which contain NH4 + greater than 1000 mg/L [101]. For example, zeolites pretreated with NaCl solution showed an excellent ion-exchange capacity [95]. It has been observed that the adsorption capacity for NH4 + by Na–zeolite (21 g/kg) was greater than by natural zeolite (19 g/kg) [102]. MgO-modified zeolites also enhanced ammonium adsorption capacity from 12.6 to 24.9 g/kg through ion-exchange, molecular adsorption, and the formation of complexes between NH4 + and functional groups of zeolite itself [101]. Comparing with natural zeolite, the use of MgCl2 -modified zeolite at pH 8–9.5 also showed much better efficiencies of NH4 + and P removal from swine wastewater [103]. With favorable characteristics of microorganism adhesion, zeolite has also been widely used as an ion exchanger for preventing ammonium inhibition in anaerobic digestion of breeding wastewater [104]. A fixed zeolite bioreactor was employed to reduce the inhibition of high NH4 + concentration (3770 mg/L) during anaerobic digestion of swine wastes. The bioreactor significantly shortened the startup time, enhanced methane gas yield more than twofold, and removed COD effectively [105]. Parameters, such as particle size, solution pH, adsorbent dosage, contact time, temperature, and initial ammonium concentration affect the NH4 + exchange capacity of the zeolites significantly [106]. In addition, the presence of competitive ions such as Na+ , K+ , Ca2+ , and Mg2+ in breeding wastewater has to be considered, which decreases the NH4 + removal efficiency by modified zeolite in the order of preference of Ca2+ > Mg2+ > K+ > Na+ at identical molar concentrations [103]. 20.4.2.3 Chemical precipitation

Chemical precipitation can be applied to recover both N and P from breeding wastewater simultaneously by forming struvite [Mg(NH4 )PO4 ⋅6H2 O]. It is notable that struvite constituent ions, Mg2+ , NH4 + , and orthophosphate (PO4 –P), are among the predominant ions present in raw and digested breeding wastewater, thereby minimizing the need to add chemicals [107]. The formed struvite has been regarded as an environmentally friendly fertilizer compared with the traditional chemical, due to the slow-release rate of nutrients and low heavy metal content compared with P rocks, which further reduces the demands for conventional fertilizer production [107]. In comparison with the conventional phosphorus fertilizer made of phosphate rocks, struvite recovered from wastewater also has lower concentrations of heavy metals and radioactive elements [107]. The recovery of N (7.5–97.3%) and P (30–99.5%) from breeding wastewaters via struvite formation has been reported by previous reports [18]. Moreover, the struvite formation was

20.4 Approaches for Resource Recovery

also reported to reduce pathogenic populations and heavy metals effectively [13]. Muhmood et al. [16] observed that along with the recovery of nutrients, 40, 45, 66, 30, 20, and 70% of Zn, Cu, Pb, Cr, Ni, and total coliforms and Escherichia coli were removed by struvite precipitation from poultry slurry, respectively. Struvite formation can be observed in a wider pH range from 7 to 11, but the optimal pH is in the range of 8.5–10.7. pH value and the molar ratio of Mg : NH4 : P of wastewater are the two main impact factors for struvite formation [108]. For example, Stratful et al. [109] indicated that the phosphate removal ratio from swine wastewater increased from 5 to 90% with the pH increasing from 7 to 9. Similar removal efficiency of soluble phosphorus (91%) was also achieved by Burns et al. [110], as adjusting the pH of the treated swine wastes to pH 9.0 with sodium hydroxide (NaOH). To keep pH in this optimal range, continuous alkali addition and/or aeration has to be conducted. For instance, the pH of swine wastewater increased up to approximately 8.5 with continuous aeration, and a large part of the soluble PO4 –P, Mg, and Ca were crystallized [111]. In breeding wastewater, the NH4 + concentration is abundant for struvite formation in comparison with PO4 3− , whereas Mg2+ is insufficient. Thus, although P can be effectively recovered from breeding wastewater via struvite crystallization, the recovery of NH4 + is quite low. As indicated by Song et al. [112], struvite crystallization achieved 90% removal of total P, but the removal of total ammonia nitrogen (TAN) was only 13% from swine wastewater at a pH range of 9.5–11. To enhance the recovery of NH4 + and PO4 3− simultaneously by struvite precipitation, high amounts of external P and Mg source should be supplemented into the system. A field experiment conducted in a swine manure holding pond indicated that 90% of P was removed from approximately 140 000 L of swine manure slurry with the addition of 2 000 L MgCl2 (64% solution), at ambient slurry temperatures ranging from 5 to 10 ∘ C [110]. Yetilmezsoy and Sapci-Zengin [113] also found that the maximum NH4 + removal was obtained as 85.4% at a pH of 9.0 with the addition of MgCl2 ⋅6H2 O and KH2 PO4 at the stoichiometric ratio (Mg2+ : NH4 + –N : PO4 3− –P = 1 : 1 : 1). The increase of Mg/P molar ratio seems to stimulate the struvite crystallization. Song et al. [112] indicated that the P removal efficiency from swine wastewater via struvite formation increased by 20% when the Mg/P molar ratio increased from 1.0 to 2.0. The NH4 –N recovery efficiency also increased from 86.4 to 97.4% by struvite from anaerobic digestion residues of poultry manure by adjusting the Mg : N : P molar ratio from 1 : 1 : 1 to 1.5 : 1 : 1 [114]. However, high costs of the chemicals added in the process hinder its large-scale applications [115]. The soluble salt MgCl2 is the most frequently used magnesium source for such processes, but it is not cost-effective due to the high market price of MgCl2 (0.12$/kg) [116]. For this reason, researches have started to look for other Mg sources as substitutes and explore the combined technologies to reduce the cost for struvite precipitation. For instance, Lin et al. [117] indicated that the total chemical cost was saved by 29.17% compared to MgCl2 application by using the residue from magnesium-hydroxide flue gas desulfurization (MFGD) process as an Mg2+ source. An evaluation of economic feasibility by Huang et al. [118] showed that the cost for struvite precipitation could be reduced by approximately 12.4% using the combined

575

576

20 Approaches Toward Resource Recovery from Breeding Wastewater

treatment of swine wastewater with MgO-saponification wastewater in comparison with the use of pure magnesium salts. The combined process of ion-exchange and precipitation was also reported as a cost-effective method for recovering N and P simultaneously [100]. Huang et al. [103] indicated that MgCl2 -modified zeolite can be used as adsorbent material and magnesium source for struvite crystallization after Mg2+ was released from the adsorption process, the NH4 + and PO4 3− removal efficiency reached 82 and 98% respectively. 20.4.2.4 Membrane Filtration

With respect to resource recovery, membrane technologies can also offer huge potentials. Microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reserve osmosis (RO) are typical types of pressure-driven membrane processes used for the treatment of wastewater based on their pore size. Nutrients in particulate form >0.1 μm in size can be harvested after MF or UF separation. The liquid fraction from MF and UF membrane can contain large amounts of dissolved nutrients. NF and RO membrane can retain these dissolved nutrients and deliver high concentrate retentate and purified water [119]. The NF membrane has been reported to be capable of retention 5–52% of the NH4 + and 97–98% of P [120]. By contrast, RO has a higher concentration efficiency for NH4 + and P, which are 75–100 and 99.5%, respectively [121]. The retention of NH4 + by RO is very pH-dependent. For instance, the reduction of swine manure pH from 7.8 to 6.5 increased total ammonia nitrogen (TAN) retention from 94 to above 99% [121]. Bilstad et al. [122] also found that the retention of TAN by RO membranes fed anaerobically digested manure ranged from 75 to 96% at pH 8.0, and was nearly 100% when pH was decreased to 4.0. The main reason might be that the reduction of pH and temperature decreases the fraction of TAN present as NH3 , thereby increasing retention. Therefore, the use of RO for nutrient concentration and breeding waste treatment is technically feasible with proper pretreatment. Moreover, wastewater properly treated by membrane technologies can be reused for irrigation [123]. An integrated system with two-step UF followed by NF was operated by Konieczny et al. [124] to recover water from swine slurry. The produced water had a quality similar to the drinking water range for some components except for NH4 + , which happens to be a valuable nutrient for irrigation. The effective rejection of NH4 + and other ions such as sodium and chloride by RO increased the potential uses of breeding wastewater for irrigation [125]. Although MF and UF are normally used as the pretreatment methods in practice, membrane fouling is still severe [74]. Therefore, the high cost associated with the technology, such as membrane cleaning and replacement, high operating pressures required for concentrated feed, as well as costs of further fertilizer production, remains the main obstacle to the application of RO for nutrients recovery from breeding wastewater [120]. Emerging membrane processes, including forward osmosis (FO), membrane distillation (MD), and electrodialysis (ED), have high potential for nutrients recovery from breeding wastewater. As reviewed by Yan et al. [126], the rejection number of PO4 − by FO was greater than that of NH4 + because of the larger hydrated radius of

20.4 Approaches for Resource Recovery

PO4 − . The accumulated nutrient in feed can be recovered via struvite precipitation by supplementing magnesium cation into the feed when magnesium-based draw solution is used. The recovered products (struvite and water) have a potential value of 1.35$/m3 . Comparatively, FO has a lower fouling propensity and higher fouling reversibility than the pressure-driven RO membrane filtration [127]. Membrane distillation (MD) is a thermally driven membrane process that can utilize the vapor pressure gradient across a porous membrane to extract water and volatile compounds from the feed. In the MD process, components can only transport through the membrane in a vapor phase, while the nonvolatile constituent in the feed solution can be rejected completely by MD. Therefore, components can be concentrated either in the feed stream or permeate streams based on the volatility and vapor pressure [128]. The nonvolatile inorganic nutrient ions, such as K+ and PO4 − , can be completely rejected by MD and concentrated in the feed stream to facilitate subsequent nutrient precipitation. MD has been successfully used for ammonia removal/recovery from wastewater because ammonia is more volatile than water and can be enriched in the permeate stream of MD processes [129]. The volatile ammonia can be subsequently adsorbed by acid solution for ammonia fertilizer production [126]. Xie et al. [128] summarized that more than 96% ammonia recovery can be achieved in MD processes in the form of aqueous solution. A high pH (> 9.7), high temperature (> 45 ∘ C), and the application of acidic stripping solution in the MD permeate stream can favor the evaporation rate of NH3 [128]. However, the permeate stream and recovered ammonia fertilizer can be contaminated by the existence of other volatile organic compounds, such as volatile fatty acids. Additionally, MD membrane pore wetting and fouling can also hinder ammonia vapor permeation, but the fouling is less than that of pressure-driven membranes [129]. Electrodialysis (ED) is an electrically driven membrane technology with the utilization of ion-exchange membranes to comprise cation-selective, anion-selective, and bipolar membranes. In the electrical field, anions and cations move toward the anode and the cathode, respectively, and are obtained separately in concentrated solution [130]. The unique ion separation mechanism of ED process provides a selective mechanism for wastewater nutrient recovery [128]. PO4 3− and NH4 + can be selectively separated from the feed solution and migrate to anode chamber and cathode chamber respectively. Accordingly, the NH4 + in swine wastewater can be concentrated to 14–21 g/L in the ED process, which is higher than that in RO process (13 g/L) [128]. Yan et al. [126] also indicated that more than 80% of PO4 3− could be recovered from wastewater by the combination of ED process and chemical precipitation. However, only small-scale ED has been used. The advantages and disadvantages of above discussed physicochemical approaches have been listed in Table 20.3; although ammonia stripping can recover NH4 + efficiently under optimal condition, high energy is required and P is still retained in the solid fraction. The operation of ion-exchange and adsorption is easy, but sorbents can be blocked easily by the suspended solids, and the competition of other ions is severe, which can significantly reduce the recovery efficiencies. Struvite formation is a promising method due to its low energy consumption, whereas its recovery efficiency is limited by the extremely high concentration

577

578

20 Approaches Toward Resource Recovery from Breeding Wastewater

of NH4 + in livestock and poultry wastewater. To recover NH4 + effectively, large amount of Mg2+ and PO4 3− need to be added in the system. In comparison with MD and ED processes, FO approaches are more feasible for recovering nutrients from breeding wastewater considering their advantages of low energy consumption, high rejection rate for nutrient, and low membrane fouling potential [126].

20.4.3 Plant-Based Approaches Currently, plant-based approaches for wastewater treatment have been gaining significant attention worldwide, due to their dual advantages of wastewater treatment and resource recovery. Microalgae are the most studied aquatic plants. They are the most popular feedstock for biofuel and bioproducts production because of their high growth rates, superior environmental adaptability, no competition with food or arable land, and valuable biochemical composition [131]. Compared to plant crops, microalgae has higher biomass productivity, requires less land area and lower cost per yield. Water and inorganic nutrients are identified as important limiting resources for microalgae culture. According to the elemental composition of microalgal biomass, it is estimated that roughly 40–90 kg of N and 3–15 kg of P would be required to produce 1 t of algae for 100% uptake [132]. Fertilizers are usually utilized to supply these chemical nutrients, which would likely affect fertilizers market prices, further lowering the economics of algal biofuels production. Benemann and Oswald [133] found that the cost of nutrients in algae culture systems accounted for 10–20% of the total cost. Raw and digested breeding wastewater, alternatively, can provide a nutrient-rich environment for microalgae growth. In this case, nutrients can be recovered by microalgae as valuable biomass, which can be further extracted from the biomass and recycled for further production, reducing the net fertilizers input [132]. This is the reason why microalgae cultivation has been also proposed as a nutrient recovery approach. The use of breeding wastewater to supply the nutrient needs for biomass production improves the economics and environmental footprint of the whole production process [134]. Aquatic plants cultivation coupled to nutrients recovery in different raw and anaerobic digested breeding wastewaters has been extensively studied by previous researchers. Results from recent studies about the microalgae/duckweed biomass production and nutrient removal from breeding wastewater is summarized in Table 20.4. The growth of microalgae and duckweed and their removal capacity of nutrients in breeding wastewaters depend on many variables. Different results can be achieved in different types and compositions of wastewater. Swine, dairy and poultry farm effluent are comprised of high concentrations of COD, ammonium, and phosphates. High organic matter and nutrient concentrations in these wastewaters limit light penetration due to turbidity and often result in toxicity, negative impact on microalgae growth. In addition, high concentrations of ammonium, heavy metals, antibiotics, and hormones in undiluted livestock and poultry wastewater also inhibit the growth of microalgae [39]. Thus, dilution of livestock and poultry wastewater is usually required prior to using them as a microalgal growth medium. In addition, variation is also caused by the different tolerance capacity of

Table 20.4

Species

Spirodela oligorrhiza Spirodela polyrrhiza Spirodela polyrrhiza

Recent studies about the microalgae/duckweed biomass production and nutrient removal from breeding wastewater.

Conditions

Initial N,P concentration (mg/L)

Dry biomass TN yield (g/m2 /d) removal (%)

6% swine lagoon water, 60% NH4 –-N = 52.1, 0.687 g/g initial duckweed coverage, TP = 15.9 and harvest twice a week 28 ∘ C, photoperiod of 16 h NH4 –N = 75.1 4.68 light (80 μmol/m2 /s) and 8 h dark 12.4 NH4 –N = 20 20–30 ∘ C, light intensity 2.89 mmol/m2 /s

TP removal (%)

Starch content (%)

Protein content (%) References

83.7

89.4

–

–

[135]

100 (NH4 –N)

–

9.2

–

[136]

1.08 g/m2 /d (NH4 –N removal rate)

0.10 g/m2 /d (PO4 –P) removal rate)

18.8

25.4

[137]

[138]

NH4 –N = 20 20 ∘ C, light intensity 40 mol/m/s at a photoperiod of 12 h/d, harvested frequency of 1–3 times a week 20–30 ∘ C, light intensities NH4 –N = 1.5 40 mmol/m2 /s mmol/L

10.7

1.12 g/m2 /d (NH4 –N removal rate)

–

–

26.3

10.1

92.9 mmol/m2 /d (NH4 –N removal rate).

2.90 mmol/m2 /d (PO4 –P removal rate)

1.88 g/m2 /d (production rate)

2.68 g/m2 /d [139] (production rate)

Spirodela punctata 7776

–

NH4 –N = 240, PO4 –P = 31.0

31.92

0.955 mg/L/h (NH4 –N uptake)

0.129 mg/L/h (PO4 –P uptake)

–

–

[140]

Spirodela punctata 7776

24 ∘ C and 16 h of light per day

NH4 –N = 343, PO4 –P = 135

18.7

0. 802 g/m2 /day (NH4 –N removal rate)

0.24 g/m2 /day (PO4 –P removal rate)

–

–

[141]

Spirodela polyrrhiza

Spirodela polyrrhiza

(continued)

Table 20.4

Species

(Continued)

Conditions

Landoltia – punctata OT Landoltia punctata

–

Landoltia – punctata OT + Lemna minor OT

Initial N,P concentration (mg/L)

Dry biomass TN yield (g/m2 /d) removal (%)

TP removal (%)

Starch content (%)

Protein content (%) References

NH4 –N = 51, P = 11.2

16.2 in fresh 100 weight (NH4 –N)

99

16.7

30.1

[142]

NH4 –N = 75.1

3.23

100 (NH4 –N)

–

8.5

–

[136]

NH4 –N = 51, PO4 –P = 12

27.75

100

99

15.5

28.8

[142]

Landoltia punctata

–

NH4 –N = 336, P = 23.8

18

98

98.8

–

35

[143]

Lemna minor

–

NH4 –N = 75.1

2.79

100 (NH4 –N)

–

9.4

–

[136]

Lemna minor

Growth cycle of 15–18 d

NH4 –N = 56.1, PO4 –P = 16.3

3.5

100

74.8

10.6

32.2

[144]

Lemna minor OT

–

–

21.72

100

99

9.37

32.2

[142]

Lemna minor 8627

23 ∘ C, Light intensity 40 mmol/m2 /s at photoperiod of 12 h/d

–

29

3.36 g/m2 /d (N uptake rate)

0.20 g/m2 /d (P uptake rate)

–

–

[145]

Lemna minor

–

NH4 –N = 700–1900

3.1

0.14 g/m2 /d (N uptake rate)

0.035 g/m2 /d (P uptake rate)

–

–

[146]

NH4 –N = 75.1

3.07

100 (NH4 –N)

–

8.5

–

[136]

Lemna gibba –

20.4 Approaches for Resource Recovery

microalgae and duckweed species even in the same wastewater media. For example, microalgae strains of Chlorella have been studied enormously in different types of breeding wastewaters and shown to be effective in N and P removal with a wide range of initial concentrations. Moreover, cultivation conditions, including light intensity, pH, temperature, and cultivating modes can also affect the microalgae and duckweed growth. Generally, microalgae species has its own optimum range of pH, temperature, and light intensity at which it can function properly [147]. As reviewed previously, the optimal pH, temperature, and light intensity for microalgae growth in swine wastewater are in the range of 6–10, 20–27 ∘ C, and 45–300 μmol/m2 /s, respectively [39]. The nutrients removal and recovery from breeding wastewater is positive with the microalgae and duckweed growth. This is the reason why N and P are important nutrient sources required for the growth of microalgae and duckweed. Inorganic nitrogen and organic nitrogen (urea and amino acids) in breeding wastewaters are required by plants for the synthesis of proteins, nucleic acid, enzymes, chlorophylls, and genetic material [148]. Phosphorus mainly in the form of orthophosphate (HPO4− and H2 PO4–) is also a key factor in the energy metabolism of algae and duckweed and is found in nucleic acids, lipids, proteins, and the intermediates of carbohydrate metabolism [149]. Thereby, N and P in breeding wastewaters are mainly recovered by microalgae and duckweed via uptake [150]. Thus, the removal and recovery of nutrients from breeding wastewaters can also be affected by the microalgae and duckweed growth factors. Unlike the microbial process for generation of gases such as H2 or CH4 , the growth of microalgae and duckweed for bioenergy purposes does not typically result in an immediate source of fuel. Instead, they serve the role of providing biomass. Microalgae and duckweed harvested from breeding wastewater rich in carbohydrate, starch, protein, and lipids could be converted to multiple forms of bioenergy like biodiesel, bioethanol, H2 , and CH4 [152]. The conversion technology can be classified into chemical reaction, biochemical conversion, and thermochemical conversion (Figure 20.2) [151]. The choice of conversion approaches mainly includes the type and quantity of biomass feedstock, the desired form of the energy, economic consideration, project specificity, and the desired end form of the product [153]. As shown in Table 20.4, most strains of microalgae harvested from breeding wastewater are exceedingly rich in lipid under the optimal conditions (30–50% of dry weight biomass), which have high potential to be converted to biofuel, such as biodiesel, by transesterification [154]. Hena et al. [155] indicated that the biomass and lipid production of consortium cultivated in dairy wastewater were 153.54 t and 29 470 L/ha/yr, respectively, and 72.70% of algal lipid obtained from consortium could be converted into biodiesel. The physical and chemical properties of biodiesel produced from algal oil are similar to biodiesel produced from crops of first generation, to diesel from petroleum, and are favorable with International Biodiesel Standard for Vehicles (EN14214) [154]. The comparison of microalgae with other biodiesel sources has been summarized by Chisti [156], the oil yield of microalgae (58 700–136 900 L/ha) is much higher than that of other oil crops (from

581

582

20 Approaches Toward Resource Recovery from Breeding Wastewater

Chemical conversion

Aquatic biomass

Biochemical conversion

Thermochemic al conversion

Transesterification

Biodiesel

Anaerobic digestion

Biomethane

Dark fermentation

Biohydrogen

Pyrolysis

Bio-oil, biochar and syngas

Hydrothermal processing

Bio-oil

Gasification

Syngas

Figure 20.2 Options for biomass conversion processes of aquatic biomass into bioenergy. Source: Modified from Kaur et al. [151]. © Elsevier.

170 L/ha [corn] to 5950 L/ha [oil palm]) with the lowest land area requirement. Thus, the microalgae harvested from breeding wastewater are promising feedstock for biodiesel production without adversely affecting the food supply and other crop products. Some microalgal strains and duckweed, as well as the residual biomass after lipid extraction is rich in carbohydrates and protein, and thus serves as potential feedstock for H2 and CH4 production via dark fermentation and anaerobic digestion [157]. Lignocellulosic biomass, including corn stover, wheat straw, rice straw, sweet sorghum, and sugarcane bagasse, is the most extensively studied plant biomass for H2 production [158]. However, its conversion is challenging because hydrogen-producing microorganisms cannot utilize the major carbohydrates directly (cellulose and hemicellulose) in lignocellulosic biomass. Intensive thermochemical pretreatment and subsequent enzymatic hydrolysis have to be conducted before the conversion, which results in high operating costs. On the contrary, the carbohydrate content in microalgae and duckweed could be more readily used by microorganisms for H2 production, which would make the conversion process easier and cost-effective [157]. For example, Ferreira et al. [159] compared H2 and biogas production ability of Scenedesmus obliquus biomass harvested from different wastewater sources. The result indicated that the biomass from swine and poultry wastewater produced higher values for both hydrogen and biogas yields (around sevenfold), because of the high sugar content present in the biomass (36.2% for swine and 23.6% for poultry wastewaters, respectively). The H2 obtained from duckweed in swine wastewater was up to 75 mL/g dry duckweed in seven days, which also is comparable with other plant biomass [158]. Moreover, the anaerobic digestion of microalgae biomass harvested from breeding wastewater can produce higher CH4 than that harvested from other growth mediums. For instance, the suitability of fresh microalgae from swine wastewater as a substrate for biogas production was evaluated by Montingelli et al. [160]; the

20.4 Approaches for Resource Recovery

biogas production (364 mL/gVS ) and CH4 content (62.6% v/v) significantly compared to those of standard positive controls. Perazzoli et al. [161] also demonstrated that the CH4 produced from anaerobic digestion of microalgae polyculture from swine wastewater was 320–389 mL/(gVS ), higher than that from other growth mediums. Furthermore, the produced biogas can supply CO2 to encourage microalgae growth and lipid accumulation concomitantly upgrading the CH4 content of biogas [162]. Srinuanpan et al. [163] indicated that the content of CO2 in biogas can be continuously reduced through cultivation of oleaginous microalgae, thereby resulting in the high content of CH4 (>98%) in biogas. Microalgae and duckweed cultured in breeding wastewater are also potential feedstock for the production of bioethanol, because of their high carbohydrate, starch, and protein content that can be easily and effectively saccharified to glucose for fermentation [144]. Xu et al. [137] reported that the starch content of Spirodela polyrrhiza harvested from swine wastewater was up to 64.9% through simple transfer of fresh duckweed to well water for 10 days, achieving a high annual starch yield of 9.42 t/ha. After enzymatic hydrolysis and yeast fermentation of the high-starch duckweed, 94.7% of the theoretical starch-to-ethanol conversion was realized, resulting in an estimated ethanol yield of 6.42 × 103 L/ha/yr, about 50% higher than that of maize-based ethanol production. Thus, high-starch microalgae and duckweed harvested from breeding wastewater can be very readily converted for bioethanol production.

20.4.4 Thermochemical Approaches Breeding wastes and the aquatic biomass harvested from breeding wastewater can be converted into liquid, gaseous, and solid fuel products via thermochemical conversion [153, 154]. Thermochemical conversion is a high-temperature chemical reforming process technology including pyrolysis, hydrothermal processing, combustion, and gasification [164]. In addition to bioenergy production, the thermal process has the advantage of recovering nutrients from breeding wastes, eliminating odor and pathogens at high temperature, less greenhouse gas emission, small footprint, and short processing time [87, 165]. Pyrolysis is a thermal process of converting biomass feedstock into bio-oil, biochar, and non-condensable gases under 350–700 ∘ C in an anaerobic and pressurized environments [165]. The characteristic of products obtained as a result of pyrolysis mainly depends on biomass types, the operating modes (e.g. slow, fast, catalytic and microwave pyrolysis) and conditions (such as, temperature, heating rate, sweep gas flow rate and duration time) of pyrolysis, which have been reviewed and studied extensively by pervious researchers [154]. Chars are the dominant product from the thermal decomposition of biomass at lower pyrolysis temperatures [166]. The carbon content and thermal stability of biochars can be increased by increasing pyrolysis temperatures. The production of bio-oils usually happens at moderate temperatures (400–550 ∘ C) with short residence times (two to three seconds). The bio-oils produced from the pyrolysis of microalgae have higher oil yield and HHV

583

584

20 Approaches Toward Resource Recovery from Breeding Wastewater

compared with those from lignocellulosic biomass [167]. The gas product can be obtained at high temperature [166]. Comparatively, hydrothermal process offers advantages by converting wet biomass material into liquid fuel carried out at moderate temperature (280–370 ∘ C) and high pressures (14–22 MPa), thereby avoiding additional costs associated with drying [168, 169]. Considering the high moisture content of livestock manure and aquatic plants (>80%), hydrothermal process has been regarded as one of the most promising methods [170]. Life cycle analysis (LCA) of hydrothermal algal biomass confirmed that the process is a viable and economic option [171]. Target products (e.g., bio-oil, biochar, and gaseous products) from hydrothermal conversion of biomass mainly depend on the reaction temperature, reaction time, catalyst presence, and reactor pressure [172]. In the process of gasification, biomass is converted into a combustible gas mixture, containing varying amounts of H2 , CO, CO2 , and CH4 . The gas mixture is a low calorific gas (typical 4–6 MJ/m3 ) that can be burnt directly or used as a fuel for gas engines or gas turbines [173]. Accordingly, the gasification technology can be classified into conventional gasification and supercritical water gasification [174]. Dry biomass; high temperature (800–1000 ∘ C); and a gasification agent, such as steam, air, CO2 , O2 , or their mixtures, are required in the conventional gasification process. Therefore, the utilization of conventional gasification on breeding wastes and aquatic plants is less studied considering their high moisture contents, which can result in much heating energy consumption [153]. Comparatively, supercritical water gasification is a recent promising technology for converting wet biomass (70–95% moisture) to gas product directly, at the temperature of 375–550 ∘ C, and 22.1–36 MPa in the absence/presence of catalysts [174]. This process is more suitable for bioenergy production from breeding wastes and aquatic plants [175]. Ro et al. [176] found that swine manure generated high amounts of positive net energy by using wet gasification. Its conversion efficiency has been reported higher than other thermal technologies [151]. In short, using different thermochemical conversion routes, breeding wastes and the harvested aquatic plants can be converted to solid, liquid, and gas biofuels as substitute to fossil fuels. However, challenges associated with reducing energy costs and improving product qualities need to be overcome to achieve the application of the thermochemical approaches to breeding wastes for the production of bioenergy [152, 175]. The energy input requirements for wet gasification of swine manure have been concluded to be larger than those of a traditional anaerobic digestion operation [176]. Ro et al. [176] indicated that without use of an efficient heat recovery system, swine manure would not generate positive energy return from wet gasification.

20.5 Current Application and Future Perspectives The production of renewable resources and bioenergy has drawn great attention due to the increasing global resource and energy demand, the depletion of fossil fuel reserves, and the issue of greenhouse gas emissions [177]. Breeding wastes with high

20.5 Current Application and Future Perspectives

levels of organic wastes and nutrients have been regarded as a promising feedstock for resources and bioenergy recovery. In recent decades, industries in various countries have been encouraged to produce bioenergy (biogas, biodiesel, and bioelectricity), recovery nutrients, and reusable water from wastewater resources. AD is an efficient technology that combines biofuel production with sustainable waste management, and has robust commercial availability. Additionally, the digestate from AD system is a high-value nutrient and water resource for fertilizer production and irrigation. Europe is the global leader in this field, the production of biogas increased from 7.93 × 106 to 1.41 × 107 between 2009 and 2016 [177]. Germany is the largest producer in EU, followed by the United Kingdom, Italy, and France [178]. Breeding wastes with low cost and high availability can be treated for biogas production. For instance, animal wastes from medium- and large-scale livestock farms in China have been successfully used as feedstock for producing biogas. It is reported that the total annual production of animal wastes has reached 3 Gt in China, but most of the waste is not treated [179]. A case study by Lu et al. [179] estimated that if all the livestock manure were treated to produce biogas, the total energy would be 3.77 EJ, which could meet 5.23% of China’s energy consumption. Therefore, even though biogas has been produced via anaerobic digestion around the world, an increasing amount of biogas produced from animal wastes is required in future. Moreover, the successful application of breeding wastes for biogas production requires the improvement of government support, design, and construction of digesters, as well as the monitoring and maintenance of the project [180]. A variety of studies examined the H2 production from livestock and poultry wastes through dark fermentation, while implementation of this technology is still in its infancy [77]. As discussed previously, the main challenge in dark fermentation is the low hydrogen yield, since only 15% of the energy from the organic source is typically obtained in the form of hydrogen and remaining 67–85% of the substrate remains unused [181]. Therefore, further studies are required to enhance the hydrogen yield, which should focus on the suitable microbial strain for biohygrogen production, the modification of bioreactor and operating process to redirect the metabolic pathway [182]. The proper recovery of nutrients present in breeding wastewater is essential to achieve environmental sustainability, which has been conducted in full-scale applications. Baldi et al. [183] reported that 35–50% of N could be recovered from anaerobic digested livestock wastes by ammonia stripping in full-scale established plants. However, the high alkalinity effluent discharged from ammonia stripping process would give rise to another environmental issue without post-treatment [13]. Post-treatment by direct chemical neutralization requires considerable economic input. Comparatively, struvite precipitation process is a relatively eco-friendly process that recovers P and N from wastewaters, which has been successfully implemented at large scales and in commercially established processes currently [18]. Laridi et al. [184] concluded that around 99 and 15% of phosphate and ammonium in swine wastewater could be recovered via struvite precipitation at a pilot-scale application. A 3.5-year operation about P recovery from swine wastewater has been conducted by Suzuki et al. [185], 171 g struvite was obtained

585

586

20 Approaches Toward Resource Recovery from Breeding Wastewater

from 1 m3 swine wastewater. The recovered struvite was approximately 95% pure, and was ready for immediate application to farmland. In recent study, 15 integrated scenarios from full-scale pig manure treatment facilities have been compared in terms of technical feasibility and economic viability [12]. It was found that nitrogen recovery through ammonia stripping/absorption, and phosphorus recovery through struvite crystallization showed a higher market value and quality compared to basic composting, but coincided with higher net costs. The wide application of membrane filtration technologies (e.g. FO approach by using MgCl2 as draw solution) has been limited by membrane fouling problem. Microalgae- and duckweed-based bioenergy has been long considered as a promising substitute to replace the fossil fuels at large scale. Breeding wastewater potentially represents one example of valuable and sustainable water and nutrient resources to cultivate these aquatic plants for bioproduct production, because it will be generated continually as long as human beings exist. To date, some studies at bench scale have been conducted to evaluate the performance of microalgae and duckweed growth in breeding wastewater for biomass production and nutrients removal under field conditions. Xu et al. [138] investigated the growth of duckweed in swine wastewater at a pilot-scale pond, and concluded that the duckweed (Spirodela polyrrhiza) grew rapidly under field conditions, with a dry biomass yield and NH4 –N removal of 10.7 and 1.12 g/m2 /d, respectively. The main challenges and issues of these plant-based approaches, including the pretreatment of breeding wastewater, the selection of suitable species for cultivating in breeding wastewater, the optimization of cultivation conditions, the harvest and pretreatment method of biomass, as well as the further conversion technologies for bioenergy and bioproducts production, also require further study. The application of thermochemical approaches might not be encouraged due to the huge energy input, unless special products are required. Furthermore, the development of integrated approaches for enhancing the resource recovery and the quality of products in a cost-effective way is significant in future research.

20.6 Conclusion The characteristics of various breeding wastewater and potential resources in them have been reviewed in this chapter. The continuously discharged breeding wastewater with high concentrations of organic pollutants and nutrients is a major threat to environmental safety. Meanwhile, breeding wastewater is a huge resource bringing many economic and environmental benefits if recovered properly. Resources including reusable water, nutrients, bioenergy, and other bioproducts are potentially to be recovered from breeding wastewater. The advantages of using bioenergy and bioproducts over conventional fossil fuels have been pointed out explicitly in this chapter. Obviously, the development of suitable technologies for recovering the valuable resources from breeding wastewater is significant for alleviating resources and energy crisis in current society and reducing the environmental threats from traditional breeding waste management practices. Anaerobic digestion is a widely

References

used and mature approach for CH4 production from breeding wastes treatment, which has been applied in full-scale plants. Whereas, more studies are required to optimize the bioreactor configuration and operating parameters to enhance CH4 production. Nutrients recovered from breeding wastewater through physicochemical approaches, such as ammonia stripping, ion-exchange and adsorption, struvite formation, and membrane filtrations, have been reviewed in this chapter; overall, all these approaches have their own advantages and drawbacks requiring further studies. Plant-based approaches have advantages for treating wastewater, nutrients recovery, and bioenergy production. Although several operations have moved from bench-scale to small-pilot operations, further research is still needed to overcome those limits and challenges to update the technology to industrial level.

References 1 Roser, M. and Ritchie, H. (2017). Meat and Dairy Production. Our World in Data. 2 Dauda, A., Folorunso, L., and Dasuki, A. (2013). Use of probiotics for sustainable aquaculture production in Nigeria. Journal of Agriculture and Social Research 13 (2): 42–52. 3 Ramankutty, N., Mehrabi, Z., Waha, K. et al. (2018). Trends in global agricultural land use: implications for environmental health and food security. Annual Review of Plant Biology 69: 789–815. 4 Mallin, M.A., McIver, M.R., Robuck, A.R. et al. (2015). Industrial swine and poultry production causes chronic nutrient and fecal microbial stream pollution. Water, Air, & Soil Pollution 226 (12): 407. 5 Food and Agriculture Organization of the United Nations (2018). Livestock manure statistics. http://www.fao.org/faostat/en/#data/EMN/visualize (accessed 10 January 2019). 6 Tullo, E., Finzi, A., and Guarino, M. (2018). Environmental impact of livestock farming and precision livestock farming as a mitigation strategy. Science of the Total Environment 650: 2751–2760. 7 Szogi, A.A., Vanotti, M.B., and Ro, K.S. (2015). Methods for treatment of animal manures to reduce nutrient pollution prior to soil application. Current Pollution Reports 1 (1): 47–56. 8 Marinho-Soriano, E., Nunes, S., Carneiro, M. et al. (2009). Nutrients’ removal from aquaculture wastewater using the macroalgae Gracilaria birdiae. Biomass and Bioenergy 33 (2): 327–331. 9 Skaggs, R.L., Coleman, A.M., Seiple, T.E. et al. (2018). Waste–to–energy biofuel production potential for selected feedstocks in the conterminous United States. Renewable and Sustainable Energy Reviews 82: 2640–2651. 10 US Department of Energy (2019). U.S. inherent energy content of feedstock by type 2017. https://www.statista.com/statistics/692096/energy-content-of-usfeedstock-by-type (accessed 13 January 2019).

587

588

20 Approaches Toward Resource Recovery from Breeding Wastewater

11 Hülsen, T., Hsieh, K., Tait, S. et al. (2018). White and infrared light continuous photobioreactors for resource recovery from poultry processing wastewater – a comparison. Water Research 144: 665–676. 12 De Vrieze, J., Colica, G., Pintucci, C. et al. (2019). Resource recovery from pig manure via an integrated approach: a technical and economic assessment for full–scale applications. Bioresource Technology 272: 582–593. 13 Ma, H.Y., Guo, Y., Qin, Y. et al. (2018). Nutrient recovery technologies integrated with energy recovery by waste biomass anaerobic digestion. Bioresource Technology 269: 520–531. 14 Steinfeld, H., Gerber, P., Wassenaar, T.D. et al. (2006). Livestock’s Long Shadow: Environmental Issues and Options. Food & Agriculture Org. 15 United States Department of Agriculture (2018). Livestock and Poultry: World Markets and Trade. Foreign Agricultural Service. 16 Muhmood, A., Wu, S., Lu, J. et al. (2018). Nutrient recovery from anaerobically digested chicken slurry via struvite: performance optimization and interactions with heavy metals and pathogens. Science of the Total Environment 635: 1–9. 17 Duda, R.M., da Silva Vantini, J., Martins, L.S. et al. (2015). A balanced microbiota efficiently produces methane in a novel high–rate horizontal anaerobic reactor for the treatment of swine wastewater. Bioresource Technology 197: 152–160. 18 Muhmood, A., Lu, J.X., Dong, R.J. et al. (2019). Formation of struvite from agricultural wastewaters and its reuse on farmlands: status and hindrances to closing the nutrient loop. Journal of Environmental Management 230: 1–13. 19 Markou, G., Iconomou, D., and Muylaert, K. (2016). Applying raw poultry litter leachate for the cultivation of Arthrospira platensis and Chlorella vulgaris. Algal Research 13: 79–84. 20 van Rijn, J. (2013). Waste treatment in recirculating aquaculture systems. Aquacultural Engineering 53: 49–56. 21 Ansari, F.A., Singh, P., Guldhe, A. et al. (2017). Microalgal cultivation using aquaculture wastewater: integrated biomass generation and nutrient remediation. Algal Research 21: 169–177. 22 Boopathy, R., Bonvillain, C., Fontenot, Q. et al. (2007). Biological treatment of low–salinity shrimp aquaculture wastewater using sequencing batch reactor. International Biodeterioration & Biodegradation 59 (1): 16–19. 23 Knapp, J. (2015). Impact of trace mineral variation within forages on the ration formulation process. https://hdl.handle.net/1813/41215. 24 Atuanya, E.I. and Aigbirior, M. (2002). Mesophilic biomethanation and treatment of poultry waste–water using pilot scale UASB reactor. Environmental Monitoring and Assessment 77 (2): 139–147. 25 Kalyuzhnyi, S., Fedorovich, V., and Nozhevnikova, A. (1998). Anaerobic treatment of liquid fraction of hen manure in UASB reactors. Bioresource Technology 65: 221–225. 26 FAO. 2018. The State of World Fisheries and Aquaculture 2018 – Meeting the sustainable development goals. Rome. Licence: CC BY-NC-SA 3.0 IGO: 1-227.

References

27 Food and Agriculture organization of the United Nations (2008). A framework for land evaluation. Soils Bulletin 32. 28 Crab, R., Avnimelech, Y., Defoirdt, T. et al. (2007). Nitrogen removal techniques in aquaculture for a sustainable production. Aquaculture 270 (1–4): 1–14. 29 Jegatheesan, V., Shu, L., and Visvanathan, C. (2011). Aquaculture effluent: impacts and remedies for protecting the environment and human health. Encyclopedia of environmental health: 123–135. http://hdl.handle.net/10536/DRO/ DU:30039572. 30 Bauen, A., Berndes G., Junginger M., et al. (2009). Bioenergy: a sustainable and reliable energy source. A review of status and prospects. https://www .cabdirect.org/cabdirect/abstract/20113044518. 31 Dauda, A.B., Ajadi, A., Tola-Fabunmi, A.S. et al. (2019). Waste production in aquaculture: sources, components and managements in different culture systems. Aquaculture and Fisheries 4 (3): 81–88. 32 Inyinbor, A.A., Bello, O.S., Oluyori, A.P. et al. (2018). Wastewater conservation and reuse in quality vegetable cultivation: overview, challenges and future prospects. Food Control 98: 489–500. 33 Hu, Y., Cheng, H., and Tao, S. (2017). Environmental and human health challenges of industrial livestock and poultry farming in China and their mitigation. Environment International 107: 111–130. 34 Seiler, C. and Berendonk, T.U. (2012). Heavy metal driven co–selection of antibiotic resistance in soil and water bodies impacted by agriculture and aquaculture. Frontiers in Microbiology 3: 399. 35 Norton-Brandão, D., Scherrenberg, S.M., and van Lier, J.B. (2013). Reclamation of used urban waters for irrigation purposes – a review of treatment technologies. Journal of Environmental Management 122: 85–98. 36 Becerra–Castro, C., Lopes, A.R., Vaz–Moreira, I. et al. (2015). Wastewater reuse in irrigation: a microbiological perspective on implications in soil fertility and human and environmental health. Environment International 75: 117–135. 37 Adnan, A., Mavinic, D.S., and Koch, F.A. (2003). Pilot–scale study of phosphorus recovery through struvite crystallization examining the process feasibility. Journal of Environmental Engineering and Science 2 (5): 315–324. 38 Bhatnagar, A. and Sillanpää, M. (2011). A review of emerging adsorbents for nitrate removal from water. Chemical Engineering Journal 168 (2): 493–504. 39 Cheng, D., Ngo, H., Guo, W. et al. (2019). Microalgae biomass from swine wastewater and its conversion to bioenergy. Bioresource Technology 275: 109–122. 40 Xing, Y., Li, Z., Fan, Y. et al. (2010). Biohydrogen production from dairy manures with acidification pretreatment by anaerobic fermentation. Environmental Science and Pollution Research 17 (2): 392–399. 41 Zhu, L.–.D. and Hiltunen, E. (2016). Application of livestock waste compost to cultivate microalgae for bioproducts production: a feasible framework. Renewable and Sustainable Energy Reviews 54: 1285–1290.

589

590

20 Approaches Toward Resource Recovery from Breeding Wastewater

42 Sarkar, N., Ghosh, S.K., Bannerjee, S. et al. (2012). Bioethanol production from agricultural wastes: an overview. Renewable Energy 37 (1): 19–27. 43 Pittman, J.K., Dean, A.P., and Osundeko, O. (2011). The potential of sustainable algal biofuel production using wastewater resources. Bioresource Technology 102: 17–25. 44 Puyol, D., Batstone, D.J., Hülsen, T. et al. (2017). Resource recovery from wastewater by biological technologies: opportunities, challenges, and prospects. Frontiers in Microbiology 7: 2106. 45 Duku, M.H., Gu, S., and Hagan, E.B. (2011). Biochar production potential in Ghana – a review. Renewable and Sustainable Energy Reviews 15 (8): 3539–3551. 46 Cao, X.D. and Harris, W. (2010). Properties of dairy–manure–derived biochar pertinent to its potential use in remediation. Bioresource Technology 101: 5222–5228. 47 Gunamantha, I.M. and Widana, G.A.B. (2018). Characterization the potential of biochar from cow and pig manure for geoecology application. International Conference on Green Agro-Industry and Bioeconomy 131 (1): 012055. 48 Wilkie, A.C. and Mulbry, W.W. (2002). Recovery of dairy manure nutrients by benthic freshwater algae. Bioresource Technology 84: 81–91. 49 Madeira, M.S., Cardoso, C., Lopes, P.A. et al. (2017). Microalgae as feed ingredients for livestock production and meat quality: a review. Livestock Science 205: 111–121. 50 Vasco–Correa, J., Khanal, S., Manandhar, A. et al. (2018). Anaerobic digestion for bioenergy production: global status, environmental and techno–economic implications, and government policies. Bioresource Technology 247: 1015–1026. 51 Dueblein, D. and Steinhauser, A. (eds.) (2008). Biogas from Waste and Renewable Resources. Chichester: Wiley. 52 Hill, D.T. and Bolte, J.P. (2000). Methane production from low solid concentration liquid swine waste using conventional anaerobic fermentation. Bioresource Technology 74: 241–247. 53 Riano, B., Molinuevo, B., and Garcia–Gonzalez, M.C. (2011). Potential for methane production from anaerobic co–digestion of swine manure with winery wastewater. Bioresource Technology 102: 4131–4136. 54 Comino, E., Riggio, V.A., and Rosso, M. (2012). Biogas production by anaerobic co–digestion of cattle slurry and cheese whey. Bioresource Technology 114: 46–53. 55 Demirer, G.N. and Chen, S. (2005). Two–phase anaerobic digestion of unscreened dairy manure. Process Biochemistry 40 (11): 3542–3549. 56 Dareioti, M.A. and Kornaros, M. (2015). Anaerobic mesophilic co–digestion of ensiled sorghum, cheese whey and liquid cow manure in a two–stage CSTR system: effect of hydraulic retention time. Bioresource Technology 175: 553–562. 57 Song, M., Shin, S.G., and Hwang, S. (2010). Methanogenic population dynamics assessed by real–time quantitative PCR in sludge granule in upflow anaerobic sludge blanket treating swine wastewater. Bioresource Technology 101: S23–S28.

References

58 Lo, K., Liao, P., and Gao, Y. (1994). Anaerobic treatment of swine wastewater using hybrid UASB reactors. Bioresource Technology 47 (2): 153–157. 59 Yang, H.N., Deng, L.W., Liu, G.J. et al. (2016). A model for methane production in anaerobic digestion of swine wastewater. Water Research 102: 464–474. 60 Mazareli, R.C.D., Duda, R.M., Leite, V.D. et al. (2016). Anaerobic co–digestion of vegetable waste and swine wastewater in high–rate horizontal reactors with fixed bed. Waste Management 52: 112–121. 61 Lopez–Fernandez, R., Aristizabal, C., and Irusta, R. (2011). Ultrafiltration as an advanced tertiary treatment of anaerobically digested swine manure liquid fraction: a practical and theoretical study. Journal of Membrane Science 375 (1–2): 268–275. 62 Bu, F., Du, S.Y., Xie, L. et al. (2017). Swine manure treatment by anaerobic membrane bioreactor with carbon, nitrogen and phosphorus recovery. Water Science and Technology 76 (8): 1939–1949. 63 Lee, S.M., Jung, J.Y., and Chung, Y.C. (2001). Novel method for enhancing permeate flux of submerged membrane system in two–phase anaerobic reactor. Water Research 35 (2): 471–477. 64 Arimi, M.M., Knodel, J., Kiprop, A. et al. (2015). Strategies for improvement of biohydrogen production from organic–rich wastewater: a review. Biomass and Bioenergy 75: 101–118. 65 Kanai, M., Ferre, V., Wakahara, S. et al. (2010). A novel combination of methane fermentation and MBR—Kubota submerged anaerobic membrane bioreactor process. Desalination 250 (3): 964–967. 66 Smith, A.L., Stadler, L.B., Cao, L. et al. (2014). Navigating wastewater energy recovery strategies: a life cycle comparison of anaerobic membrane bioreactor and conventional treatment systems with anaerobic digestion. Environmental Science & Technology 48 (10): 5972–5981. 67 Deng, L., Yang, H., Liu, G. et al. (2014). Kinetics of temperature effects and its significance to the heating strategy for anaerobic digestion of swine wastewater. Applied Energy 134: 349–355. 68 Zhai, N., Zhang, T., Yin, D. et al. (2015). Effect of initial pH on anaerobic co–digestion of kitchen waste and cow manure. Waste Management 38: 126–131. 69 Vedrenne, F., Béline, F., Dabert, P. et al. (2008). The effect of incubation conditions on the laboratory measurement of the methane producing capacity of livestock wastes. Bioresource Technology 99: 146–155. 70 Callaghan, F., Wase, D., Thayanithy, K. et al. (2002). Continuous co–digestion of cattle slurry with fruit and vegetable wastes and chicken manure. Biomass and Bioenergy 22 (1): 71–77. 71 Cordoba, V., Fernandez, M., and Santalla, E. (2016). The effect of different inoculums on anaerobic digestion of swine wastewater. Journal of Environmental Chemical Engineering 4 (1): 115–122. 72 Sakar, S., Yetilmezsoy, K., and Kocak, E. (2009). Anaerobic digestion technology in poultry and livestock waste treatment – a literature review. Waste Management & Research 27 (1): 3–18.

591

592

20 Approaches Toward Resource Recovery from Breeding Wastewater

73 Christensen, M.L., Hjorth, M., and Keiding, K. (2009). Characterization of pig slurry with reference to flocculation and separation. Water Research 43 (3): 773–783. 74 Aspe, E., Marti, M.C., and Roeckel, M. (1997). Anaerobic treatment of fishery wastewater using a marine sediment inoculum. Water Research 31 (9): 2147–2160. 75 Turon, V., Trably, E., Fouilland, E. et al. (2016). Potentialities of dark fermentation effluents as substrates for microalgae growth: a review. Process Biochemistry 51 (11): 1843–1854. 76 Kotsopoulos, T.A., Fotidis, I.A., Tsolakis, N. et al. (2009). Biohydrogen production from pig slurry in a CSTR reactor system with mixed cultures under hyper–thermophilic temperature (70 C). Biomass and Bioenergy 33 (9): 1168–1174. 77 Guo, X.M., Trably, E., Latrille, E. et al. (2010). Hydrogen production from agricultural waste by dark fermentation: a review. International Journal of Hydrogen Energy 35 (19): 10660–10673. 78 Kinidi, L., Tan, I.A.W., Wahab, N.B.A. et al. (2018). Recent development in ammonia stripping process for industrial wastewater treatment. International Journal of Chemical Engineering, article ID 3181087. 79 Ho, S.–.H., Nagarajan, D., Ren, N.q. et al. (2018). Waste biorefineries – integrating anaerobic digestion and microalgae cultivation for bioenergy production. Current Opinion in Biotechnology 50: 101–110. 80 Mohan, S.V., Babu, V.L., and Sarma, P. (2008). Effect of various pretreatment methods on anaerobic mixed microflora to enhance biohydrogen production utilizing dairy wastewater as substrate. Bioresource Technology 99 (1): 59–67. 81 Lee, M., Hidaka, T., and Tsuno, H. (2008). Effect of temperature on performance and microbial diversity in hyperthermophilic digester system fed with kitchen garbage. Bioresource Technology 99 (15): 6852–6860. 82 Zhu, J., Li Y., Wu X., et al. (2008). Swine manure fermentation to produce biohydrogen. The 30th Symposium on Biotechnology for Fuels and Chemicals. New Orleans, LO (4−7 May 2008). 83 Wang, L., Addy, M., Liu, J. et al. (2019). Integrated process for anaerobically digested swine manure treatment. Bioresource Technology 273: 506–514. 84 Villamar, C.–.A., Vera–Puerto, I., Rivera, D. et al. (2018). Reuse and recycling of livestock and municipal wastewater in Chilean agriculture: a preliminary assessment. Water 10 (6): 817. 85 Jia, S., Zhang, X.–.X., Miao, Y. et al. (2017). Fate of antibiotic resistance genes and their associations with bacterial community in livestock breeding wastewater and its receiving river water. Water Research 124: 259–268. 86 Lei, X.H., Sugiura, N., Feng, C.P. et al. (2007). Pretreatment of anaerobic digestion effluent with ammonia stripping and biogas purification. Journal of Hazardous Materials 145 (3): 391–397. 87 Shi, L., Simplicio, W.S., Wu, G.X. et al. (2018). Nutrient recovery from Digestate of anaerobic digestion of livestock manure: a review. Current Pollution Reports 4 (2): 74–83.

References

88 Gustin, S. and Marinsek–Logar, R. (2011). Effect of pH, temperature and air flow rate on the continuous ammonia stripping of the anaerobic digestion effluent. Process Safety and Environmental Protection 89 (1): 61–66. 89 Bonmat𝚤, A. and Flotats, X. (2003). Air stripping of ammonia from pig slurry: characterisation and feasibility as a pre–or post–treatment to mesophilic anaerobic digestion. Waste Management 23 (3): 261–272. 90 Liao, P.H., Chen, A., and Lo, K.V. (1995). Removal of nitrogen from swine manure wastewaters by ammonia stripping. Bioresource Technology 54 (1): 17–20. 91 Zhang, L., Lee, Y.W., and Jahng, D. (2012). Ammonia stripping for enhanced biomethanization of piggery wastewater. Journal of Hazardous Materials 199: 36–42. 92 Limoli, A., Langone, M., and Andreottola, G. (2016). Ammonia removal from raw manure digestate by means of a turbulent mixing stripping process. Journal of Environmental Management 176: 1–10. 93 Park, S. and Kim, M. (2015). Innovative ammonia stripping with an electrolyzed water system as pretreatment of thermally hydrolyzed wasted sludge for anaerobic digestion. Water Research 68: 580–588. 94 Ukwuani, A.T. and Tao, W.D. (2016). Developing a vacuum thermal stripping – acid absorption process for ammonia recovery from anaerobic digester effluent. Water Research 106: 108–115. 95 Huang, H.M., Yang, L.P., Xue, Q. et al. (2015). Removal of ammonium from swine wastewater by zeolite combined with chlorination for regeneration. Journal of Environmental Management 160: 333–341. 96 Kumar, S. and Jain, S. (2013). History, introduction, and kinetics of ion exchange materials. Journal of Chemistry 2013: 1, , Article ID 957647–13. 97 Wang, S. and Peng, Y. (2010). Natural zeolites as effective adsorbents in water and wastewater treatment. Chemical Engineering Journal 156 (1): 11–24. ´ ´ 98 Zielinski, M., Zielinska, M., and De˛bowski, M. (2016). Ammonium removal on zeolite modified by ultrasound. Desalination and Water Treatment 57 (19): 8748–8753. 99 Zhan, Y., Zhang, H., Lin, J. et al. (2017). Role of zeolite’s exchangeable cations in phosphate adsorption onto zirconium–modified zeolite. Journal of Molecular Liquids 243: 624–637. 100 Amini, A., Aponte–Morales, V., Wang, M. et al. (2017). Cost–effective treatment of swine wastes through recovery of energy and nutrients. Waste Management 69: 508–517. 101 Guo, J. (2016). Adsorption characteristics and mechanisms of high–levels of ammonium from swine wastewater using natural and MgO modified zeolites. Desalination and Water Treatment 57 (12): 5452–5463. 102 Guo, X., Zeng, L., Li, X. et al. (2007). Removal of ammonium from RO permeate of anaerobically digested wastewater by natural zeolite. Separation Science and Technology 42 (14): 3169–3185. 103 Huang, H.M., Xiao, D., Pang, R. et al. (2014). Simultaneous removal of nutrients from simulated swine wastewater by adsorption of modified zeolite

593

594

20 Approaches Toward Resource Recovery from Breeding Wastewater

104

105

106

107

108

109

110

111

112

113

114

115

116

117

combined with struvite crystallization. Chemical Engineering Journal 256: 431–438. Montalvo, S., Guerrero, L., Borja, R. et al. (2012). Application of natural zeolites in anaerobic digestion processes: a review. Applied Clay Science 58: 125–133. Wang, Q., Yang, Y., Yu, C. et al. (2011). Study on a fixed zeolite bioreactor for anaerobic digestion of ammonium–rich swine wastes. Bioresource Technology 102 (14): 7064–7068. Wasielewski, S., Rott, E., Minke, R., and Steinmetz, H. (2018). Evaluation of different clinoptilolite zeolites as adsorbent for ammonium removal from highly concentrated synthetic wastewater. Water 10 (5): 584. Tao, W.D., Fattah, K.P., and Huchzermeier, M.P. (2016). Struvite recovery from anaerobically digested dairy manure: a review of application potential and hindrances. Journal of Environmental Management 169: 46–57. Uludag–Demirer, S., Demirer, G.N., and Chen, S. (2005). Ammonia removal from anaerobically digested dairy manure by struvite precipitation. Process Biochemistry 40 (12): 3667–3674. Stratful, I., Scrimshaw, M.D., and Lester, J.N. (2001). Conditions influencing the precipitation of magnesium ammonium phosphate. Water Research 35 (17): 4191–4199. Burns, R.T., Moody, L.B., Walker, F.R. et al. (2001). Laboratory and in–situ reductions of soluble phosphorus in swine waste slurries. Environmental Technology 22 (11): 1273–1278. Suzuki, K., Tanaka, Y., Osada, T. et al. (2002). Removal of phosphate, magnesium and calcium from swine wastewater through crystallization enhanced by aeration. Water Research 36 (12): 2991–2998. Song, Y.H., Yuan, P., Zheng, B.H. et al. (2007). Nutrients removal and recovery by crystallization of magnesium ammonium phosphate from synthetic swine wastewater. Chemosphere 69 (2): 319–324. Yetilmezsoy, K. and Sapci–Zengin, Z. (2009). Recovery of ammonium nitrogen from the effluent of UASB treating poultry manure wastewater by MAP precipitation as a slow release fertilizer. Journal of Hazardous Materials 166 (1): 260–269. Yilmazel, Y. and Demirer, G. (2011). Removal and recovery of nutrients as struvite from anaerobic digestion residues of poultry manure. Environmental Technology 32 (7): 783–794. Zhang, T., Ding, L., Ren, H. et al. (2009). Ammonium nitrogen removal from coking wastewater by chemical precipitation recycle technology. Water Research 43 (20): 5209–5215. Suzuki, K., Tanaka, Y., Kuroda, K. et al. (2005). Recovery of phosphorous from swine wastewater through crystallization. Bioresource Technology 96 (14): 1544–1550. Lin, H., Lin, Y., Wang, D. et al. (2018). Ammonium removal from digested effluent of swine wastewater by using solid residue from

References

118

119

120

121

122

123

124

125

126

127

128

129

130 131

magnesium–hydroxide flue gas desulfurization process. Journal of Industrial and Engineering Chemistry 58: 148–154. Huang, H.M., Chen, Y.Q., Jiang, Y. et al. (2014). Treatment of swine wastewater combined with MgO–saponification wastewater by struvite precipitation technology. Chemical Engineering Journal 254: 418–425. Mehta, C.M., Khunjar, W.O., Nguyen, V. et al. (2015). Technologies to recover nutrients from waste streams: a critical review. Critical Reviews in Environmental Science and Technology 45 (4): 385–427. Masse, L., Masse, D.I., and Pellerin, Y. (2007). The use of membranes for the treatment of manure: a critical literature review. Biosystems Engineering 98 (4): 371–380. Masse, L., Masse, D.I., and Pellerin, Y. (2008). The effect of pH on the separation of manure nutrients with reverse osmosis membranes. Journal of Membrane Science 325 (2): 914–919. Bilstad, T., Madland, M., Espedal, E. et al. (1992). Membrane separation of raw and anaerobically digested pig manure. Water Science and Technology 25 (10): 19–26. Quist–Jensen, C.A., Macedonio, F., and Drioli, E. (2015). Membrane technology for water production in agriculture: desalination and wastewater reuse. Desalination 364: 17–32. Konieczny, K., Kwiecinska, A., and Gworek, B. (2011). The recovery of water from slurry produced in high density livestock farming with the use of membrane processes. Separation and Purification Technology 80 (3): 490–498. Guo, X.J. and Jin, X. (2014). Purification of UF–treated anaerobically digested manure wastewater by two–pass reverse osmosis. Desalination and Water Treatment 52 (16–18): 3027–3034. Yan, T., Ye, Y.Y., Ma, H.M. et al. (2018). A critical review on membrane hybrid system for nutrient recovery from wastewater. Chemical Engineering Journal 348: 143–156. Lee, S., Boo, C., Elimelech, M. et al. (2010). Comparison of fouling behavior in forward osmosis (FO) and reverse osmosis (RO). Journal of Membrane Science 365 (1–2): 34–39. Xie, M., Shon, H.K., Gray, S.R. et al. (2016). Membrane–based processes for wastewater nutrient recovery: technology, challenges, and future direction. Water Research 89: 210–221. Zarebska, A., Nieto, D.R., Christensen, K.V. et al. (2014). Ammonia recovery from agricultural wastes by membrane distillation: fouling characterization and mechanism. Water Research 56: 1–10. Xu, T.W. and Huang, C.H. (2008). Electrodialysis–based separation technologies: a critical review. AICHE Journal 54 (12): 3147–3159. Chiu, S.Y., Kao, C.Y., Chen, T.Y. et al. (2015). Cultivation of microalgal chlorella for biomass and lipid production using wastewater as nutrient resource. Bioresource Technology 184: 179–189.

595

596

20 Approaches Toward Resource Recovery from Breeding Wastewater

132 Barbera, E., Bertucco, A., and Kumar, S. (2018). Nutrients recovery and recycling in algae processing for biofuels production. Renewable & Sustainable Energy Reviews 90: 28–42. 133 Benemann, J.R. and Oswald, W.J. (1994). Systems and Economic Analysis of Microalgae Ponds for Conversion of CO2 to Biomass. New York: NASA STI/Recon Technical Report N. 134 Muylaert, K., Beuckels, A., Depraetere, O. et al. (2015). Wastewater as a source of nutrients for microalgae biomass production. In: Biomass and Biofuels from Microalgae, 75–94. Springer. 135 Xu, J.L. and Shen, G.X. (2011). Growing duckweed in swine wastewater for nutrient recovery and biomass production. Bioresource Technology 102 (2): 848–853. 136 Toyama, T., Hanaoka, T., Tanaka, Y. et al. (2018). Comprehensive evaluation of nitrogen removal rate and biomass, ethanol, and methane production yields by combination of four major duckweeds and three types of wastewater effluent. Bioresource Technology 250: 464–473. 137 Xu, J.L., Cui, W.H., Cheng, J.J. et al. (2011). Production of high–starch duckweed and its conversion to bioethanol. Biosystems Engineering 110 (2): 67–72. 138 Xu, J., Cheng, J.J., and Stomp, A.M. (2012). Nutrient removal from swine wastewater by growing duckweed: a pilot study. Transactions of the ASABE 55 (1): 181–185. 139 Xu, J.L., Cheng, J.J., and Stomp, A.M. (2012). Growing Spirodela polyrrhiza in swine wastewater for the production of animal feed and fuel ethanol: a pilot study. Clean–Soil Air Water 40 (7): 760–765. 140 Cheng, J.Y., Bergmann, B.A., Classen, J.J. et al. (2002). Nutrient recovery from swine lagoon water by Spirodela punctata. Bioresource Technology 81 (1): 81–85. 141 Chaiprapat, S., Cheng, J.J., Classen, J.J. et al. (2005). Role of internal nutrient storage in duckweed growth for swine wastewater treatment. Transactions of the ASAE 48 (6): 2247–2258. 142 Zhao, Z., Shi, H.J., Liu, Y. et al. (2014). The influence of duckweed species diversity on biomass productivity and nutrient removal efficiency in swine wastewater. Bioresource Technology 167: 383–389. 143 Mohedano, R.A., Costa, R.H.R., Tavares, F.A. et al. (2012). High nutrient removal rate from swine wastes and protein biomass production by full–scale duckweed ponds. Bioresource Technology 112: 98–104. 144 Ge, X.M., Zhang, N.N., Phillips, G.C. et al. (2012). Growing Lemna minor in agricultural wastewater and converting the duckweed biomass to ethanol. Bioresource Technology 124: 485–488. 145 Cheng, J., Landesman, L., Bergmann, B.A. et al. (2002). Nutrient removal from swine lagoon liquid by Lemna minor 8627. Transactions of the ASAE 45 (4): 1003–1010. 146 Pena, L., Oliveira, M., Fragoso, R. et al. (2017). Potential of duckweed for swine wastewater nutrient removal and biomass valorisation through anaerobic

References

147

148

149

150

151

152 153 154

155

156 157

158 159

160 161

co–digestion. Journal of Sustainable Development of Energy Water and Environment Systems–JSDEWES 5 (2): 127–138. Qiu, R., Gao, S., Lopez, P.A. et al. (2017). Effects of pH on cell growth, lipid production and CO2 addition of microalgae Chlorella sorokiniana. Algal Research–Biomass Biofuels and Bioproducts 28: 192–199. Cao, L.P., Zhou, T., Li, Z.H. et al. (2018). Effect of combining adsorption–stripping treatment with acidification on the growth of Chlorella vulgaris and nutrient removal from swine wastewater. Bioresource Technology 263: 10–16. Cai, T., Park, S.Y., and Li, Y.B. (2013). Nutrient recovery from wastewater streams by microalgae: status and prospects. Renewable & Sustainable Energy Reviews 19: 360–369. Singh, A.K. and Pandey, A.K. (2018). Microalgae: an ecofriendly tool for the treatment of industrial wastewaters and biofuel production. In: Recent Advances in Environmental Management, 167–195. CRC Press. Kaur, M., Kumar, M., Sachdeva, S. et al. (2018). Aquatic weeds as the next generation feedstock for sustainable bioenergy production. Bioresource Technology 251: 390–402. Cantrell, K.B., Ducey, T., Ro, K.S. et al. (2008). Livestock waste–to–bioenergy generation opportunities. Bioresource Technology 99 (17): 7941–7953. Verma, R. and Suthar, S. (2015). Utility of duckweeds as source of biomass energy: a review. BioEnergy Research 8: 1589–1597. Brennan, L. and Owende, P. (2010). Biofuels from microalgae – a review of technologies for production, processing, and extractions of biofuels and co–products. Renewable & Sustainable Energy Reviews 14: 557–577. Hena, S., Fatimah, S., and Tabassum, S. (2015). Cultivation of algae consortium in a dairy farm wastewater for biodiesel production. Water Resources and Industry 10: 1–14. Chisti, Y. (2007). Biodiesel from microalgae. Biotechnology Advances 25 (3): 294–306. Xu, J.L. and Deshusses, M.A. (2015). Fermentation of swine wastewater–derived duckweed for biohydrogen production. International Journal of Hydrogen Energy 40 (22): 7028–7036. Cheng, C.L., Lo, Y.C., Lee, K.S. et al. (2011). Biohydrogen production from lignocellulosic feedstock. Bioresource Technology 102 (18): 8514–8523. Ferreira, A., Marques, P., Ribeiro, B. et al. (2018). Combining biotechnology with circular bioeconomy: from poultry, swine, cattle, brewery, dairy and urban wastewaters to biohydrogen. Environmental Research 164: 32–38. Montingelli, M., Tedesco, S., and Olabi, A. (2015). Biogas production from algal biomass: a review. Renewable and Sustainable Energy Reviews 43: 961–972. Perazzoli, S., Bruchez, B.M., Michelon, W. et al. (2016). Optimizing biomethane production from anaerobic degradation of Scenedesmus spp. biomass harvested from algae–based swine digestate treatment. International Biodeterioration & Biodegradation 109: 23–28.

597

598

20 Approaches Toward Resource Recovery from Breeding Wastewater

162 Singh, M., Reynolds, D.L., and Das, K.C. (2011). Microalgal system for treatment of effluent from poultry litter anaerobic digestion. Bioresource Technology 102 (23): 10841–10848. 163 Srinuanpan, S., Cheirsilp, B., and Prasertsan, P. (2018). Effective biogas upgrading and production of biodiesel feedstocks by strategic cultivation of oleaginous microalgae. Energy 148: 766–774. 164 Stevens, C. (2011). Thermochemical Processing of Biomass: Conversion into Fuels, Chemicals and Power. Chichester: Wiley. 165 Raheem, A., Hassan, M.Y., and Shakoor, R. (2016). Bioenergy from anaerobic digestion in Pakistan: potential, development and prospects. Renewable & Sustainable Energy Reviews 59: 264–275. 166 Akhtar, J. and Amin, N.S. (2012). A review on operating parameters for optimum liquid oil yield in biomass pyrolysis. Renewable & Sustainable Energy Reviews 16: 5101–5109. 167 Suali, E. and Sarbatly, R. (2012). Conversion of microalgae to biofuel. Renewable and Sustainable Energy Reviews 16 (6): 4316–4342. 168 Goyal, H.B., Seal, D., and Saxena, R.C. (2008). Bio–fuels from thermochemical conversion of renewable resources: a review. Renewable & Sustainable Energy Reviews 12: 504–517. 169 Rather, M.A., Khan, N.S., and Gupta, R. (2017). Hydrothermal carbonization of macrophyte Potamogeton lucens for solid biofuel production production of solid biofuel from macrophyte Potamogeton lucens. Engineering Science and Technology–an International Journal–JESTECH 20 (1): 168–174. 170 Tekin, K., Karagöz, S., and Bekta¸s, S. (2014). A review of hydrothermal biomass processing. Renewable and Sustainable Energy Reviews 40: 673–687. 171 Orfield, N.D., Fang, A.J., Valdez, P.J. et al. (2014). Life cycle design of an Algal Biorefinery Featuring Hydrothermal Liquefaction: effect of reaction conditions and an alternative pathway including microbial regrowth. ACS Sustainable Chemistry & Engineering 2 (4): 867–874. 172 Yokoyama, S. and Matsumura, Y. (2008). The Asian biomass handbook: a guide for biomass production and utilization. The Japan Institute of Energy, pp.: 61–62. 173 McKendry, P. (2002). Energy production from biomass (part 3): gasification technologies. Bioresource Technology 83 (1): 55–63. 174 Chen, W.H., Lin, B.J., Huang, M.Y. et al. (2015). Thermochemical conversion of microalgal biomass into biofuels: a review. Bioresource Technology 184: 314–327. 175 Amin, S. (2009). Review on biofuel oil and gas production processes from microalgae. Energy Conversion and Management 50 (7): 1834–1840. 176 Ro, K.S., Cantrell, K., Elliott, D. et al. (2007). Catalytic wet gasification of municipal and animal wastes. Industrial & Engineering Chemistry Research 46: 8839–8845. 177 Achinas, S., Achinas, V., and Euverink, G.J.W. (2017). A technological overview of biogas production from biowaste. Engineering 3 (3): 299–307.

References

178 Raboni, M. and Urbini, G. (2014). Production and use of biogas in Europe: a survey of current status and perspectives. Revista Ambiente & Água 9 (2): 191–202. 179 Lu, J.B., Zhu, L., Hu, G.L. et al. (2010). Integrating animal manure–based bioenergy production with invasive species control: a case study at Tongren pig farm in China. Biomass & Bioenergy 34: 821–827. 180 Bensah, E.C. and Brew–Hammond, A. (2010). Biogas technology dissemination in Ghana: history, current status, future prospects, and policy significance. International Journal of Energy and Environment 1 (2): 277–294. 181 Logan, B.E. (2004). Extracting hydrogen electricity from renewable resources. Environmental Science & Technology 38 (9): 160A–167A. 182 Bharathiraja, B., Sudharsanaa, T., Bharghavi, A. et al. (2016). Biohydrogen and biogas – an overview on feedstocks and enhancement process. Fuel 185: 810–828. 183 Baldi, M., Collivignarelli, M.C., Abba, A., and Benigna, I. (2018). The valorization of ammonia in manure Digestate by means of alternative stripping reactors. Sustainability 10 (9): 3073. 184 Laridi, R., Auclair, J.C., and Benmoussa, H. (2005). Laboratory and pilot–scale phosphate and ammonium removal by controlled struvite precipitation following coagulation and flocculation of swine wastewater. Environmental Technology 26: 525–536. 185 Suzuki, K., Tanaka, Y., Kuroda, K. et al. (2007). Removal and recovery of phosphorous from swine wastewater by demonstration crystallization reactor and struvite accumulation device. Bioresource Technology 98: 1573–1578. 186 Adam, M.R., Othman, M.H.D., Abu, S.R. et al. (2019). Current trends and future prospects of ammonia removal in wastewater: a comprehensive review on adsorptive membrane development. Separation and Purification Technology 213: 114–132. 187 Hjorth, M., Christensen, K.V., Christensen, M.L. et al. (2010). Solid–liquid separation of animal slurry in theory and practice. A review. Agronomy for Sustainable Development 30 (1): 153–180. 188 Wu, Z.Y., Zou, S.Q., Zhang, B. et al. (2018). Forward osmosis promoted in–situ formation of struvite with simultaneous water recovery from digested swine wastewater. Chemical Engineering Journal 342: 274–280. 189 Shi, L., Hu, Y.S., Xie, S.H. et al. (2018). Recovery of nutrients and volatile fatty acids from pig manure hydrolysate using two–stage bipolar membrane electrodialysis. Chemical Engineering Journal 334: 134–142.

599

601

21 Resources Recovery and Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production Dongyun Du and Jia Li School of Resource and Environmental Engineering, South-Central University for Nationalities, Wuhan, Hubei Province, P.R. China

21.1 Introduction Manganese is considered as an important alloy element and has been widely used in stainless steel, nonferrous industry, electronics, batteries, catalyst, coating, forage additive, and many other fields. Currently, the stainless steel industry accounts for about 90% of total manganese production. Metallic manganese with 99.7–99.9% purity is mainly produced by electrolysis [1]. China accounts for 98 and 90% of the world’s overall EMM production capacity and output and has become the world’s largest country for electrolytic manganese metal (EMM) production, consumption, and export [2]. However, 93.6% of overall manganese ores are of low grade and are treated as relatively scarce resources in China [3]. Therefore, China’s EMM industry highly depends on importation. With the increasing development of EMM industry, a large amount of CO2 , manganese-bearing wastewater, and residue is produced [4]. Improper waste disposal has caused serious environmental pollutions. Hence, reducing CO2 discharge and reutilizing manganese wastes have become key factors for the sustainable development of EMM industry. This chapter focuses on the recovery and reutilization of valuable elements in the liquid and solid wastes from manganese production industry. Section 21.2 introduces EMM production processes and generation of associated wastes. Section 21.3 describes a novel method for manganese recovery from manganese wastewater and for capture of CO2 via laboratory- and pilot-scale tests. Section 21.4 evaluates both chemical and bioleaching methods for the silicon activation in terms of different operating parameters. Section 21.5 summarizes the current research and provides insight for future extended work and directions.

Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

602

21 Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production

21.2 EMM Production Process and Associated Wastes 21.2.1 EMM Production Process Figure 21.1 indicates the production process of EMM. Rhodochrosite and manganese oxide ore are two main raw materials for EMM industry. Rhodochrosite mostly consists of manganese carbonate and is used as electrolyte through acid leaching, neutralization, sedimentation, and pressure filtration. Lastly, SeO2 and (NH4 )2 SO3 are added for electrolyzation inside the electrolytic cell. Manganese oxide ore mainly consists of manganese dioxide and Mn(IV) needs to be reduced to Mn(II) to react with sulfuric acid to form MnSO4 for electrolyzation [5]. There are two reduction approaches: (i) roasting manganese ore with added coal powder to produce CO for Mn(IV) reduction under 800 ∘ C; and (ii) using the two-ore method to leach manganese oxide ore powder and pyrite powder with sulfuric acid at the same time; Mn(IV) is reduced to Mn(II) for the production of manganese sulfate solution. Figure 21.1 indicates the production process.

21.2.2 EMM Wastewater In the production process of electrolytic manganese, the following types of manganese-containing wastewater are generated: (i) ore processing wastewater, including mine pond water, dumping site leachate, ore-dressing wastewater, and tailing leachate (as shown in Figure 21.2); (ii) electrolytic manganese production

CO2

Crushed

H2SO4

Rhodochrosite

Ammonia Pressure filtertion

MnO2

Anodic liquid

Oxidation, Neutralize EMR field Vulcanizing agent Purify +

Manganese

Electrolysis

–

Pressure filtertion

Passivation, clean, peeled off

EMR EMW

Sewage treatment system

Wastewater Containing Pb, Ni

Figure 21.1 Production process of electrolytic manganese, generation of electrolytic manganese residue (EMR) and electrolytic manganese wastewater (EMW).

21.2 EMM Production Process and Associated Wastes

Mine pond water

Drilling

Ore-dressing wastewater

Manganese concentrate

Blasting

Waste disposal site Leachate

Tailings reservoir

Ore grinding

Leachate

Fine crushing

Transporting

Oredressing dill

Stope

Dirt haul

Shovel loading

Ore dressing

Medium crushing

Primary rushing Ore transportation

Figure 21.2 The schematic diagram of manganese ore mining and mineral processing wastewater.

wastewater (as shown in Figures 21.1 and 21.3); and (iii) manganese residue leachate (as shown in Figures 21.1 and 21.3). Ore mining process is mainly in an open environment. The wastewater generation process is complex. The wastewater contains a large amount of suspended solids and free manganese, and the quantity and quality of the wastewater are unstable. As shown in Figure 21.2, mine pond water is water accumulated in the mine pits or ditches during surface or underground mining process; ore-dressing wastewater is water produced during gravitational separation, magnetic separation, or floatation process; leachate from the dumping site and the tailing pond is wastewater generated by flushing or rain water runoff. Figure 21.3 shows the scenes for tailing leachate and ore-dressing wastewater. EMM wastewater is mainly divided into two parts: production water and wastewater from electrolysis process and thereafter. As shown in Figure 21.3d, production water includes wastewater from floor flushing and filter cloth cleaning (i.e. generated from the liquid production section) and other wastewater (i.e. recycled cooling water, electrode plate passivation and plate washing wastewater, generated from electrolysis and post-sequence section). All wastewater is usually recycled after simple treatment. After recycling for 7–10 times, all wastewater is collected and sent to the wastewater treatment facility. EMM wastewater contains numerous substances with high concentration of Mn(II), Cr(VI) NH3 –N, as well as high content of Ca(II) and Mg(II). The concentration ranges of different components of EMM wastewater, which are way higher than the Integrated Wastewater Discharge Standard, are listed as follows [6]: Ca2+ 100–600 mg/L, Mg2+ 100–1500 mg/L, Mn2+ 100–2000 mg/L, Cr(VI) 60–1000 mg/L, and NH3 –H 50–6000 mg/L. Stockpiling (Figure 21.3c) is still the most popular technology for electrolytic manganese residue (EMR) disposal. When surface runoff or groundwater gets

603

604

21 Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production

(a) Rhodochrosite power

(b)

(c)

Liquid production section

Eloctrolysis and post-sequence section

Filter cloth cleaning wastewater

Workshop cleaning wastewater

(d)

Metal manganese sheet

Circulating cooling water

Passivation wastewater

Plate washing wastewater

Figure 21.3 The field pictures or schematic diagram of (a) manganese ore mining, (b) mineral processing wastewater, (c) stockpile of electrolytic manganese residue, and (d) generation of electrolytic manganese wastewater.

into contact with EMR, leaching process will produce wastewater containing high concentration of manganese. EMR leachate is usually collected and treated at the wastewater treatment facility with other wastes.

21.2.3 Electrolytic Manganese Residue (EMR) EMR is the main solid waste produced from the EMM industry. It contains a high content of Ca, Si, and a trace amount of heavy metals. About 6–10 tons of EMR are produced when 1 ton of manganese metal is generated. Statistically, about 7.0 × 106 tons of EMR is discharged every year with 5.0 × 107 tons of disposed EMR. EMR leachate spill has been found with high manganese concentration after water flushing due to the increasing soluble manganese. Manganese residue comprises mainly manganese (up to 4%, mass fraction), nitrogen (up to 1.7%), and a variety of heavy metals. Up till now, little attention has been paid to their utilization [7–9]. Currently, EMR is used for road-base construction, railroad ballast, asphalt pavements, cement and concrete industry [10]. Efficient recovery and resource utilization of EMR is essential for the sustainable development of the EMM industry. The main components of the EMR are SiO2 (23.4%), CaO (15.0%), Fe2 O3 (8.6%), and Mn (4.8%); the content of soluble manganese Mn2+ is about 4%. After backwashing, the remaining water and slag may be used to produce manganese carbonate products, selenium-enriched fertilizers, and

21.3 Manganese Recovery from Manganese-Bearing Wastewater

high-performance environmental materials [10]. Silicon is recognized as the fourth nutrient element after nitrogen, phosphorus, and potassium by the international soil community. Although silicon is a major component of soil (accounting for about 70% of the total soil mass), the effective silicon content accounts for only 0.03%. This is because 99% of the silicon in the soil is present in the silicate minerals, which have very low water solubility, are inactive, and cannot be directly absorbed and utilized by plants. There is an increasing demand for effective silicon in soil due to the fact the over 50% of the farmland in China lack in it. Slow weathering by soil silicon and supplement of effective silicon from straw reutilization are unable to balance the need with demand. Manganese slag is an important source of silicon and contains some manganese sulfate and ammonium sulfate. Therefore, it is imperative to look for ways to activate the silica minerals and release effective silicon to produce high-efficiency silicon-containing compound fertilizer. Reutilization of EMR is a practical way to lighten environmental problems caused by solid wastes. The development of technologies that use manganese slag to prepare high-value products is the future direction to handle and dispose of solid waste, which is also essential for the sustainable development of the electrolytic manganese industry.

21.3 Manganese Recovery from Manganese-Bearing Wastewater 21.3.1 Wastewater Treatment Strategy Manganese (Mn) as an important metal has been widely used in various fields such as nonferrous metallurgy, steel production, catalysis, and batteries [11]. Manganese production typically involves acid leaching and electrolysis. At the same time, large amount of wastewater is produced, which contains Ca2+ , Mg2+ , and Mn2+ ions [12, 13]. The produced wastewater is toxic and needs to be treated due to the high content of manganese or the environment could get polluted and the sustainability of the industry will be affected [3]. Moreover, carbon dioxide (CO2 ) is discharged due to the acid leaching and fossil fuels combustion. About 1 metric ton of CO2 is discharged for 1 ton of manganese produced. Consequently, about 1.1 million tons of CO2 are being produced annually worldwide from manganese production, which accounts for 0.002% of total global CO2 emission [4]. CO2 is well known as the greenhouse gas (GHG), plays a significant role in climate change. Therefore, recycling and reutilizing CO2 generated from electrolytic manganese production could significantly release the burden of carbon footprint of the manganese industry. Traditional treatment technologies for manganese-bearing wastewater are chemical deposition, coagulation, sedimentation, ferrite co-precipitation, ion-exchange, electrolysis, and biological methods. However, the technologies mentioned above have drawbacks include high costs, long treatment time, low efficiency, and secondary pollutions [14–16]. Wang et al. [14] investigated a novel method for recovery of dissolved manganese from manganese residue by using a combination of ammonia and CO2 and the efficiency reached up to 94%. Silva et al. [15] applied limestone

605

21 Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production

and sodium carbonate to treat manganese-bearing wastewater and 99% of manganese removal has been achieved. Hoyland et al. [16] have been able to successfully remove more than 98% of Mn by biofiltration. However, all these treatment options encountered high costs and industry application difficulties. Thus, a selective separation and recovery of Mn using CO2 generated during the acid leaching process will be introduced in detail, which overcomes the drawbacks mentioned above. The following section focuses on introducing pilot-scale tests for developing reliable techniques and industrial equipment for the production of quality industrial-grade manganese carbonate to reduce pollution from manganese-related industry and CO2 emission. The process developed aims at concurrent recovery of Mn from wastewater generated at electrolytic manganese industries, along with the reduction of CO2 emission.

21.3.2 Onsite CO2 Emission An electrolytic manganese plant in Guangxi, China was chosen as a feasibility study on comprehensive utilization of CO2 was first conducted by analyzing the acid leaching process and the emission concentrations of CO2 at an industrial scale. It has been found that the quantities and emission patterns of CO2 are closely related to the operation sequences of acid leaching. Briefly, there are nine acid-leaching slots in the plant. The slots can effectively and continuously control the acid leaching of rhodochrosite. The exhaust fumes from these nine slots are sent to a purification tower before discharge. Figure 21.4 shows the CO2 concentration profile from a single slot. The detention times of the feeding stage, the acid addition stage, and the leaching operation stage were 1.5, 1.0, and, 3.0 hours, respectively. CO2 emissions firstly started from the 50 40 Carbon dioxide (% vol)

606

30

20 10 0 feeding acid addition 0

60

120

180

acid leaching 240 300 Time (min)

360

420

480

Figure 21.4 CO2 concentration profile in one acid-leaching reactor (slot). Source: Yu et al. [17]. © 2019, Elsevier.

21.3 Manganese Recovery from Manganese-Bearing Wastewater

Reactor NO.

Gas emission duration (h)

Day 1

Day 2

Day 3

Figure 21.5 The Gantt chart of the gas production process in three days. Source: Yu et al. [17]. © 2019, Elsevier.

feeding stage, and the CO2 concentration reached a maximum of 42% (v/v) shortly after the completion of the acid addition. It then started to decrease to an insignificant level after about 2.5 hours. For each batch slot, the average total CO2 emission duration was around 5 hours. CO2 emission duration from the three stages (i.e. feeding, acid addition, and acid leaching) was monitored for eight days. Figure 21.5 illustrates the CO2 emission patterns for all nine reactors during the first three days. The average gas emission duration is 4.99 hours for each slot but it varies between 3.50 and 6.76 hours for each one of them. The combined emissions from these nine reactors followed similar pattern and were continuous throughout the monitoring period of eight days. The results showed that CO2 emissions from electrolytic manganese plant can provide a relatively consistent CO2 source for the removal of manganese in wastewater.

21.3.3 Effect of CO2 Dosage Experiments were conducted by adding CO2 to the Mn-bearing wastewater in a reactor to recover the soluble Mn in the solution. Experiments were conducted by adding CO2 to the Mn-bearing wastewater in a reactor to recover the soluble Mn in the solution. The Mn-bearing wastewater includes the wastewater produced in the liquid production section and electrolysis and post-sequence section, which are illustrated in Figure 21.3. Firstly, CO2 was pumped into the reactor containing 1 L of Mn-bearing wastewater at the flow rate of 0.4 L/min at the set reaction temperature and stirring speed. As the pH of the wastewater decreased nearly to 5, 20 mL of 2.5 mol/L NaOH solution was slowly added to the wastewater at the speed of 0.02 mL/s by automatic potentiometric titrator to keep the pH of solution in the range of 6.2–7.0. When all of NaOH solution was added, the pumping of CO2 to the wastewater was stopped and the stirring of wastewater was continued for a certain

607

21 Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production

100

Figure 21.6 Metal recovery efficiency under different values of pH. Source: Yu et al. [17]. © 2019, Elsevier.

Ca Mg Mn

80 recovery (%)

60 40 20 0 6.2

6.4

6.6 pH

6.8

7.0

(300)

(214)

(122)

(024) (018) (116)

(202)

(110)

(113)

(014)

time before filtration. The concentrations of Ca, Mg, and Mn in the filtrate were analyzed by ICP. The filter residue was dried in vacuum oven at 75 ∘ C for 12 hours and then analyzed by XRD and SEM. Figure 21.6 shows that manganese recovery (99%) was the highest compared with calcium and magnesium recovery (54 and 12%, respectively) when pH ≥ 6.6. Thus, pH of 6.6 was then chosen as the control pH during pilot tests as the treated wastewater with its end pH close to the natural value can be directly discharged. Additionally, X-ray powder diffraction (XRD) revealed that precipitates formed mainly composed of rhodochrosite (Figure 21.7). Figure 21.8 shows the concentration of carbonate species at different pHs and pCO2 = 0.15 atm.

(012)

608

45 °C

JCPDS No. 44-1472 20

30

40

50

60

MnCO3

70

2 θ (°)

Figure 21.7 XRD patterns of sediment from wastewater containing multiple ions. Source: Yu et al. [17]. © 2019, Elsevier.

21.3 Manganese Recovery from Manganese-Bearing Wastewater

0 cT,CO

3

–2 H2CO3* 3 2–

– O3

HC

–4

CO

Log(Conc.)

Figure 21.8 Effect of pH on concentrations of carbonate species in water at 25 ∘ C and pCO2 = 0.15 atm. Dashed lines represent the total carbonate concentration, CT,CO3 . Source: Yu et al. [17]. © 2019, Elsevier.

H+

OH

–6

–8 4

6

8

10

12

Equilibrium pH

Carbonate precipitates will form when pH is adjusted to a critical value. The precipitates generally formed through the combination of free carbonic acid CO3 2− and the free active species Me2+ (Eq. (21.1)). Me2+ + CO2− 3 −−−−→ MeCO3 (s)

(21.1)

According to the solubility product of each species, MnCO3 has a smaller solubility product (K sp = 1.8 ⋅ 10−11 ) than CaCO3 (K sp = 2.8 ⋅ 10−9 ) and MgCO3 (K sp = 3.5 ⋅ 10−8 ). Therefore, MnCO3 is preferred to be precipitated. The precipitation pH of CaCO3 , MgCO3 , and MnCO3 is 7.7, 8.2, and 6.6, respectively; therefore, at pH 6.6, selective separation and recovery of manganese can be achieved completely by precipitation. Moreover, during control experiment, manganese can be also recovered at alkaline condition with a pH over 9.5 without the presence of CO2 . Thus, the critical advantage of using CO2 is that it reduces the amount of alkalinity needed.

21.3.4 Pilot Treatment System and Its Performance A pilot treatment system was designed in the plant (Figure 21.9). The pilot system consisted of a flow-regulation tank (L × W × H : 1.34 m × 1.70 m × 2.40 m), a mineralization reactor, where is Mn precipitated out at the presence of CO2 from the wastewater (D × H : 1.20 m × 5.12 m), a sedimentation tank (L × W × H : 1.34 m × 1.70 m × 2.40 m), a recirculation tank (L × W × H : 1.00 m × 1.00 m × 3.10 m), a filter press, and an automatic control system. In order to optimize industrial design, four operational modes were experimentally tested as follows: Batch mode: The batch mode was designed to test if MnCO3 can be steadily produced. The mode includes the one-time feed of Mn-bearing wastewater into the mineralization reaction with closed effluent control valve. Then, CO2 aeration, pH adjustment, and precipitation took place inside the reactor.

609

610

21 Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production Gas to atmosphere

Mineralization reactor

Regulation tank Recireclation tank

Effluent

Backwashing leachate of the EMR

Filter press

Reagent pool MnCO3

Sedimentation tank

Automatic monitoring and control system

CO2

Figure 21.9 A pilot system for acid leaching tail gas CO2 capture and recovery of MnCO3 from Mn-bearing wastewater. Source: Yu et al. [17]. © 2019, Elsevier.

Batch mode with external recirculation: Unlike the batch mode, the closed valve was opened during the batch mode with external recirculation. Also, the internal mixer was turned on. Besides, the mineralization reactor and regulating tank were connected to study the influence of the selective separation and manganese recovery under external circulation conditions. Continuous mode: The continuous mode has a continuous flow-through system, which can feed manganese wastewater into the mineralization reactor. Then the effluent is transferred into the sedimentation tank after CO2 aeration and pH adjustment. During this process, MnCO3 will settle and the final MnCO3 product can be obtained after filtration. The main purpose of this mode is to test the feasibility of Mn recovery. Continuous mode with external recirculation: Similar to the continuous mode, this mode was used to further purify the MnCO3 products and reduce the alkalinity of the system. The pH was maintained at 6.6 according to the lab results by an automatic control system. Alkali was added and Ca2+ , Mg2+ , and Mn2+ in the supernatant were measured after injection. When the reactions terminated, the residuals were vacuum filtrated and dried. Then, the dried samples were collected for analysis by scanning electron microscope (SEM), XRD, and the thermal gravity–differential thermal analysis (TG–DTA) (the volume of wastewater was 3 m3 in all four modes). According to Figure 21.10, the recovery efficiencies of Ca2+ , Mg2+ , and Mn2+ from the wastewater for all four operational modes are increased with increasing reaction time. For batch mode, the recovery efficiency for Mn2+ reached 99.92 and 99.99% after 360 minutes with final concentrations of Mn2+ of 4 and 1 mg/L, respectively without and with recirculation. The results showed that recirculation process could only slightly improve the performance of metals removal. For continuous mode, the Mn2+ recovery efficiencies reached 99.71 and 99.93% with the final concentrations of Mn2+ of 16 and 4 mg/L, respectively, without and with recirculation. When compared with continuous mode only, the recirculation process improved the system performance with better mass transfer and lower effluent concentration of Mn2+ (to

21.3 Manganese Recovery from Manganese-Bearing Wastewater

Ca Mg Mn Recovery rate (%)

Recovery rate (%)

Ca Mg Mn

(a)

Time (min)

(b)

Ca Mg Mn

Ca Mg Mn Recovery rate (%)

Recovery rate (%) (c)

Time (min)

Time (min)

(d)

Time (min)

Figure 21.10 Ion recovery efficiencies: (a) batch mode, (b) batch mode with recirculation, (c) continuous mode, (d) continuous mode with recirculation. Source: Yu et al. [17]. © 2019, Elsevier.

4 mg/L). The total volume of the treated wastewater is 3 m3 for batch modes and the total volume of the treated wastewater is 9 m3 for continuous modes. For all four operation modes, the much higher recovery efficiency of Mn2+ indicates Mn2+ was selectively removed. For batch modes, 123–130 L of alkali was used and added under a consumption rate of 38–41 L alkaline solution /m−3 wastewater to maintain the pH of the system. For continuous modes, 500–540 L of alkali was used and added under a consumption rate of 45 L/m3 alkaline solution/m−3 wastewater to maintain the pH of the system. The batch mode consumes much more alkali than the continuous modes by the automatic control system.

21.3.5 Characteristics of Precipitates Formed Figure 21.11 shows the XRD patterns of the precipitates formed from four different operation modes. The results indicate that the precipitates were mainly solid products, which were mainly rhodochrosite (MnCO3 ). SEM results (Figure 21.12) show that all four precipitates contain hexagonal crystals with compact structure and larger granule, which are the obvious characteristics of manganese carbonate lattice. From the EDS analysis spectra (Figure 21.13), C, O, Mn, and Ca were found, and the high peaks of Mn and O further verified the product to be MnCO3 . The small

611

21 Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production

(a) 10000 (b) Intensity (a.u.)

612

(c)

5000

(d) 0

20

30

40

50

60

70

2 θ (°)

Figure 21.11

XRD patterns of the precipitates. Source: Yu et al. [17]. © 2019, Elsevier.

(a)

(b)

(c)

(d)

Figure 21.12 SEM images of solid products. (a) Batch mode, (b) batch circulating mode, (c) continuous mode, (d) continuous circulating mode. Source: Yu et al. [17]. © 2019, Elsevier.

21.3 Manganese Recovery from Manganese-Bearing Wastewater

Figure 21.13 Elsevier.

SEM–EDS images of sample from batch mode. Source: Yu et al. [17]. © 2019,

amount of calcium found could be due to the formation of CaCO3 since the Ksp of CaCO3 is very close to the Ksp of MnCO3 . The precipitates yielded (in kg/ton water) were 11.18, 11.43, 11.32, and 11.34 for batch mode, batch mode with recirculation, continuous mode, and continuous mode with recirculation, accordingly.

21.3.6 Thermal Stability of Formed MnCO3 Thermal gravity–differential thermal analysis (TG–DTA) of the precipitates was performed under nitrogen gas at a heating rate of 10.0 K/min. Results indicate that the loss of weight at 220 ∘ C is caused by the loss of carbonate minerals associated water. Therefore, the analysis was focused on the weight loss below the temperature of 220 ∘ C. The TG and the DTA results of the four precipitates show that loss of weight due to thermal decomposition at 220−650 ∘ C of batch, batch with recirculation, continuous, and continuous mode with recirculation was 37.8, 39.5, 33.2, and 37.1%, respectively. For continuous modes, the TG and DTA curves have small weight loss intervals which could be possibly caused by the substitution of Mn by Ca, Mg, and Fe. The content of manganese carbonate was measured by the ammonium oxidation/reduction method, the content of chloride was measured by the silver nitrate titration method, and the content of sulfate was measured by the barium chloride sediment titration method. The quality of CO2 which was selectively separated from manganese carbonate in electrolytic manganese production from wastewater was also analyzed to explore the economic benefits. Table 21.1 shows the test results and the technical indicators of industrial carbonate [18]. The table illustrates that the manganese carbonate (in Mn % by weight) content of the precipitate generated from the continuous mode does not reach the industrial grade II. While the grade of other three precipitates generated are higher than the industrial grade II. The batch-with-recirculation mode offers the best quality of manganese carbonate in all six samples.

613

614

21 Reuse from Liquid and Solid Wastes Generated from Electrolytic Manganese Production

Table 21.1 weight).

Components of the manganese carbonate (in % by

Mn (≥)

Cl− (≤)

SO4 2− (≤)

Industrial grade I

42.5

0.01

0.6

Industrial grade II

41.5

0.02

0.8

Batch mode

41.9

0.01

0.6

Batch mode with recirculation

42.9

0.01

0.6

Continuous mode

41.2

0.01

0.6

Continuous mode with recirculation

41.6

0.01

0.6

Source: Yu et al. [17]. © 2019, Elsevier.

21.3.7 Potential Application in Industry The technique of using captured CO2 from flue gases to selectively recover Mn from the wastewater is discussed in detail in this chapter. This could be potentially applied in the treatment of manganese-containing wastewater from landfill leachate of EMR, backwashing leachate of EMR, mining, electrolytic manganese wastewater, etc. In China, about 1.1 million tons of EEM is produced annually and more than 350 tons of manganese-bearing wastewater generated. In the whole world, approximately 3.8 billion tons of manganese-bearing wastewater are being produced annually with an average manganese concentration of ∼1000 mg/L [13, 18, 19]. Given the similar average concentrations of Mn in global manganese-bearing wastewater with the concentration of soluble Mn in this study, around 0.15 million tons of soluble manganese can be gained from the leachate and manganese-bearing wastewater [20, 21]. Therefore, the harmlessness of EMR has large potentials for recovering soluble manganese. This work provides a new approach for sludge utilization as a valuable resource [14, 22, 23]. Additionally, at least 428 000 tons of CO2 is estimated to be captured by the soluble manganese from manganese-bearing wastewater and EMR each year. According to economic analysis, with an initial Mn concentration of 5 g/L, a total operating cost of $4 ton−1 was estimated, which includes alkali consumption, power cost, and labor cost. However, accounting for the production of 10.35 kg of industrial grade manganese carbonate, a total of $14 ton−1 gain was estimated for the process, indicating that applying this technology for treating manganese-bearing wastewater could be cost-effective. Therefore, this technology may be feasible, cost-effective, and valuable for both controlling wastewater pollution and reducing the emission of CO2 .

21.4 Activation/Recovery of Silicon from EMR: Two Methods

21.4 Activation/Recovery of Silicon from EMR: Two Methods 21.4.1 Characterization of EMR The dried EMR was crushed and then passed through a 100-mesh sieve (particle size 50 kPa are usually fitted to particles with a density of >2000 kg/m3 and a diameter of Pt and the maximum leaching rate was achieved at pH of 10 for all metals. However, due to

737

738

25 Resource Recovery from Electronic Waste

the toxicity and adverse contamination problems of cyanide, more research has been focused on the development of less toxic non-cyanide reagents and the most commonly used non-cyanide lixiviants are halide, thiourea, and thiosulfate. Halide leaching method can achieve efficient metal extraction and lixiviants used include chloride, fluorine, bromide, iodine, and astatine leaching. Except for fluorine and astatine, all regents have been used for fold extraction by forming both Au(I) and Au(III) complexes, but only chloride has been applied on industrial scale. Furthermore, this method has some drawbacks such as high consumption of reagent and high economics, hindering its application at large scale. Compared to cyanide extraction, chloride leaching of gold is more challenging owing to two major reasons: (i) special stainless steel and rubber-lined equipment has high resistance to highly corrosive acidic and oxidizing conditions, and (ii) chlorine gas is highly volatile and poisonous and thus must be handled carefully to avoid health risks [32, 36]. Thiourea ((NH2 )2 CS) leaching has advantages such as low toxicity, promising recovery rates (up to 99% gold extraction), faster kinetics, and high selectivity. However, the rate of leaching depends on the concentrations of thiourea and oxidants. Thiourea is not stable in alkaline solution and also readily dissolves in acidic solution. Thus, ferric ion in sulfuric acid is used in the reaction to oxidize thiourea reagent and forms a stable ferric sulfate complex [(FeSO4 .CS(NH2 )2 )+ ] through forming an intermediate formamidine disulfide complex. Moreover, it is more expensive than cyanide due to high consumption during gold extraction, and the rate of gold dissolution is strongly determined by pH. As a result, this process needs further development for its commercial applications [31, 37, 38]. As a chemical widely used in photography and in the pharmaceutical industries, thiosulfate (S2 O3 2− ) has been proposed for dissolving gold in presence of cupric ion (Cu2+ ) and ammonia (NH3 ). Cu2+ and NH3 act as a catalyst in thiosulfate solution to improve metal recovery by forming a stable cupric tetra-amine complex, which can stabilize the gold–thiosulfate complex [Au(S2 O3 )2 3− ]. Nevertheless, to optimize the concentration ratio of ammonia to thiosulfate, it is essential to achieve efficient metal recovery, as the increase in thiosulfate concentration can lead to formation of unwanted byproducts (i.e. sulfate, trithionate, and tetrathionate), and reduce recovery efficiency [39–41]. 25.5.1.3 Biohydrometallurgical Processes

Over last two decades, using biometallurgy to recover metals has been proved to be one of very promising technologies, which is environment-friendly and has great potential to lower operating cost and energy consumption. The most commonly used biohydrometallurgical processes are bioleaching and biosorption. With the assistance of special microorganisms, bioleaching has been applied for recovering metals from metallic sulfides, which are the major bearing minerals for many base metals and precious metals. In addition to successfully recovering copper and gold at industrial scale, studies on bioleaching have been focused on the extraction of other metals such as nickel (Ni), cobalt (Co), molybdenum (Mo), zinc (Zn), lead (Pb), and silver (Ag) [42, 43]. As a number of individual and mixed culture of acidophilic chemolithotrophic bacteria and acidophilic heterotrophic bacteria are

25.5 Recovery of the Valuable Materials from E-Waste

able to facilitate metal dissolution and increase the rate of metal extraction through a series of bio-oxidation and bioleaching reactions, they have been widely used for bioleaching applications. For example, thiobacilli group of bacteria (Acidithiobacillus thiooxidans and Sulfobolus sp.), iron-oxidizing bacteria (Leptospirillum sp.), mesophilic bacteria (Acidithiobacillus ferrooxidans), and moderately thermophilic strains (Sulfobacilllus thermosulfidooxidans and Thermoplasma acidophilum) have been extensively investigated for metal extraction [44–46]. Biosorption process is a passive physicochemical adsorption process by utilizing certain types of inactive, dead, microbial biomass or derived product, which can interact with metal ions so as to concentrate heavy metals even from very dilute aqueous solution. Depending on the specific properties of biosorbents, physicochemical metal removal mechanisms include ion-exchange, complexation, coordination, and chelation between metal ions and ligands. Other metal removal mechanisms are related with metabolism, including metal precipitation as sulfides, sequestration by metal-binding proteins, peptides or siderophores, transport and internal compartmentalization [47]. Since biosorption has merits such as high efficiency, low operating cost, and low sludge production, a variety of microbial biomass have been utilized for accumulation and recovery of heavy metals and precious metals, including algae (e.g. Chlorella vulgaris, Fucus vesiculosus, Ascophyllum nodosum, etc.), bacteria (e.g. Zoogloea ramigera (aerobic bacterium), Desulfovibrio desulfuricans (sulfate-reducing bacterium), Cyanobacterium violaceum (cyanogenic bacterium), Streptomyces rimosus, etc.), yeast (e.g. Saccharomyces cerevisiae, Candida albicans, Pichia stipites, etc.), fungi (e.g. Penicillium simplicissimum, Aspergillus niger, Mucor racemosus, Rhizopus arrhizus, Ganoderma lucidum, etc.), proteins (e.g. Hen eggshell membrane, lysozyme, bovine serum albumin [BSA], etc.), and others (e.g. alfalfa, chitosan derivatives, condensed-tannin gel, acid-washed Ucides cordatus, etc.) [32, 48, 49]. To achieve optimum metal recovery, bioleaching and biosorption normally are applied as an integrated approach. Furthermore, appropriate chemical modification methods (i.e. pretreatment, binding site enhancement, binding site modification and polymerization) also can increase and activate binding sites of biomass surface, resulting in enhanced sorption capacity for metals due to a large number of functional groups presenting on the biomass surface [50]. The structure of the biosorbents and the surface area of biomass determine the efficiency of metal biosorption. Table 25.4 summarizes different metals’ affinity for ligands during biosorption process [49]. Class A metal ions prefer to bind the type I ligands through oxygen, while Class B metal ions show high affinity for type III ligands, as well as form strong binding with type II ligands. Borderline metal ions could bind these three types of ligands with different preferences [51]. Current research has been focused on the combination of two or more technologies in order to achieve efficient extraction of metals, as each of the major technology does have merits and issues and the integrated treatment method can provide complementary solution for better metal recovery. Table 25.5 compares the pros and cons of pyrometallurgical, hydrometallurgical, and biohydrometallurgical processes.

739

740

25 Resource Recovery from Electronic Waste

Table 25.4

Different metals’ affinity for ligands during biosorption process.

Type of ligand

Ligands

Metal classes

Type I: Ligands prefer to Class A

F− , O2− , OH− , H2 O, CO3 2− , SO4 2− , ROSO3 − , NO3 − , HPO4 2− , PO4 3− , ROH, RCOO− , C=O, ROR

Class A: Li, Be, Na, Mg, K, Ca, Sc, Rb, Sr, Y, Cs, Ba, La, Fr, Ra, Ac, Al, lanthanides, actinides

Type II: Other important ligands

Cl− , Br− , N3 − , NO2 − , SO3 2− , NH3 , N2 , RNH2 , R2 NH, R3 N, =N-, -CO-N-R, O2 , O2 − , O2 2−

Borderline ions: Ti, V, Cr, Mn, Fe, Co, Ni, Cu, Zn, Ga, Cd, In, Sn, Sb, As

Type III: Ligands prefer to Class B

H− , I− , R− , CN− , CO, S2− , RS− , R2 S, R3 As

Class B: Rh, Pd, Ag, Lr, Pt, Au, Hg, Tl, Pb, Bi

The symbol R represents an alkyl radical such as CH2 − , CH3 CH2 − , etc. Source: Wang and Chen [49]. © Elsevier.

25.5.2 Recovery of Plastics from E-Waste Plastics are man-made synthetic materials out of the natural recycling cycle and the use of plastics in EEE has been increasing both in types and quantities, which can represent a significant portion of e-waste generated (approximately 10–30% of WEEE). Although plastics have a high recycling potential and the recycling of plastics in WEEE is implemented in many countries, the recycling task can be tougher and more complicated compared to recycling of plastics from other sources (e.g. packaging, municipal solid waste, building and construction) because of the following two reasons: Firstly, plastics in WEEE may consist of more than 15 different types of polymers (around 20% of the total weight of EEE), including acrylonitrile–butadiene–styrene (ABS), high-impact polystyrene (HIPS), polypropylene (PP), polyethylene (PE), polystyrene (PS), styrene–acrylonitrile (SAN), polyesters, polyurethane (PU), polyamide (PA), blends of polycarbonate (PC)/ABS and blends of HIPS/poly (p-phenylene oxide) (PPO), etc. Secondly, most of synthetic organic polymers used in electrical and electronic products contain toxic additive, namely brominated flame retardants (BFRs) or polyvinyl chloride (PVC) [52–55]. It is estimated that approximately 2.7 Mt/year of plastics containing flame retardants are discarded in WEEE globally. BFRs are the major environmental and health concern in terms of plastic recycling. During thermal treatment, the formation of polybrominated dibenzo-p-dioxins (PBDDs) and polybrominated dibenzofurans (PBDFs) can occur because the use of BFRs as flame retardant additives in EEE plastics improves fire safety. As one of BFRs, PBDEs can be released through (i) open dumping and stockpiling, (ii) crushing and grinding processes, and (iii) open burning or uncontrolled thermal treatment of printed wiring boards (PWBs) [56, 57]. The current technologies for plastics recycling from WEEE have been insufficient due to their complexity and large quantity. In most of the cases, the waste plastics are recycled into lower grade products. Moreover, in the development of efficient methods for plastic recycling, it is important to obtain sufficient information about

25.5 Recovery of the Valuable Materials from E-Waste

Table 25.5 The comparisons of benefits and problematic issues of pyrometallurgical, hydrometallurgical, and biohydrometallurgical processes. Process

Benefits

Pyrometallurgy

●

●

●

High temperature results in fast reaction Easy separation of valuable metals and waste Converting 95% of feed to useful products

Problematic issues ● ●

●

●

● ● ●

Hydrometallurgy

● ●

●

●

●

●

Biohydrometallurgy

● ●

●

●

●

●

Less energy intensive Easy to operate and control Not releasing hazardous substances Less greenhouse gas emission More efficient to recover precious metals Can be used as subsequent process for pyrometallurgy Energy-efficient process High technical applicability and economic feasibility Reducing chemical usage and sludge production Simple operation and more environment-friendly Able to extract metals at very low concentration High efficiency in detoxifying effluents

●

● ●

●

●

●

●

●

●

●

●

High energy demand Releasing hazardous substances such as dioxins (due to the presence of halogenated flame retardants), furans, and volatile metals Transferring iron and aluminum into slag and aluminum impair slag properties Increasing loss of precious metals and base metals from slag due to ceramic components and glass in e-waste Impossible to recover plastics Partial purification of precious metals Not effective for extracting metals at very low level Producing a large amount of waste acid liquid that needs further management to reduce possible environmental contamination Slow and time-consuming process Requiring fine grinding for efficient leaching process Adding chemicals and reagents may be highly toxic and increase treatment cost Not effective for extracting metals at very low level Cannot recover all of the metals present in e-waste Limited utility for commercial application of recovering metals from a large quantity of e-waste due to slow nature of process compared to pyrometallurgy and hydrometallurgy The efficiency of targeted metal recovery largely depending on the matrix composition and chemical speciation Further development is needed to treat high complexity of electronic waste, as well as achieve high metal uptake capacity Microorganisms are sensitive to pH and temperature change Passivation and other secondary reactions may negatively affect the metal solubilization resulting in poor metal yields and slow leaching rates during bioleaching

741

742

25 Resource Recovery from Electronic Waste

components and potential hazardous substances in WEEE [53]. Generally, plastics recycling from e-waste can be divided into four categories [58]: 1. Primary recycling: It is also called mechanical recycling. Waste plastic is mechanically reprocessed into a product that has properties equivalent to the original product. Mechanical recycling is probably the most commonly used method for plastics recycling, which mainly is composed of size reduction (e.g. granulation, shredding, etc.), sorting and cleaning (e.g. manual, X-ray fluorescence [XRF], optical-infrared, etc.), as well as melt processing (e.g. extrusion, injection molding, etc.). The recycled plastic can be reused in manufacturing. 2. Secondary recycling: Compared with the property of original product, waste plastic is mechanically reprocessed into a lower grade product. 3. Tertiary recycling: This type of recycling mainly focuses on extracting BFRs from e-waste plastics, which can then be transferred into useful products. Chemicals and solvents (e.g. isopropanol, toluene, n-hexane, etc.) are used for extraction and treatment. The rule to select prime solvents is that the more similar the solvent and polymer are in polarity or solubility parameters, the easier it dissolves the polymer in the solvent. Furthermore, when the interaction between the polymer and solvent exceeds the cohesion of the polymers, the polymer is easily dissolved in the solvent [59]. 4. Quaternary recycling: It is a waste-to-energy process using thermal treatment (e.g. pyrolysis, gasification and plasma arc) to deal with waste plastics. Although it can reduce negative impacts on the environment and human health, the production of hazardous compounds (i.e. dioxins, furans and other POPs) is a major obstacle for using plastics from e-waste as a source for energy production. Moreover, the quality of energy produced is strongly governed by the chemical composition of the recycled materials and operating conditions (i.e. temperature, residence time and reactor type). Most technologies are still in research phase and some reported technologies such as KDV technology from Alphakat, Germany and Haloclean technology have not been applied in industrial scale. The major drawback of primary and secondary recycling is the limited time a thermoplastic can go through thermal or melt processing, as the polymer chains of a certain type of plastic can degrade and get shorter and shorter after certain number of melt processing cycles. Furthermore, mechanical recycling has been developed well for recycling single thermoplastics such as high-density PE and low-density PE, whereas it is not efficient to process blended polymeric materials such as blends of PC/ABS, because mixed recycled polymers exhibit weak mechanical properties and show unpredictable rheological properties, which preclude their usage in high-value applications. Thus, regarding mechanical recycling processes, two strategies should be considered in order to turn recycled plastics into suitable feedstock for high-value applications. One is to select proper additives (e.g. impact modifiers), that can be added to recycled resin to improve mechanical properties. The other is to mix appropriate amount of recycled material with the virgin material for getting a blend that exhibits a good balance between properties and processability properties [53, 58, 60]. For preparing feedstock for energy recovery,

25.5 Recovery of the Valuable Materials from E-Waste

despite the utilization of some promising thermal technologies (e.g. co-pyrolysis and catalytic pyrolysis), most of the methods are merely applied for treating simple pure plastics, such as PP, PE, PS, and PVC [54].

25.5.3 Recovery of Lithium-Ion Batteries from E-Waste As an electrochemical device able to convert chemical energy to electrical energy, a battery consists of an anode, a cathode, an electrolyte, separators, and the external case. Electrodes and electrolyte determine the characteristics of battery system and its application. Separators normally are composed of polymeric materials, paper, or paperboard, while external case consists of steel, polymeric materials, or paperboard. Batteries commonly used include two basic types, namely single-use primary cells (non-rechargeable batteries, e.g. zinc–carbon batteries, alkaline– manganese batteries, lithium batteries, etc.) and rechargeable secondary cells (rechargeable batteries, e.g. nickel–cadmium (Ni–Cd) batteries, nickel–metal-hydride (Ni–MH) batteries, lithium-ion (Li-ion) batteries, etc.) [61]. Since being commercialized by SONY in 1990s, lithium-ion batteries (LIBs) have experienced increasing usage for portable electronic devices and become the most leading power source for consumer electronics, especially dominating smartphone and laptop computer markets. Moreover, the large LIBs have not only become more popular in powering hybrid and electric vehicles, but also been used in energy-storage applications for solar power and wind power electric generation systems. Nowadays, lithium and LIBs represent approximately 37% of the global rechargeable battery market and the member of consumer products using LIBs keeps increasing constantly. It is also estimated that plug-in hybrid electric vehicle (PHEV), electric vehicle (EV), and hybrid electric vehicle (HEV) account for 20% vehicle market in 2020. Compared to other types of batteries, LIBs have the advantages of high energy density, high storage efficiency (83%), long life span, low self-discharge, small volume, light weight, non-memory effect, wide range of application temperatures, and advantages in environmentally compatible operations [62–64]. Lithium is a precious and limited resource (0.007% of the earth’s crust), and current commercial lithium production mainly relies on brines (59%) and hard igneous rocks (25%). It is forecasted that global lithium demand for battery industry will grow from 12 160 tons in 2020 to 21 520 tons by 2025, which represent 37 and 66% of the present total lithium production capacity. Thus, the lithium production bottleneck could become a vital issue for the lithium industries in the near future, which can be prevented by 100% LIBs recycling with minimum 90% lithium recovery [62, 65]. Spent LIBs usually contain 5–20% cobalt (Co), 5–10% nickel (Ni), 5–7% lithium (Li), 5–10% other metals (copper [Cu], aluminum [Al], iron [Fe], etc.), 15% organic compounds, and 7% plastic. It has also been stated that 4000 tons of spent LIBs contain 1100 tons of heavy metals and more than 200 tons of toxic electrolyte (containing harmful substances such as flammable organic solvents and fluorine-containing lithium salts). Hence, recycling valuable metals (i.e. lithium, nickel, cobalt, copper, and manganese) from spent LIBs can be both economically profitable and environmentally friendly [64, 66]. To date, the recycling of spent

743

744

25 Resource Recovery from Electronic Waste

LIBs is limited (less than 3%) and there are no existing regulations to guide the recycling of spent LIBs at a large scale [67]. Most of the available technologies only focus on recycling cobalt, nickel, and copper from spent LIBs. Dewulf et al. [68] mentioned the use of recycled cobalt and nickel from spent LIBs to manufacture active cathode material could save 51.3, 45.3, and 57.2% of natural resources, fossil fuels and nuclear energy demand, respectively. Before recycling, spent LIBs must be discharged first in order to avoid short circuiting or spontaneous combustion because of residual energy in spent LIBs. A commonly used method of discharging is to immerse spent LIBs in a salt solution. Based on the assembly of LIBs and the diversity of electrode materials, the present recycling and resource recovery from spent LIBs requires two principal processes: physical processes and chemical processes. Physical processes include pretreatments (e.g. manual dismantling, crushing, screening, magnetic separation, washing, thermal pretreatment, etc.), mechanochemical processes, dissolution processes and ultrasonic-assisted separation, while chemical processes include acid leaching, bioleaching, solvent extraction, chemical precipitation, and electrochemical processes [63, 64]. However, there are no obvious boundaries between the different processes and technologies applied, as the applied recycling processes can vary case by case. Other the other hand, metal-extraction processes play a vital role in the entire recycling process and may involve one or more treatment methods. Generally, the metal recycling from spent LIBs can be classified into three key processing steps, namely, pretreatment processes, metal-extraction processes, and product preparation processes. The purposes of pretreatment are to separate and enrich components and materials that have similar physical properties (e.g. shape, density, conductivity, magnetic property, etc.), as well as improve recovery efficiency and reduce energy consumption of the following pyrometallurgical or hydrometallurgical processes [63]. Although ongoing research has successfully developed some commercialized recycling processes, the biggest challenge is the technology and chemistry of LIBs are ever-evolving, and thus, new recycling technologies need to be developed to satisfy the recycling requirements, especially for lithium recovery from the economic point of view. To develop a promising recycling process that can be applied in industrial scale, the process should be able to recover all valuable components from spent LIBs with low energy consumption and least environmental pollution. Georgi-Maschler et al. have developed a selective pyrometallurgical treatment for recovering cobalt and lithium from spent LIBs simultaneously (Figure 25.4). The process utilizes four physical processes and an electric arc furnace for pyrometallurgical treatment, which can transform the material fractions of spent LIBs into a cobalt alloy and lithium-containing concentrates. After that, lithium is recovered as lithium cobalt oxide (LiCoO2 , 60% Co) by using a subsequent hydrometallurgical step. Additionally, electrolyte and other metals, such as iron-nickel fraction, aluminum, and copper, also can be diverted out for further recovery. However, the economic efficiency of the process strongly depends on the cobalt price and the economic evaluation has indicated that a minimum cobalt price of €20/kg is required for every 1000 tons of LIBs scrap processed per year [69].

25.5 Recovery of the Valuable Materials from E-Waste Li-ion battery scrap

pre-treatment (sorting and exposing of single battery cells)

plastic fraction, electronic fraction, attaching parts, other battery types

battery cells

vacuumthermal treatment (distillation and pyrolysis)

electrolyte condensate

deactivated cells

mechanical processing (crushing and material separation)

Fe-Ni fraction, Al fraction, electrode foil fraction

electrode material (fine fraction < 2.0 mm)

binder, slag components

agglomeration (pelletizing or briquitting) sulphuric acid, sodium carbonate, water

EM pellets

slag components (if necessary)

pyrometallurgical processing (graphite minimization and carbo-reductive melting)

Li concentrates

Co-base alloy

hydrometallurgical processing (leaching and precipitation)

residue, sodium sulphate

Li carbonate

Figure 25.4 Flowchart of recycling process proposed for Li-ion batteries. Source: Georgi-Maschler et al. [69]. © Elsevier.

25.5.4 Recycling Waste Solar PV Panels To abate primary energy consumption and the impact of climate change, solar energy, as one of the low-carbon pillars of renewable technologies, has been rapidly gaining ground to support global energy supply with an annual growth rate of more than 20%. A typical solar-energy system consists of a solar panel using several solar cells in series, a solar controller to automatically prevent overcharge of the battery, and a battery or group of batteries to store the energy. An inverter may be included to convert the direct current into alternating current if the required output power is 220 V (AC) or 110 V. Solar panels, also known as solar cells or photovoltaics (PVs), are the key element of solar systems, converting solar radiation directly into electrical current by use of carefully engineered semiconductor materials [70].

745

746

25 Resource Recovery from Electronic Waste

Solar panels have evolved through three generations: (i) first-generation crystalline silicon (crystalline-Si), sharing 90% of the global PV market (40% monocrystalline, 48% multi-crystalline [also referred to as polycrystalline], and 2% ribbon Si); (ii) second-generation thin-film, about 9% of the global PV market (2% amorphous silicon [amorphous-Si], 5% cadmium telluride [CdTe], 2% copper indium gallium selenide [CIGS]); and (iii) third-generation concentrator photovoltaics (CPVs), dye-sensitized solar panels, organic solar panel, and hybrid panels. Although monocrystalline PVs have higher efficiency than multi-crystalline PVs, they both are currently mostly used commercial solar-panel materials worldwide due to their higher conversion efficiency and guaranteed the highest investment return than thin-film PVs, as well as great recycling value. The average lifetime of installed PV panels is around 25–30 years [71]. Thus, a substantial amount of end-of-life PV waste is expected to be generated after significant volumes of PV installations have been carried out in early 1990s, increasing from 100 000 tons in 2016 to 60–78 Mt in 2050 [72]. However, waste PV panels are classified as general or industrial waste in most countries around the world. Therefore, the environmental issues related to end-of-life solar panels cannot be ignored. PV components as discarded electronic devices have also been added into the European Union’s WEEE Directive officially in July 2012. However, till now, the collection and recycling of waste solar panels have been very low, making up only 10% of total PV waste volume. The main reasons hindering PV panels recycling are PV panels are not economically beneficial, especially for Si-based PVs, and the low profitability of current recycling processes also negatively affect the initiative for collecting and recycling from manufacturers’ and consumers’ point of view. Furthermore, the regulations for proper PV waste management have not been well established in most of the countries around world, except for the European Union, Japan, and the United States. In addition to reducing the amount of e-waste that ends up in landfill and prevent environmental pollution and health hazards, the recycling of waste PVs can also save energy required and production cost during manufacturing new PV panels, because silicon production is an energy-intensive process. The energy and cost needed to recover silicon from waste PVs are only 1/3 of energy for using pure silicon metal in PV panel manufacturing. Moreover, metal contents (i.e. aluminum, copper, silver, tin, lead and cadmium) in waste PV panels can be recovered as well to alleviate the depletion of natural resources. Thus, the current technology development is mainly targeted at recycling of Si-based PVs for silicon extraction and recycling metals, whereas little attention has been paid to other types of PVs [70, 71]. At present, processing and recycling of waste solar panels can be performed using three steps as follows [70]: 1. Component repair: After overhauling the system and analyzing faults in the junction box, components are replaced to prevent electrical failure and this step does not involve separation or material recycling. 2. Module separation: Waste panels are dismantled through manual dismantling and components are separated according to physical and chemical properties of

25.6 Future Perspectives and Conclusion

materials and structure by using methods such as mechanical separation, organic solvent dissolution, chemical etching, heat treatment, cryogenic breaking, and electrostatic separation. 3. Extraction of silicon and metals from components: Silicon can be recycled using cement-based thermal insulation system and chemical method, while metals can be recovered using hydrometallurgical processes after grinding. Thin-film solar modules can be recycled using a combination of mechanical and chemical treatment. If the waste panels contain battery piece, precious metals such as indium, gallium, cadmium, selenium, etc. can be recovered using hydrometallurgy. The Italian company Sasil S.r.l has developed an innovative pilot-scale recycling process called “Full Recovery End of Life Photovoltaics” (FRELP) to recover materials from waste crystalline-Si panels [73]. Firstly, the waste PV modules are transferred into an automated system for dismantling and removing frames, junction box, and cables. Copper and aluminum are further treated for recycling plastics are used for energy recovery. Secondly, glass is treated by using thermomechanical method to separate glass and ethylene vinyl acetate (EVA) foil. The glass scraps go through a refinement process by using optical sorting. Thirdly, after size reduction, the detached EVA foil is incinerated. Ash produced from incineration is sieved, and then treated by acid leaching (HNO3 ) and filtration processes to recover metallurgical grade silicon. Subsequently, silver and copper are also recovered through electrolysis. Afterward, the residues from the electrolysis process are neutralized and filtered for disposal in landfill. Berger et al. [74] also presented an advanced CdTe thin-film modules recycling technique proposed by First Solar, which has been implemented in full-scale application and includes the following recycling stages: (i) glass-laminate material of end-of-life modules are broken mechanically into large pieces using a shredder for size reduction, and further crushed into 4–5 mm pieces using a hammer mill; (ii) semiconductor film is removed in a slowly rotating leach drum using sulfuric acid and hydrogen peroxide; (iii) glass is separated from the liquids by emptying the content of the leach drum into a classifier; (iv) a vibrating screen is used to separate glass from the larger pieces of EVA foil; (v) glass is rinsed to remove any residual semiconductor material remaining on the glass; (vi) metals are precipitated and concentrated in a thickening tank using sodium hydroxide and the resulting filter cake is packaged for metals recovery by a third party; and (vii) the semiconductor materials are purified for processing into products for new solar modules.

25.6 Future Perspectives and Conclusion As a huge problem and an alarming issue for modern society, environmentally sound e-waste management is indeed an imperative task to alleviate negative environmental impacts and prevent public health risks. E-waste is also an important secondary source for recycling and recovery in terms of creating an energy and resource-efficient economy.

747

748

25 Resource Recovery from Electronic Waste

On an international scale, current available global e-waste statistics are not able to track the real amount of WEEE generated as well as the amount of cross boundary transport e-waste to less developed/developing countries, who generally have less e-waste management infrastructure. Moreover, the types of e-waste covered by legislations in different countries can vary considerably, resulting in difficulties in coordinating collected and recycled e-waste amounts. Thus, establishing a strong partnership and collaboration between countries, government, policy makers, industries, and businesses to produce reliable and comparable global, national, and regional e-waste data not only can identify e-waste recycling opportunities as well as related pollutants and health effects, but also can assist in measuring and improving the existing policies and legislations so as to propose best available practices for better global e-waste management. In addition, the international community should consider the adverse health effects of e-waste exposure as a priority issue and set up corresponding measures to monitor the health effects due to e-waste exposure, especially for children and vulnerable populations. The international health community, academia, policy experts, and nongovernmental organizations, together with national governments, should develop policy solutions and educational programmers, as well as make provision for reducing e-waste exposure and its health effects. [1, 75]. On national level, governments, manufacturers, producers, companies, businesses, operators of recycling services and the general public are the key players. Laws and regulations enforced by governments should be compatible with international agreements and programs, improve downstream monitor of e-waste, promote recycling, encourage reuse, and most important be well implemented. From an upstream perspective, manufacturers, producers, companies, and businesses should improve waste recycling practices through Extended Producer Responsibility (EPR) by taking specific responsibility for the development of financial incentives to increase collection rates, as well as the treatment, recycling, and disposal of e-waste. They can promote eco-friendly design and better design-for-recycling products, which are easy to handle, dismantle, separate, and recycle. For countries or regions having intensive informal e-waste recycling activities, safety and preventive measures and environmental awareness programs about consequences of human health exposure to e-waste pollutants should be in place, providing ventilating and dust control, training in safe work practices, use of personal protective equipment and medical surveillance. Besides, exposure to open dumping and burning during informal recycling should be significantly reduced and the remediation of contaminated sites because of informal recycling activities is also necessary for preventing additional exposure. The formal recycling can create an energy and resource-efficient economy. However, one major challenge of e-waste recycling is collection, which depends on the awareness, contribution, and support of the general public, governments, businesses, and other social organizations. To improve e-waste recycling, recycling capacity and relevant infrastructure for e-waste collection, separation, and processing also need to be substantially developed based on the concept of circular economy, especially in developing countries. Although numerous research activities

References

have been carrying on to develop effective recycling methods in order to integrate mechanical, chemical and biological approaches for e-waste recycling (particularly for metals), the further research should focus on how to improve resource recovery efficiency, minimize the loss of precious metals and avoid environmental and health risks during recycling. Furthermore, development of novel biotechnology and more efficient leaching mechanisms can make biological treatment attractive, especially for precious metal recovery. Additionally, building up proactive strategic programs and scientific assessment for the developed recycling technologies can also facilitate their implementation in terms of technical, environmental, economic, and social perspectives.

References 1 Baldé, C.P., Forti, V., Gray, V. et al. (2017). The Global E-Waste Monitor – 2017. Bonn/Geneva/Vienna: United Nations University (UNU), International Telecommunication Union (ITU) & International Solid Waste Association (ISWA). 2 Kumar, A., Holuszko, M., and Espinosa, D.C.R. (2017). E-waste: an overview on generation, collection, legislation and recycling practices. Resources, Conservation and Recycling 122: 32–42. 3 Awasthi, A.K., Wang, M., Awasthi, M.K. et al. (2018). Environmental pollution and human body burden from improper recycling of e-waste in China: a short-review. Environmental Pollution 243: 1310–1316. 4 Step Initiative (2014). One Global Definition of E-Waste. Bonn, Germany: United Nations University. 5 Chen, A., Dietrich, K.N., Huo, X. et al. (2011). Developmental neurotoxicants in e-waste: an emerging health concern. Environmental Health Perspectives 119: 431–438. 6 Crocker, R. (2012). Getting to zero waste in the new mobile communications paradigm: a social and cultural perspective. In: Designing for Zero Waste: Consumption, Technologies and the Built Environment (eds. S. Lehmann and R. Crocker), 115–130. London: Earthscan. 7 Statista (2015). Average lifespan of consumer electronics and tech devices in 2015. https://www.statista.com/statistics/688455/consumer-electronics-techdevice-average-lifespan (accessed 17 December 2019). 8 Frazzoli, C., Mantovani, A., and Orisakwe, O.E. (2011). Electronic waste and human health. In: Encyclopedia of Environmental Health (ed. J.O. Nriagu), 269–281. Elsevier. 9 Ilankoon, I.M.S.K., Ghorbani, Y., Chong, M.N. et al. (2018). E-waste in the international context – a review of trade flows, regulations, hazards, waste management strategies and technologies for value recovery. Waste Management 82: 258–275. 10 Dissanayake, V. (2014). Electronic waste. In: Encyclopedia of Toxicology, 3e (ed. P. Wexler), 568–572. Academic Press.

749

750

25 Resource Recovery from Electronic Waste

11 Pascale, A., Bares, C., and Laborde, A. (2018). Electronic waste. In: Encyclopedia of the Anthropocene, vol. 5 (eds. D.A. Dellasala and M.I. Goldstein), 269–275. Elsevier Inc. 12 UNEP (United Nations Environment Programme) (2013). Metal Recycling: Opportunities, Limits, Infrastructure, A Report of the Working Group on the Global Metal Flows to the International Resource Panel. Reuter, M.A., Hudson, C., van Schaik, A., Heiskanen, K., Meskers, C., Hagelüken, C. 13 ITU (International Telecommunication Union) (2016a). Measuring the Information Society Report 2016. https://www.itu.int/en/ITU-D/Statistics/Documents/ publications/misr2016/MISR2016-w4.pdf (accessed 17 December 2019). 14 Golev, A., Schmeda-Lopez, D.R., Smart, S.K. et al. (2016). Where next on e-waste in Australia? Waste Management 58: 348–358. 15 Electronics TakeBack Coalition (2014). Facts and Figures on E-Waste and Recycling. http://www.electronicstakeback.com/wp-content/uploads/Facts_and_ Figures_on_EWaste_and_Recycling.pdf (accessed 17 December 2019). 16 Tesfaye, F., Lindberg, D., Hamuyuni, J. et al. (2017). Improving urban mining practices for optimal recovery of resources from e-waste. Minerals Engineering 111: 209–221. 17 Li, J., Zeng, X., Chen, M. et al. (2015). “Control-Alt-Delete”: rebooting solutions for the e-waste problem. Environmental Science & Technology 49: 7095–7108. 18 Mwakabona, H.T. (2017). The growing global demand for element gallium: electrical and electronic waste and coal fly ash as alternative sources for its sustainable supply. International Journal of Science and Research 6: 1287–1293. 19 Cui, J. and Forssberg, E. (2003). Mechanical recycling of waste electric and electronic equipment: a review. Journal of Hazardous Materials 99: 243–263. 20 Xavier, L.H., Giese, E.C., Ribeiro-Duthie, A.C. et al. (2020). Sustainability and the circular economy: a theoretical approach focused on e-waste urban mining. Resources Policy, In press, https://doi.org/10.1016/j.resourpol.2019.101467, Article 101467. 21 Zeng, X., Gong, R., Chen, W.Q. et al. (2016). Uncovering the recycling potential of “New” WEEE in China. Environmental Science & Technology 50: 1347–1358. 22 Thierry, M., Salomon, M., van Nunen, J. et al. (1995). Strategic issues in product recovery management. California Management Review 37: 114–135. 23 Zoeteman, B.C.J., Krikke, H.R., and Venselaar, J. (2010). Handling WEEE waste flows: on the effectiveness of producer responsibility in a globalizing world. The International Journal of Advanced Manufacturing Technology 47: 415–436. 24 Li, J., Zeng, X., and Stevels, A. (2015). Ecodesign in consumer electronics: past, present and future. Critical Reviews in Environmental Science and Technology 45: 840–860. 25 Ikhlayel, M. (2018). An integrated approach to establish e-waste management systems for developing countries. Journal of Cleaner Production 170: 119–130. 26 Osibanjo, O. and Nnorom, I.C. (2007). The challenge of electronic waste (e-waste) management in developing countries. Waste Management & Research 25: 489–501.

References

27 Heeks, R., Subramanian, L., and Jones, C. (2015). Understanding e-waste management in developing countries: strategies, determinants, and policy implications in the Indian ICT sector. Information Technology for Development 21: 653–667. 28 Dai, Y., Huo, X., Zhang, Y. et al. (2017). Elevated lead levels and changes in blood morphology and erythrocyte CR1 in preschool children from an e-waste area. Science of the Total Environment 592: 51–59. 29 Widmer, R., Oswald-Krapf, H., Sinha-Khetriwal, D. et al. (2005). Global perspectives on e-waste. Environmental Impact Assessment Review 25: 436–458. 30 Ruan, J. and Xu, Z. (2016). Constructing environment-friendly return road of metals from e-waste: combination of physical separation technologies. Renewable and Sustainable Energy Reviews 54: 745–760. 31 Chauhan, G., Jadhao, P.R., Pant, K.K. et al. (2018). Nigam novel technologies and conventional processes for recovery of metals from waste electrical and electronic equipment: challenges & opportunities – a review. Journal of Environmental Chemical Engineering 6: 1288–1304. 32 Cui, J. and Zhang, L. (2008). Metallurgical recovery of metals from electronic waste: a review. Journal of Hazardous Materials 158: 228–256. 33 Namias, J. (2013). The future of electronic waste recycling in the United States: Obstacles and domestic solutions. Master thesis. Columbia University, New York. ´ Ž., Korac, ´ M., and Ranitovic, ´ M. (2011). Hydrometallurgical pro34 Kamberovic, cess for extraction of metals from electronic waste-part 2: development of the processes for the recovery of copper from printed circuit boards (PCB). Metalurgija – Journal of Metallurgy 17: 139–149. 35 Dorin, R. and Woods, R. (1991). Determination of leaching rates of precious metals by electrochemical techniques. Journal of Applied Electrochemistry 21: 419–424. 36 Hilson, G. and Monhemius, A.J. (2006). Alternatives to cyanide in the gold mining industry: what prospects for the future? Journal of Cleaner Production 14: 1158–1167. 37 Prasad, M.S., Mensah-Biney, R., and Pizarro, R.S. (1991). Modern trends in gold processing – overview. Minerals Engineering 4: 1257–1277. 38 Gurung, M., Adhikari, B.B., Kawakita, H. et al. (2013). Recovery of gold and silver from spent mobile phones by means of acidothiourea leaching followed by adsorption using biosorbent prepared from persimmon tannin. Hydrometallurgy 133: 84–93. 39 Aylmore, M.G. and Muir, D.M. (2001). Thiosulfate leaching of gold – a review. Minerals Engineering 14: 135–174. 40 Breuer, P.L. and Jeffrey, M.I. (2000). Thiosulfate leaching kinetics of gold in the presence of copper and ammonia. Minerals Engineering 13: 1071–1081. 41 Tripathi, A., Kumar, M., Sau, D.C. et al. (2012). Leaching of gold from the waste mobile phone printed circuit boards (PCBs) with ammonium thiosulphate. International Journal of Metallurgical Engineering 1: 17–21. 42 Ehrlich, H.L. (1997). Microbes and metals. Applied Microbiology and Biotechnology 48: 687–692.

751

752

25 Resource Recovery from Electronic Waste

43 Morin, D., Lips, A., Pinches, T. et al. (2006). BioMinE – integrated project for the development of biotechnology for metal-bearing materials in Europe. Hydrometallurgy 83: 69–76. 44 Yang, T., Xu, Z., Wen, J. et al. (2009). Factors influencing bioleaching copper from waste printed circuit boards by Acidithiobacillus ferrooxidans. Hydrometallurgy 97: 29–32. 45 Ilyas, S., Ruan, C., Bhatti, H.N. et al. (2010). Column bioleaching of metals from electronic scrap. Hydrometallurgy 101: 135–140. 46 Vestola, E.A., Kuusenaho, M.K., Närhi, H.M. et al. (2010). Acid bioleaching of solid waste materials from copper, steel and recycling industries. Hydrometallurgy 103: 74–79. 47 White, C., Wilkinson, S.C., and Gadd, G.M. (1995). The role of microorganisms in biosorption of toxic metals and radionuclides. International Biodeterioration & Biodegradation 35: 17–40. 48 Romera, E., González, F., Ballester, A. et al. (2006). Biosorption with algae: a statistical review. Critical Reviews in Biotechnology 26: 223–235. 49 Wang, J. and Chen, C. (2009). Biosorbents for heavy metals removal and their future. Biotechnology Advances 27: 195–226. 50 Vijayaraghavan, K. and Yun, Y.S. (2008). Bacterial biosorbents and biosorption. Biotechnology Advances 26: 266–291. 51 Nieboer, E. and Richardson, D.H.S. (1980). The replacement of the no-descript term “heavy metals” by a biologically and chemically significant classification of metal ions. Environmental Pollution. Series B: Chemical and Physical 1: 3–26. 52 Vilaplana, F. and Karlsson, S. (2008). Quality concepts for the improved use of recycled polymeric materials: a review. Macromolecular Materials and Engineering 293: 274–297. 53 Taurino, R., Pozzi, P., and Zanasi, T. (2010). Facile characterization of polymer fractions from waste electrical and electronic equipment (WEEE) for mechanical recycling. Waste Management 30: 2601–2607. 54 Ma, C., Yu, J., Wang, B. et al. (2016). Chemical recycling of brominated flame retarded plastics from e-waste for clean fuels production: a review. Renewable and Sustainable Energy Reviews 61: 433–450. 55 Tansel, B. (2017). From electronic consumer products to e-wastes: global outlook, waste quantities, recycling challenges. Environment International 98: 35–45. 56 Kajiwara, N. and Takigami, H. (2011). Contamination status of polybrominated diphenyl ethers released through primitive e-waste recycling and related future issues. Material Cycles and Waste Management Research 22: 159–168. 57 Peeters, J.R., Vanegas, P., Tange, L. et al. (2014). Closed loop recycling of plastics containing flame retardants. Resources, Conservation and Recycling 84: 35–43. 58 Sahajwalla, V. and Gaikwad, V. (2018). The present and future of e-waste plastics recycling. Current Opinion in Green and Sustainable Chemistry 13: 102–107. 59 Zhao, Y.B., Lv, X.D., and Ni, H.G. (2018). Solvent-based separation and recycling of waste plastics: a review. Chemosphere 209: 707–720. 60 Liang, R. and Gupta, R.K. (2001). Effect of residual impurities on the rheological and mechanical properties of engineering polymers separated from mixed

References

61

62 63 64 65 66

67

68

69 70 71

72

73

74

75

plastics. Proceeding of the 59th ANTEC 2001 (6–10 May), Dallas, USA. Society of Plastics Engineers, p. 2753. Bernardes, A.M., Espinosa, D.C.R., and Tenório, J.A.S. (2004). Recycling of batteries: a review of current processes and technologies. Journal of Power Sources 130: 291–298. Swain, B. (2017). Recovery and recycling of lithium: a review. Separation and Purification Technology 172: 388–403. Huang, B., Pan, Z., Su, X. et al. (2018). Recycling of lithium-ion batteries: recent advances and perspectives. Journal of Power Sources 399: 274–286. Zheng, X., Zhu, Z., Lin, X. et al. (2018). A mini-review on metal recycling from spent lithium ion batteries. Engineering 4: 361–370. Gruber, P.W., Medina, P.A., Keoleian, G.A. et al. (2011). Global lithium availability. Journal of Industrial Ecology 15: 760–775. Ordoñez, J., Gago, E.J., and Girard, A. (2016). Processes and technologies for the recycling and recovery of spent lithium-ion batteries. Renewable and Sustainable Energy Reviews 60: 195–205. Dehghani-Sanij, A.R., Tharumalingam, E., Dusseault, M.B. et al. (2019). Study of energy storage systems and environmental challenges of batteries. Renewable and Sustainable Energy Reviews 104: 192–208. Dewulf, J., Van Der Vorst, G., Denturck, K. et al. (2010). Recycling rechargeable lithium-ion batteries: critical analysis of natural resource savings. Resources, Conservation and Recycling 54: 229–234. Georgi-Maschler, T., Friedrich, B., Weyhe, R. et al. (2012). Development of a recycling process for Li-ion batteries. Journal of Power Sources 207: 173–182. Xu, Y., Li, J., Tan, Q. et al. (2018). Global status of recycling waste solar panels: a review. Waste Management 75: 450–458. Aman, M.M., Solangi, K.H., Hossain, M.S. et al. (2015). A review of safety, health and environmental (SHE) issues of solar energy system. Renewable and Sustainable Energy Reviews 41: 1190–1204. Ardente, F., Latunussa, C.E.L., and Blengini, G.A. (2019). Resource efficient recovery of critical and precious metals from waste silicon PV panel recycling. Waste Management 91: 156–167. Wambach, K. (2017). Life cycle inventory of current photovoltaic module recycling processes in Europe. http://www.iea-pvps.org/index.php?id=9&eID=dam_ frontend_push&docID=4239 (accessed 17 December 2019). Berger, W., Simon, F.G., Weimann, K. et al. (2010). A novel approach for the recycling of thin film photovoltaic modules. Resources, Conservation and Recycling 54: 711–718. Grant, K., Goldizen, F.C., Sly, P.D. et al. (2013). Health consequences of exposure to e-waste: a systematic review. The Lancet Global Health 1: 350–361.

753

755

Index a absorbed dose 528, 530, 548, 549 absorption process 573, 710 CO2 removal 711–713 acid/alkali leaching 737 acidogenesis 24, 279, 284, 290, 313, 314, 317, 318, 321, 322, 324–326, 330, 331, 333 acidogenic bacteria 279 activated carbons (ACs) 217, 222, 224–226 activated rice husk (ARH), gold sorption of 477, 480 activated sludge (AS) 6, 8, 123, 127, 217–219, 223, 275–276, 293, 305–306, 317–318, 321, 326–329, 331–333, 374, 388, 490–491, 507, 550, 566 adenosine triphosphate (ATP) 339 adsorbents reusability and phosphorus recovery 364–365 adsorption process 407, 416, 479, 573, 574, 710 CO2 removal 712 description 409 Freundlich isotherm 409 isotherm models 409, 410–411 kinetic models 407–409 mass transfer 409 NO3 − and PO4 3− removal 401 adsorption yield (AY) 423, 427 advanced oxidation process (AOP) 525, 544

aerobic FO-based systems 14, 377, 383, 388, 389 aerobic granular sludge 217, 219, 220–222, 229–230, 235 laboratory-scale application 220 aerobic osmotic membrane bioreactor (OMBR) 377, 380, 381–383 algae based birefinery system 212, 213 algal energy and bio-product formation 200–202 algal process system LCA 210–211 technological issues 208–210 alkaline pretreatment 283, 289, 293 amine scrubbing 711 ammonia stripping 143, 144, 149, 150, 489, 499, 570–573, 577, 585–587 ammonium removal methods, in wastewater treatment 489 anaerobic digestion (AD) process 123, 147, 149, 565 bioenergy recovery 279–280 of microalgae biomass 582 theory of 279 anaerobic FO-based systems 377, 383, 388, 389 anaerobic membrane bioreactors (AnMBRs) 565, 566 anaerobic OMBR (AnOMBR) 377, 383 animal husbandry 142 animal wastes 154, 560, 585

Sustainable Resource Management: Technologies for Recovery and Reuse of Energy and Waste Materials, First Edition. Edited by Wenshan Guo, Huu Hao Ngo, Rao Y. Surampalli, and Tian C. Zhang. © 2021 WILEY-VCH GmbH. Published 2021 by WILEY-VCH GmbH.

756

Index

anion exchange membrane (AEM) 494, 495, 506 AnOMBR-based systems, for nutrient recovery 383 aquaculture wastewater 561– 563 characteristics 561 aqueous phases, of HTL 91, 92 Aspergillus oryzae palletization 173 Aurantiochytrium sp. 177, 183 auto-flocculation 173, 208 automatic rotating filter, for wastewater 647 automatic strainer, for wastewater 647, 648 auxochromes 535

b Bacillus licheniformis 173 bacterial decomposition 339 BES-FO hybrid system 504, 510 Best Available Techniques (BAT) 152 biobased jet fuel 168 biochar 437, 438, 564 ash/mineral content 441 characteristics 439–442 feedstocks 442, 443, 444 fractions 35 macroporosity 441 pH of 451 phosphorous recovery 442–443 cationic impregnation 449–450 magnetic biochar composites 454–455 metal (hydr)oxide and layered double hydroxide composites 450–454 post-pyrolysis modification 449 pre-pyrolysis modification 443, 455 properties 440 pyrolysis temperature 440 surface area 441 biochemical oxygen demand (BOD) 114, 117, 123, 219, 533 removal 114

bio-crude 21, 27– 30, 32, 34–36, 40, 258 fractions 35 yield 29, 30, 36 biodiesel 25, 26 jet fuel 168, 169 microalgae, selection of 169–170 bioelectrochemical system (BES) 15, 208 advantages 490 ammonium recovery 498, 499 application for resource recovery 507–508 challenges 508–510 chemical recovery 505–507 energy recovery 496–498 evolution of 493–496 metals recovery 501–503 nutrient recovery 498–501 phosphate recovery 498–500 types 493 water recovery 503–505 bioelectrochemical treatment system (BET) 208 bioelectrogenesis 208 bioenergy 11–12, 15, 26, 83–98, 125, 147, 149, 156, 199, 200, 206, 275–294, 560, 564, 581, 584–587 bio-energy production, limitations 98 aqueous by-products utilization 97–98 poor quality, of crude bio-oil 97 bio-fertilizer 23, 208 biofilter digester 114 bio-flocculant 172, 173, 236–237 biofuel microalgae cultivation 170–172 microalgae harvesting 172–174 bio-flocculation 172–173 magnetic separation 173–174 oil, extraction of 174–175 biogas potential, human faeces 110 biohydrogen 7, 24, 209, 564 bio-hydrology 481

Index

biohydrometallurgical process 738–740, 741 benefits and issues 739, 741 biohydrometallurgy 467, 468, 735 bioleaching 16, 142, 263, 264, 467, 472, 481, 601, 624–631, 738, 739, 744 bioleaching residues SEM analysis 630–631 XRD analysis 629–630 bioleaching treatment 467 bioleaching, of Si flask experiments 624, 625 heavy metals recovery 628–629 optimization of parameters 626–628 silicate bacteria culture medium 624 silicon bacteria growth, time courses of 625 biological approach, for resource recovery anaerobic digestion 565–569 dark fermentation (DF) 569–570 biological methods 139, 340, 400, 401, 530, 544, 605, 632 NO3 − and PO4 3− removal 401 biological mineralization 208 biological nitrogen removal (BNR) 218, 308, 316 biomass 83 cultivation 202, 206 harvesting 172, 173, 175, 178, 179, 183 biomass conversion process 27, 581, 582 of aquatic biomass 581, 582 bio-oils 32, 35, 36, 84, 88, 90, 91, 97, 255, 259–260, 266, 267, 583 from food waste 37–39 bioplastics 7, 26, 207 biorefinery and integrated approaches 207–208 biosensors 465, 466 biosolids 8

bio-sorbents 142, 218, 229–236, 237–238, 402, 472–478, 480, 481, 739 biosorbent, removal of pollutant heavy metal sorption 226, 237 organic pollutant sorption 237–238 biosorption process 739, 740 biosorbent source and selection 473 commercial sorbents 477–478 description 472 gold biosorption by industrial biomass 473 algae 474–476 bacteria 474 fungal microbes 476 plant material 476–477 of gold particle from biomass 478–479 blowers 279, 706, 707 blue baby syndrome 141, 400 breeding wastes management 560–563 breeding wastewater 585 anaerobic digestion process 565–569 characteristics 560–563 dark fermentation (DF) 569–570 microalgae and duckweed culture 583 microalgae/duckweed biomass production and nutrient removal 578, 579 nutrient recovery from biological approach 565–570 physicochemical approach 570–578 organic wastewater in 565–569 plant-based treatment approach 578–583 resources in bioenergy 564 nutrients 563–564 water 563

757

758

Index

breeding wastewater (contd.) thermochemical approach 583–584 value-added bioproducts 564–565 brominated flame retardants (BFRs) 734, 740, 742 bubbling fluidized bed (BFB) 254, 256, 257 bubbling fluidized bed reactor (BFBR) 254

c CAA (clean air act) 153, 699 Ca-based phosphate (hydroxyapatites) 262 calcium peroxide (CaO2 ) 289, 290 Ca-modified biochar 455 capital expenditure (CAPEX) 30, 204, 209–211 catalyst concentration 36–37 catalytic pyrolysis 255, 743 cation exchange capacity (CEC) 114, 260, 261, 565 cation-exchange membrane (CEM) 386, 494, 495, 500, 506 cationic impregnation, of biochar 449–450 CdTe thin-film modules recycling technique 747 cell-independent sorption 474 cellulose triacetate (CTA) membrane 386 centrifugal blowers 707 CFD-DEM 684 channel heat exchanger-type 644, 645 channel heat exchanger-type 1 644, 645 characteristics of, Fe3 O4 , SiO2 @Fe3 O4 & ZrO2 @SiO2 @Fe3 O4 345–348 chars 441, 583 chemical fertilizers 113, 128, 373, 398, 490 chemical leaching 142, 467, 737 chemical looping 255

chemical oxygen demand (COD) 5, 8, 114, 141, 219, 276, 290, 309 chemical precipitation 13, 14, 226, 232, 262, 305, 306, 307, 309, 320, 340, 374, 375, 377, 379, 380, 381, 382, 386, 387, 388, 389, 401, 489, 490, 499, 501, 502, 509, 540, 560, 574, 577, 744 chemically enhanced primary sedimentation (CEPS) 13, 305 chemisorption 232, 408, 411, 416, 425, 473 chitin 235, 474, 478, 481 Chlorella pyrenoidosa 170 Chlorella regularis 173 circular economy 10, 12, 47, 52, 64, 76, 77, 157, 158–159, 160, 161, 724, 731, 748 business model 158–159 circular economy concept 731 circulating fluidized bed (CFB) 253–257 circulating fluidized bed incineration technology (CFBIT) 253–257 circulating fluidized bed reactor (CFBR) 255 citric acid-soluble manganese 622 cleaning system, with circulating rubber balls 654, 655 climate change (CC) 1, 22, 50, 76, 126, 151, 157, 199, 250, 605, 708, 745 Clostridiaceae 25 coagulants 305, 308, 309, 311, 312, 540, 542–544 coagulation 151, 208, 232, 307, 309–311, 400, 401, 525, 530, 533, 540, 542, 543, 605 coefficient of performance (COP) 7, 640, 655 collection piping system, LPG recovery 706–707 combined heat and power (CHP) systems 6, 219, 279, 715 competent applications, of algae 205–207

Index

compounded annual growth rate (CAGR) 205 comprehensive nutrient management plan (CNMP) 153 comprehensive separation efficiency, of hydrocyclones 666, 667 compressed natural gas (CNG) 715 Computational Fluid Dynamics coupled with Discrete Element Method (CFD-DEM) 684 concentrated animal feeding operations (CAFOs) 153, 559 condensate knock-out system 707, 708 condensate management system 704, 707 condensation 88, 93, 96, 261, 264, 438, 657, 709, 710, 713 cone pyrolyser 255 Congo Red (CR) adsorption 238 consumer electronics, average lifespans for 725 continuous stirred-tank reactor (CSTR) 565, 566, 570 Convention on Long-Range Trans-boundary Air Pollution (CLRTAP) 151 conventional human excreta management systems 110, 111 conventional microalgal biomass harvesting techniques 172 conversion methods, analysis of 202 algal biomass composition 202–203 conversion routes 203–204 product yield and market value 204–205 co-precipitation of layered double hydroxides (LDH) 449, 452 co-precipitation, of metal hydroxides 451 COP value, of WWSHP system 635–637, 639–644, 646–648, 651, 652, 654–660

copper metal recovery 470 CO2 removal 711 absorption process 711–712 adsorption process 712 cryogenic process 713 membrane separation 712–713 Corynebacterium glutamicum 474 cross-draft gasifier 256 cryogenic process 713 crystal violet (CV) 230, 231, 232 crystallization processes 340, 341 NO3 − and PO4 3− removal 401 CWA (clean water act) 153 cyanide leaching 472, 474

d dark fermentation (DF) 208, 288, 560, 569–570, 582, 585 Decentralised wastewater treatment systems 123 Department of Environment Food and Rural Affairs (DEFRA) 58 depolymerization 27, 32, 34, 89, 90, 93, 438 desorption of PO4 3− 417, 419, 425 desorption process 366, 473, 477, 478, 481 desorption, of NO3 − 417, 420, 425 digester 153, 155, 156, 276, 279, 284, 500, 566 direct liquefaction 27 direct wastewater source heat pump system 637, 638 Docosahexaenoic acid (DHA) 12, 168–169 domestic wastewater treatment 6, 123, 217, 491, 507 dry-expansion shell-and-tube evaporator 641 Dual Inconel pipe counter-current heat exchanger 29 duckweed-based bioenergy 586 dyeing process 534, 535, 538, 549 dynamic light scattering (DLS) 345

759

760

Index

e earthworm composting 138 e-beam process 526 ecological sanitation 11, 111, 112, 113, 115, 117, 119, 120, 126, 129, 130, 131 economical use, of LFG 714 economics of resource recovery 127–128, 157–158 effluent organic matter (EfOM) 227, 228, 229 Eicosapentaenoic acid (EPA) 12, 168–169 Eisenia foetida 138 electrical conductivity (EC) 143, 440 electrical hydrocyclones 677 electrochemical hydrocyclones 677, 678 electrodialysis (ED) 7, 375, 401, 576, 577, 578 electrokinetic remediation technology 264 electrolytic manganese metal (EMM) production process 601, 602–605 wastewater 602–604 electrolytic manganese residue (EMR) 602, 604 characterization of 615–616 chemical composition 615 components 604 leaching, of toxic elements 622–624 morphology evolution 622, 623 reutilization 605 silicon activation 617–618 distribution characteristics 615, 616 TG, DTG and DSC curves 617 transformations during bioleaching process 629–631 XRD pattern 615–616, 619–620 electrolytic manganese wastewater (EMW) 16, 602, 604, 614 electromagnetic hydrocyclones 677–678

electromagnetic spectrum, components of 526–527 electron beam (EB) process defined 526 irradiation 526 and membrane bioreactor process, on COD removal 547 for wastewater treatment 528–533 aeration biological treatment 545 economic feasibility 549–552 industrial applications 548–549 lab-scale tests 544–548 limitations 551–552 quantitative effects of 528 transformer oil treatment 530 electronic waste (e-waste) 9, 17, 723 categories 724–725 chronic hazards 726 collection 731–734 defined 724 generation 725 generation rates 723 global management 732 hazardous substances 727–729 health impacts of 726–729 health risks 726 lithium ion batteries, recovery of 743–745 management regulations 724 metal concentrations 730 metals recovery from 734–735 physical separation technologies 734 plastics recovery 740–743 recycling benefits 729–731 challenges 733, 748 in developing counties 733 pre-processing 734 reasons for 734 volume 723 worldwide generation 725 energy consumption system 635 energy recovery processes 39

Index

energy-dispersive X-ray spectroscopy (EDX) 237, 316, 345, 380 enhanced biological P removal (EBPR) 305, 307, 489 enhanced-separation hydrocyclone technologies 668 challenges 683–685 perspectives 684–685 systematic and comprehensive studies 685 Enterococcus faecalis 152 environment sanitation and financing 159–160 ethanol 1, 24, 25, 30, 88, 91, 209, 210, 343, 344, 505, 564, 583 EU Common Alerting Protocol (CAP) 151, 152 Eudrilus esugeniae 138 excretion, frequency of 109 expanded granular sludge bed (EGSB) reactor 219, 565 Extended Producer Responsibility (EPR) 732, 748 extracellular polymeric substances (EPS) 13, 173, 218, 276, 282, 283 extraction wells, LFG 704, 708

f faeces 11, 110 Fe-impregnated biochar materials 450 Fe0 /Fe3 O4 composites characteristics of 350–351 synthesis of 344– 345 Fe3 O4 and La(OH)3 /Fe3 O4 nanocomposites, characteristics of 348– 349 Feed inlet system 28, 29 fenvalerate, removal efficiency of 530 fertilizers 578 chemical 373 direct land application of 374 filamentous fungi 173, 476 fish-hook effect 675

fixed bed gasifier (FXBG) 255, 256 flocculant-assisted hydrocyclones 677, 679 flocculation mechanism 237 flue gas desulfurization method 465 fluid flow, in hydrocyclones 16, 665–666 fluidized bed furnace (FBF) 250–253 fluidized bed reactor (FBR) 40, 254, 255 fluorescence spectroscopy 237 FO-based system CTA vs. TFC membrane 386 for nutrient recovery challenges 387–388 recommendations 385–387 focused pulsed (FP) pretreatment 282 Fomitopsis carnea 476 food waste (FW) 1, 7–8, 10, 21–40, 52, 83, 148, 260, 318, 321–324, 326–328, 331–333 composition and reaction 36–39 Formaldehyde (FA) 90, 477, 543 forward osmosis (FO) process 14, 375, 499, 576 configurations 377 near coastal area, advantages of 379 for nutrients recovery 377–385 economic feasibility 379 pH values and chemical dose 386 vs. pressure-driven membrane technologies 385 fossil fuels 24, 31, 38, 83, 84, 124, 125, 149, 168, 178, 179, 201, 258, 279, 294, 584, 586, 605, 744, 635 fossil resource depletion (FD) 157 foulant growth law, in WWHEs 657 Fourier transmission infrared spectra (FT-IR) 226, 345, 620 free ammonia (FA) 13, 283–284, 289 pretreatment 283–284

761

762

Index

free nitrous acid (FNA) 13, 283, 289 pretreatment 283 Freundlich equation 232 Freundlich isotherm 233, 409 Freundlich parameters 480 fuel cells and biosensors 466 Full Recovery End of Life Photovoltaics (FRELP) 747 furnace chamber 253

g gamma irradiation pretreatment 284, 288 Gantt chart 607 gas chromatography (GC) 38, 90, 258 gasification technology 257, 584 gel permeation chromatography 90 genetically modified microalgae/ thraustochytrids 184–185 glassy polymers 712, 713 global E-waste Monitor 2017 723, 725 global e-waste production 467 global food waste production advanced food waste management methods acidogenesis 24 biodiesel 25– 26 bioplastics 26 solventogenesis 24–25 conventional food waste management practices composting 24 fertilizer/animal feed 23 incineration 23 land filling 23 global water demand 397 glycol absorption process 710 gold biocompatibility property 466 biosorption by industrial biomass 473 bacteria 474 as catalyst 464 chemical processing 464–465

demand for 463 electronic industry 466 medical and biomedical application 466 nanostructure-based electrodes 466 recovery from e-waste 470–472 gold biosorption by industrial biomass algae 474–476 fungal microbes 476 plant material 476–477 gold colloids 466 gold nanoparticles 466, 476 granular activated carbon (GAC) 404, 438, 441, 530, 564 granular sludge 13, 120, 217–225 granular sludge biosorbent 229–236 heavy metal contained wastewater AnGS, for Pb(II) and Cu(II) removal 234–235 Cu(II) sorption 232–233 Ni(II) sorption, onto AGS/AnGS 233–234 Zn(II) sorption 232 treatment biosorption, of dye wastewater 230–232 role of EPS 229–230 granulation process 219–221 gravity-film heat exchanger 642, 643 green chemistry 464 greenhouse gases (GHGs) emission 22, 31, 32, 137, 699

h Haber–Bosch process 373, 389 halide leaching method 738 hazardous substances, in e-waste 158, 727, 729, 742 heat exchanger with cleaning function by strong flushing 653 with fouling fluidized-removing 654 with nylon brushes 652

Index

heat pumps 7, 16, 635, 636, 637–640, 659 heat recovery from raw wastewater 637–639 helical heat exchangers 643 hemicellulose 23, 34, 85, 87, 88, 90, 92, 93, 138, 438, 582 heterotrophic microalgae/ thraustochytrids 182 heterotrophic or phototrophic microbes 6 hierarchy of resource use (HRU) 2, 3 high karat gold 463 high-income country (HIC) 4, 5 higher heating value (HHV) 91, 122, 258, 260, 583 homoacetogens 279 horizontal gas collection wells advantages and disadvantages 704, 706 configuration 704, 705 household waste composting 68, 69 household waste recycling 68, 69 Howdon domestic wastewater treatment plant 507 human dietary changes 559 human toxicity (HT) 157 hydraulic oscillation system 29 hydraulic retention time (HRT) 155, 312, 316, 318, 508, 547, 549, 565, 566, 569, 570 hydro-denitrogenation 37 hydro-deoxygenation reactions 37 hydro-desulfurization 37 hydro-metallurgical treatment 467 hydro-thermal liquefaction (HTL) 21–33 FW treatment 32–37 reaction time 35 solid to solvent ratio 35– 36 temperature 34–35 reactor 28 hydrochar 90, 91, 454 hydrocyclones 16, 663 applications 680–683

comprehensive separation efficiency 667 cut size 667 enhanced by adjusting back pressure 679–680 enhanced by control particles 679 enhanced by flotation 679 enhanced by monitoring and automatic control 680 fluid flow characteristics 665– 666 geometric parameters of 668 cone angle 671 conical-section shape 671–672 cylindrical-section diameter 668 cylindrical-section length 668 inclination angle 672 inlet shape 669 inlet size 668–669 inlet-section angle 669 overflow diameter 670 ratio of underflow diameter to overflow diameter 671 underflow diameter 670 underflow-pipe shape 671 vortex-finder length 669 vortex-finder shape 669–670 vortex-finder thickness 669 inclination angle 672 operating conditions 677 operating parameters of 674 feed concentration 675 feed density difference 675 feed flow rate 674 feed fluid viscosity/ rheology 676 feed particle arrangement 676 feed particle shape 675–676 feed particle size 675 feed pressure 674–675 publications and citations 663–664 reduced separation efficiency 667 separation efficiency 666–667 split ratio 667 total static pressure drop 667

763

764

Index

hydrocyclones (contd.) water-injection 673 with inner cone 672–673 with reflux device 673 with solid rod 672 working principle of 663, 665 hydrogen peroxide 465, 466, 467, 525, 544, 747 hydrogen production from sludge 284 biological pretreatment 290 chemical pretreatment 289 acid pretreatment 289 alkaline pretreatment 289 CaO2 pretreatment 290 free ammonia pretreatment 289 free nitrous acid pretreatment 289 ozone pretreatment 289 TCC 290 wet oxidation pretreatment 289– 290 physical pretreatment 284 freezing/thawing pretreatment 288 gamma irradiation pretreatment 288 microwave pretreatment 288 sterilization pretreatment 288 thermal pretreatment 284–288 ultrasonic pretreatment 288 hydrogen sulfide removal 711 hydrogen-oxidizing bacteria 9 hydrogenotropic methanogens 279 hydrogentrophic methanogenesis 279 hydrometallurgical process 8, 467, 736–740, 741, 744, 747 benefits and issues 739, 741 hydrometallurgy 467, 468, 469, 481, 735, 747 hydrothermal carbonization (HTC) 203, 26 hydrothermal liquefaction (HTL) 26, 84, 257 of food waste 26–32

fast HTL 30 GHG 31–32 iso-thermal HTL 30 products 30–31 reactor operation 27–29 technology 10 hydrothermal liquefaction, of lignocellulosic biomass 91 bio-oil 92–93 catalysts 95–96 cellulose, and degradation 87–88 composition of 85 development, of HTL technology 85 heating rate 93–94 hemicellulose, and degradation 88 lignin, and degradation 88–90 liquid-to-solid ratio 96–97 pressure 94–95 products description bio-oil 90 other by-products 91 solid residue 90– 91 residence time 94 temperature 93 hydrothermal process 38, 583, 584 hydrothermal pyrolysis 255 hydrothermal technology 85 hydrothermal treatment (HTT) 26, 27, 454 hydrous iron oxides (HFO) 327, 332, 333 hydrous pyrolysis 27 hydroxyacetone 88 hydroxyapatite formation 150 5-hydroxymethyl furfural (HMF) 34, 37 p-hydroxyphenyl propanoid units 88

i incineration 23, 248, 250, 257, 261, 275, 417, 481, 736 indirect wastewater source heat pump system 637–638

Index

inductively coupled plasma atomic emission spectroscopy (ICP-AES) 345 industrial wastewater 492 characteristics 492–493 COD concentration 492 inner cones, of hydrocyclones 672–673 inorganic coagulants 540, 543, 544 integrated microalgae/thraustochytrids cultivation 183–184 integrated microalgae/thraustochytrids system 186 intergovernmental panel on climate changes (IPCC) 151 internal circulation (IC) 219 International Energy Agency (IEA) 205 intra-particle diffusion process 408, 416–417, 480 ion exchange method 573, 574 NO3 − and PO4 3− removal 401 ionizing process 526 ionizing radiation 526, 528 iron ore reduction 255 iron-based adsorbents, for phosphate removal 364 isotherm models 407–410

j Johkasou system 123–124

l Lagerstroemia speciosa 477, 480 La(OH)3 /Fe3 O4 nanocomposites, synthesis of 344 land-farming 247–248, 266 Landfill Allowance Trading Scheme (LATS) 66, 68, 71 landfill gas (LFG) 16, 699 applications 715 collection 699 collection efficiency 699–700, 703

LandGEM model 701–702 vs. LFG recovery 704, 705 regression model 703 theoretical model 702–703 conversion 716–717 generation and recovery projection 704 recovery system 700, 704 collection piping system 706–707 condensate management system 707 extraction wells 704, 705 wellheads 704, 706 site assessment 700–701 cover and liner system, type of 701 depth of landfill 701 moisture 701 waste characteristics 701 utilization combined heat and power systems 715 direct use of 714 electricity generation 714–715 high Btu applications 715 Landfill Methane Outreach Program (LMOP) 716 landfills 699, 719 and state-level projects 716 tax escalator 71 landfilling process 23, 699 LandGEM model 701–702 Langmuir isotherm model 228, 230, 422 Langmuir model 14, 224, 226, 232–233, 351, 353, 356, 409, 421, 480 lannate, removal efficiency of 530 layered double hydroxides (LDH) 449, 452–455 structure of 453 leachate 217, 250, 306, 385, 398, 400, 478, 481, 492, 503, 602–604, 614, 622, 632, 706 leaching of ambient acid (AAL) 470

765

766

Index

leaching of high pressure (HPAL) 470 leaching process 263, 472, 476, 604, 606, 736 leaching, of toxic elements 622 LFG extraction site plan sample 707 life cycle analysis (LCA), of hydrothermal algal biomass 584 Life Cycle Assessment (LCA), of sanitation systems 125–126 ligand exchange mechanism 416, 451 light brown suspension 276 lignocellulosic biomass 11, 83–98, 203, 254, 260, 582, 584 linear combination fitting (LCF) 327–328 Lipomyces sp. 26 liquefied natural gas (LNG) 713, 715 lithium-ion batteries (LIBs) 743 advantages 743 recycling process, flowchart of 745 spent 743, 744 livestock wastewater 560–561 anaerobic digestion, performance of 567 characteristics 561 local authorities (LAs) 10–11, 47, 50, 52–54, 56–58, 64–65, 71–75 low energy mainline (LEM) 6 low-cost adsorbents for nutrient removal 403 adsorbent dose effect 403–404 contact time 404–405 initial adsorbate concentration 403 interfering anions effect 406–407 pH effect 405–406 temperature effect 404 for pollutants removal 402 regeneration and reuse of 426–427 low-cost WW decontamination technologies 647

anti-fouling technology 648–654 filtration technology 647–648 low-karat gold 463 lower heating value (LHV) 259 lower middle-income country (LMIC) 4, 397 low-cost de-foulant methods 657 Lumbricus rubellus 138 Lyngbya majuscula 475, 476 lysis-thickening centrifuge pretreatment 281–282

m magnesium-hydroxide flue gas desulfurization (MFGD) process 575 magnetic biochar composites 454–455 magnetic fluids hydrocyclones 678 magnetic iron based-oxide materials, characterization methods 345 magnetic nanomaterials 343 magnetic separation 173–174, 343, 347, 349–350, 365–366, 603, 734, 744 Maillard reaction 35–36 Malachite green (MG) 224 manganese carbonate, thermal stability of 602, 604, 606, 611, 613–614, 632 manganese citrate 622 manganese slag 605 manganese-bearing wastewater manganese recovery CO2 dosage effect 607–609 ion recovery efficiency 611 manganese carbonate, thermal stability of 613, 614 onsite CO2 emission 606–607 operational modes 611 pilot treatment system and performance 609–611 precipitates, characteristics of 611–613 treatment strategy 605–606

Index

operational modes 609 treatment strategy 605–606 manure management anaerobic digestion/co-digestion 147 centralized and de-centralized models 148–149 composting/co-composting 147–148 energy production 149 mineral reutilization 150–151 ammonia stripping 150 struvite crystallisation 150 nutrient recovery, from manure 142–147 sanitization and hygiene aerobic composting 139–141 heavy metal recovery, from livestock manure 142 nitrogen and phosphorus recovery 141 material recycling 10, 746 material synthesis, development and Fe3 O4 nanoparticles, synthesis of coprecipitation method 343 solvothermal method 343 MEC-FO system 385, 389 for ammonium and water recovery 385 for nutrients recovery 384 MEC/MRC-AnOMBR system 384 mechanical plastic recycling 742 mechanical-baking coupling method 622 mechanical-grate incineration technology (MGIT) 253 membrane bioreactors (MBRs) 6, 13, 127, 306, 308–309, 317–326, 377, 546, 565 membrane distillation (MD) 39, 375, 576–577 membrane separation process 375, 712 membrane technique NO3 − and PO4 3− removal 401 metal biosorption process 472

metal coagulants 542 metal extraction biohydrometallurgical process 738–740 hydrometallurgical process 736–738 metallurgical process for 735 pyrometallurgical process 736 metal impregnation, of biochar 449 metal recovery 142, 463, 469, 472, 476, 490, 501–503, 608, 733–734, 738, 739, 749 metal regeneration 472 metallic manganese 601 methane (CH4 ) 122, 280–284, 505, 566, 699–719 methane emissions LandGEM model 701–702 regression model 703 theoretical model 702–703 methane production, from sludge 277, 280 biological pretreatment 284 chemical pretreatment alkaline pretreatment 283 FA pretreatment 283–284 FNA pretreatment 283 ozonation 283 physical pretreatment 280 FP pretreatment 282 mechanical pretreatment 281–282 microwave pretreatment 282 THP 280, 281 ultrasonic pretreatment 282 methane recovery 708–709 absorption process 710 adsorption process 710 condensation 710 particulates filtration 711 stages of 710 systems 704–708 and utilization 717 economic challenges 718 impurities 718 low methane production 718

767

768

Index

methane recovery (contd.) regulatory challenges 719 social challenges 719 technical challenges 718 methane utilization 713–717 methanol 25, 30, 34–35, 91, 174, 204, 259, 716 methyl orange (MO) 230 MFC-FO systems 389 Mg-accumulated tomato tissue biochar 443 Mg-Al-Cl-LDH-hydrochar composite material 454 Mg-Al-NO3 -LDH-functionalised biochars 454 Mg-modified biochar 455 microalgae 167 biorefinery 209 as feed ingredients 565 and thraustochytrids DHA and EPA productions 179–183 microalgae/thraustochytrids-based biofuel production 12 microalgal biofilm system 183–184 microalgal oil biodiesel production 176 jet fuel production 176, 177 microbial desalination cell (MDC) 208, 493–495, 504–505, 507 microbial electrolysis cell (MEC) 208, 383, 385, 389, 493–495, 498–500, 503, 505, 507 microbial electrolysis desalination cell (MEDC) 506–507 microbial electrosynthesis (MES) 493–496, 505–506, 510 microbial fuel cells (MFCs) 119, 208, 383, 389, 493–495, 497, 500, 507 microbial solar cell (MSC) 493, 496 microfiltration (MF) 382, 576 membrane 382 microwave pretreatment 282, 288, 293 microwave-assisted pyrolysis 255

Millennium Development Goals (MDGs) 128–129 mineral concentrates 150–151 mixed integer nonlinear programming (MINLP) 204 mixed-liquor suspended solids (MLSS) 220–221, 317, 545–547 modular systems 644, 646 molecular polymerization 93 monounsaturated fatty acids (MUFAs) 169, 177, 181, 186 multi-component contaminants, treatment of 235–236 multi-hydrocyclone arrangements 668, 673 multi-scale governance (MSG) 50 multilateral environmental agreements (MEA) 151 multiple hearth furnace (MHF) 250, 252 municipal solid waste (MSW) 8 landfills 700–708 management 699–700 municipal wastewater 5–6, 16, 130, 217, 219, 221–222, 305, 307, 319, 389, 397, 491, 530, 544 global quantity of 492

n Nannochloropsis oculata 170 nanofiltration (NF) 7, 306, 576 National Synchrotron Radiation Research Center (NSRRC) 327 net energy ratio (NER) 211, 496 nickel metal recovery 469–470 nitrate pollution 400 nitrogen-containing wastewaters 217 NO3 − removal 400, 402 adsorbents for 414–415 and recovery using Prosopis juliflora weed 421 low-cost adsorbents adsorbent dose effect 403–404

Index

contact time 404–405 initial adsorbate concentration 403 interfering anions effect 406–407 pH effect 405–406 temperature effect 404 non-conventional human excreta management systems 111 non-conventional sanitation systems 110, 112, 126 non-ionizing radiation 526 non-point source, of nitrate pollution 400 NPDES (National Pollutant Discharge Elimination System) 153 N-removal processes 119 nucleophilic addition reaction 535, 538, 542 nucleophilic substitution reaction 535, 538, 539 nutrients 398, 563, 564 pollution nitrate pollution 400 NO3 − and PO4 3− removal methods 400–401 phosphate pollution 398–399 potential sources of 398, 399 recovery 1 chemical precipitation 374 MEC-FO system for 384 OMBR-based system for 381 organic contaminants 375 systems 156 removal disadvantages of 374 utilization 118

o Oleaginous yeasts 26 OMBR-based system, for nutrients recovery 380–383 OMBR-MF hybrid system, for nutrient recovery 382 OMBR-MF system 382–383 open raceway pond (ORP) system 179, 183, 208, 210, 322, 329

operating expenditure (OPEX) 30, 204, 209, 210 operational LFG energy projects 714, 716, 717 ore mining process 603–604 organic coagulants 540, 543–544 organic matter (OM) 7, 15, 83, 110, 113, 115, 119, 122, 123, 130, 137, 218–219, 227, 247, 250, 253, 284, 288, 494, 496, 530, 559–564, 569, 578, 718 osmo-heterotrophic unicellular marine protists 167 ozonation 283, 290, 293, 525, 528, 533, 544 ozonation pretreatment 283

p palladium/gold composite material 466 particulate matter formation (PMF) 157 particulates filtration 711 partition-release-recover (PRR) 6 Perionyx excavatus 138 permanent magnetic hydrocyclones 677, 678 Persimmon tannin 477 personal care products (PPCPs) 290 Phaeodactylum tricornutum 180, 181, 185 pH and zeta potential analysis 352–353 phosphate accumulating organisms (PAO) 219, 305, 380, 388 phosphate adsorption processes and mechanisms 341 of ZrO2 @Fe3 O4 357 of ZrO2 @SiO2 @Fe3 O4 357–358 phosphate adsorption equilibrium 351 phosphate adsorption isotherms of ZrO2 @Fe3 O4 nanoparticles 351 of ZrO2 @SiO2 @Fe3 O4 351 of La(OH)3 /Fe3 O4 nanocomposites 353

769

770

Index

phosphate adsorption kinetics of ZrO2 @Fe3 O4 351 of ZrO2 @SiO2 @Fe3 O4 351 of La(OH)3 /Fe3 O4 nanocomposites 353 phosphate recovery and reusability of Fe0 /Fe3 O4 composite 361 of La(OH)3/Fe3 O4 nanocomposites 360 phosphate removal current adsorbents 341–342 and recovery of 342–343 phosphonates 339 phosphorus (P) 13, 305, 339 capture and reuse 437–439 biochar use for 439–442 pollution and eutrophication 340 phosphorus, from wastewater and sludge acidogenic fermentation Al dosage 315–316 experimental methods 312 Fe dosage 312–315 organic carbon 316 membrane bioreactor experiment 317–319 P removal 319–321 P recovery technologies on CEPS 307–308 chemically-enhanced membrane bioreactors 308–309 chemical precipitation 307 thermal treatment 307 WWTP 306–307 P removal and recovery, mechanisms of acidogenic fermentation 330–331 experiment 326–327 Fe speciation 329–330 Fe-P complex, solubility 331–332 microbial iron reduction 331 P speciation 327–329 P removal, from wastewater experimental methods 309 results 310–311

sludge fermentation and P recovery 321–326 acidification vs. acidogenesis 325–326 batch fermentation 321–324 long-term performance 324–325 phosphorus removal capacity 362–363 by Fe0 /Fe3 O4 /Fe2+ system 355–356 kinetics 363–364 and recovery technologies 340–341 photobioreactor (PBR) system 179, 181, 183, 211 physical methods NO3 − and PO4 3− removal 401 physicochemical approach, for nutrient recovery advantages and disadvantages 577 ammonia stripping 570–573 chemical precipitation 574–576 ion exchange and adsorption 573–574 membrane filtration technologies 576–578 physisorption 408, 411, 416, 473 Pit-latrines 121 plant-based approach for wastewater treatment 578–583 plate heat exchanger 640, 644, 646 PO4 3− removal 402 adsorbents for 412 and recovery using Prosopis juliflora weed 421–426 low-cost adsorbents adsorbent dose effect 403–404 contact time 404–405 initial adsorbate concentration 403 interfering anions effect 406–407 pH effect 405–406 temperature effect 404 methods 400

Index

point sources, of nitrate pollution 400 pollution control and monitoring 464, 465 polyaluminum chloride (PACl)-based enhanced primary treatment 308 polycondensation 93 polyethyleneimine modified Lagerstroemia speciosa leaf powder 480 polyethyleneimine modified chitosan fiber (PCSF) 480 polyethylenimine modified bacterial biosorbent (PBBF) 480 Polyhydroxyalkanoates (PHAs) 26, 308 polyhydroxybutyrate (PHB) 324 polymerization/condensation reactions 261 polyphosphate 305, 327, 339, 380 polyphosphate-accumulating organisms (PAOs) 219, 305, 380, 388 polysaccharide polymers 88 polyvinyl chloride (PVC) 464, 740, 743 positive displacement blowers 707 potential Environmental Applications 217–238, 365–366 poultry feed 562 poultry wastewater 561–562 anaerobic digestion, performance of 567 characteristics 561 pre-water treatment 637 precious and critical elements 470, 471 precious metal recovery 463, 472 cyanide leaching 472 from bioderived materials 470, 472 precious metal recovery (PMR) 463, 469, 472, 476, 749 precious metal recycling, importance of 467–469, 481

pressure pipe heat exchangers 643–644 pressure swing adsorption (PSA) process 713 pretreatment technologies 291 evaluation 290–294 primary plastic recycling 742 primary pollutant concentrations 311 primary sludge 275–276, 308, 312, 332 primary wastewater 636–637 printed circuit boards (PCB), leaching and metal recovery process 469 product yield and market value 204–205 Prosopis juliflora weed (PJAC) 421 NO3 − and PO4 3− removal and recovery adsorbate and adsorbent preparation 421 desorption study 425–426 equilibrium adsorption study 421–425 pyro-metallurgy 467–469, 735–736 pyrolysis 253–255, 260, 438, 454, 583 pyrometallurgical process 8, 736 benefits and issues 741 pyrometallurgy 467–469, 735 metal extraction 736

q quaternary plastic recycling 742

r radiolysis process 528 rare earth element (REE) 342 raw domestic sewage 309 raw wastewater 306, 545, 547, 549, 636 Reactive Brilliant Red K-2G (RBR) 238 reactive dyeing mechanism 535, 538–539, 541, 602 reactive dyes 533–535, 538–539, 547

771

772

Index

reduced separation efficiency, of hydrocyclones 666, 667 regression model 38, 703 Renewable Energy Directive (RED) 201 Renewable Transport Fuel Certificates (RTFC) 201 residence time 27–30, 94, 98, 254–255, 258, 260, 583, 671, 673–674, 708, 742 resource recovery 1 backgrounds economical aspect 4–5 environmental impacts 4 HRU 2 population growth 2–4 resource scarcity 4 internationally coordinated framework 10 novel technologies, development of 9 social and economic feasibility 9–10 waste E-waste 9 electronic waste and hazardous waste 7 global food waste 7 industrial waste 7 municipal solid waste 7 wastewater 5 heat recovery 7 nutrient recovery 6 organic carbon recovery 6–7 resource recovery oriented sanitation ecological sanitation 112–113 biofilters 114–115 failures and success 115–116 Rottebehaelter and centrifugal separation sanitation 113–114 human excreta and sustainable future economics, of resource recovery sanitation 127–128 sanitation access and resource recovery 128–129

in rural areas 116–117 anaerobic digestion 119–121 community scale 121 nutrient recovery, from urine 117–119 in urban context energy matters 121–123 industrial scale units 124–125 Johkasou systems 123–124 resource recovery sanitation in developing countries 151–153 anaerobic treatments 154–155 chemical treatments 154 commercial scale resource recovery 155–156 composting 155 pasteurization 154 storage 154 Rhizoclonium hieroglyphicum 475–476 Rhizopus oryzae 476, 739 Rhodosporidium sp. 26 Rhodotorula sp. 26 rice husk (RH) gold sorption 477 rotary lobe blowers 707 rotating cone pyrolysis reactor (RCPR) 254–255 rotating hemispherical filter wastewater collection device 650 rotating particle separator (RPS) 264 rotating sleeve filter wastewater collection device 650 Rottebehaelter and centrifugal separation sanitation 113–114 Rottebehaelter system 112, 113 rubbery polymers 712

s Saccharomyces cerevisiae 24, 739 Salmonella 152, 154 sanitation 11, 109–131, 149, 151–155, 157–160, 397, 719 saturated fatty acids (SFAs) 169, 177, 181, 186

Index

saturation magnetization 345–346, 349–350 scanning electron microscopy (SEM) 220–221, 225, 230, 237, 316, 345, 350, 379–380, 475–477, 608, 610–613, 615, 630–631, 676 scanning electron microscopy/energydispersive X-ray spectroscopy (SEM/EDS) 316 Scenedesmus sp. 170, 174 Schizochytrium sp. 177 secondary plastic recycling 742 secondary wastewater 276, 463, 481, 637 seed sludge 220–221, 312, 318 selective phosphate adsorption, of La(OH)3 /Fe3 O4 nanocomposites 358–360 selective phosphate removal, of Fe0 /Fe3 O4 /Fe2+ system 360, 361 semi-continuous reactors 309, 316 separation efficiency, hydrocyclones 666–667 sewage filter structure 649 sewage sludge characteristics of primary sludge 276 WAS 276 composition 251 sewage treatment plants (STPs) 126, 217, 275, 279–284, 289, 293–294, 491, 530, 636, 639, 643–644 side-stream co-fermentation 306, 333 silicon 16, 91, 601, 605, 615–632, 746–747 silicon activation ball milling effect 617–618 flow chart of 618 mass ratio effect 619 roasting temperature effect 619–620 roasting time effect 620–621

silicon bioleaching 624–632, 738–739, 744 simultaneous saccharification and fermentation (SSF) 24–25 SiO2 @Fe3 O4 nanoparticles, synthesis of 343, 344 sludge microflora 288 sludge volume index (SVI) 220 sludge-based activated carbon dye wastewater, treatment MG sorption 225 mineral acid modification 225–226 heavy metal wastewater, treatment 226 Cu(II) sorption 227–229 heavy metal sorption 226–227 production method 222–223 H3 PO4 223–224 ZnCl2 223 sodium acetate 343 soft reinforcement 71 solar PV panels 745–747 average lifetime 746 generations 746 waste PV panels 746 processing and recycling 746–747 solid residue (SR) 27–28, 30, 87, 90–91, 93–97, 736 solids retention time (SRT) 308–309, 312, 316, 321, 325, 326, 332, 566 soluble chemical oxygen demand (SCOD) 289–290, 309, 311, 313, 315–316, 324 soluble microbial products (SMPs) 233, 321 solventogenesis 24–25 sorption isotherms 232, 234 sorption process 478–479 Freundlich model 480 Langmuir model 480 spent adsorbent handling and management 417 spent LIBs 743–744

773

774

Index

split ratio 16, 663, 666, 667, 670, 672, 674, 676, 679–680, 683–685 spraying type wastewater evaporator 642 Statutory Management Requirements (SMRs) 152 steel slag, in LFG treatment 712 stockpiling 603, 740 struvite formation 1, 150, 307, 380, 499, 501, 574–575, 577, 587 struvite precipitation 6, 119, 128, 374–375, 383, 385–387, 500, 575, 577, 585 SulfaTreat 711 supercritical liquefaction, waste biomass 95 supercritical water gasification 97, 584 sustainability 24 manure management systems LCA 156–157 transitions literature 48–51 swirling decontamination method 651 synthesis gas 258

t Tailor-made food waste 38 tannin gel 476, 739 target transformation (TT) 327 tech devices, average lifespans for 725, 726 techno-economic analysis (TEA) 10, 199–200, 203, 211–212 Teflon-lined stainless-steel autoclave 343, 344 temperature-phased anaerobic digestion (TPAD) 13, 284, 293, 294 tertiary plastic recycling 742 tetraethyl orthosilicate (TEOS) 344 textile dyes chemical structure formula 534 classification 534–535, 536

dyeing process 535, 538 nomenclature of 535, 537 reactive dyeing mechanism 535, 538–539 textile wastewater 525, 539 characteristics 539–540 composition 539 treatment 525 treatment methods 540, 544 advanced oxidation process (AOP) 544 biological methods 544 physico-chemical treatment 540, 542–544 thermal hydrolysis pretreatment (THP) 280, 281 thermal hydrolysis vendor system 281 thermal-chemical treatment, of sewage sludge ashes to value-added materials 261 bio-oils 259–260 biochar 260–261 characterization of 250 combustible gas 258–259 gasification performance-cost-benefit analysis 257 typical gasification processes 255–257 heavy metals removal and recovery 263–264 impact on, environmental sustainability 248–250 incineration performance-cost-benefit analysis 253 typical incineration processes 250–253 liquefaction performance-cost-benefit analysis 258 typical liquefaction processes 257–258 nutrient recovery 261–263

Index

pyrolysis performance-cost-benefit analysis 255 typical pyrolysis processes 253–255 technology limitations deactivation, of catalyst 264 high moisture content 265–266 NOx and SOx emission 265 tar formation 264–265 thermochemical conversion 203, 207, 581, 583–584 thin-film composite (TFC) membrane 386 thiourea ((NH2 )2 CS) leaching 738 Thraustochytrids 167 for biodiesel production 177–178 microalgae, challenges 178–179 total ammonia nitrogen (TAN) concentration 150, 496, 575, 576 total bio-oil (TBO) 95–96 total chemical oxygen demand (TCOD) 309–311 total nitrogen (TN) 5, 6, 141, 220, 260, 262, 309, 545, 563, 565 total phosphorus (TP) 5, 6, 141, 545 removal 220 total primary energy supply (TPES) 200 total static pressure drop, of hydrocyclones 667 total suspended solids (TSS) removal 114, 508 toxic triphenylmethane 224 Trachydiscus minutus 181 trade names, of dyes 535 traditional sewage sludge treatment 247, 250 transesterification, of triglyceride 175–177, 179, 204, 581 transformer oil treatment, using EB 530 transmembrane pressure (TMP) 318, 320 transmission electron microscopy (TEM) 345–346, 348

triacylglycerides (TAGs) 168, 170, 183, 224 triazine and vinyl sulphone dyes, reaction of 539 triclocarban (TCC) 289, 290 trisodium citrate 343 trisodium citrate dihydrate 343 Tungstated zirconia catalyst 176 Turbinaria conoides 475, 476 two parameter isotherm models 409 2000 Waste Strategy, implementation 66–67 financial instruments 70–72 LA implementation, of waste policy 67–69 local authorities and the public 72–74 regional governance 72 strategy, legacy of 74–75 targets 70

u ultrafiltration (UF) 150, 383, 503, 576 United Nation economic commission for Europe (UNECE) 151 United Nations Children’s Emergency Fund (UNICEF) 128 United States environmental protection agency (US EPA) 124, 465, 700, 701, 703–705, 707, 709, 712, 714, 716 unmodified biochars 442, 443, 455 upflow anaerobic sludge blanket (UASB) reactor 6, 117, 120, 219, 233, 565–566 Upper middle-income country (UMIC) 4, 397 Urine Diversion Dehydration Toilets (UDDTs) 116, 126 U.S. Environmental Protection Agency 155, 340

v valuable soil conditioner 110 Ventilated Improved Pit (VIP) 126

775

776

Index

vertical gas collection wells advantages and disadvantages 706 configuration 705 vibrating sample magnetometer (VSM) 345, 347, 350–351 volatile fatty acids (VFA) 7, 24, 97, 279, 308, 496, 566, 577 volatile organic compounds (VOCs) 151, 551, 577, 718

w waste activated sludge (WAS) 275–276, 293 waste electrical and electronic equipment (WEEE) generation 723 metal extraction from 735 recovery options 732 waste management, in England 51–52 waste management, revolution English waste, EU 58–64 in England 51–52 influences, in the UK 64–66 interviewees, selection of 54–58 research design 53, 54 research methods 53 secondary data 58 sustainability transitions literature 48–51 waste printed circuit boards (WPCB) 9, 733 waste-to-energy (WTF) 157–158 wastewater (WW) characteristics of 636–637 collection device, in open type 651 wastewater heat exchangers (WWHE) 639–640 classification of 640–641 flow heat transfer law 657 foulant growth law in 657 soft-foulant separation 657 style of 641–646 wastewater P-recovery processes 305

wastewater source heat pump (WWSHP) 635, 637 application 639–640 case study 654–657 challenges 657 energy source 639–640 operating costs 658–659 principle 637–639 with de-foulant hydrocyclone 651, 652 wastewater treatment 489 ammonium removal methods 489 biolectrochemical systems 490 electricity consumption of 489 energy consumption of 489 wastewater treatment plants (WWTPs) 1, 6–7, 9, 117, 123, 124, 210, 222, 228, 255, 263, 282, 305–306, 308, 374, 380, 437, 454–456, 490, 491, 507, 547, 552, 654–655 wastewater treatment process activated sludge 218–219 granular sludge aerobic granular sludge 220–222 anaerobic granular sludge 219 pilot-scale application 220–222 wastewaters decomposition, by EB technology 530, 531 water contamination 5, 490 water hyacinth-based biochar 454 water scarcity 130, 397 water scrubbing process 711 water scrubbing unit 712 water soluble (WS) 35, 37, 84, 96, 110, 209, 622 water source heat pump (WSHP) 7, 16, 635, 637–640 water-injection hydrocyclones 673 wet air oxidation (WAO) 465 woody biomass 85 World Health Organization (WHO) 128, 233, 397, 400

Index

x

z

X-ray absorption near edge structure (XANES) spectroscopy 13, 327–330, 332–333 X-ray absorption spectrometer (XAS) 327 X-ray diffraction (XRD) 230, 261, 316, 324, 327, 345–346, 349–350, 608, 610–612, 616, 620, 629–630 X-ray photoelectron spectroscopy (XPS) 345 X-ray powder diffraction (XRD) 316, 608

zeolites 96, 119, 138, 142, 186, 438, 573–574, 576 zeta potential analysis 352–354 zirconium oxide (ZrO2 ) 37, 342, 344–346, 352, 357, 364–365 ZrO2 @Fe3 O4 nanoparticles phosphate recovery and reusability 357–358 synthesis of 344 ZrO2 @SiO2 @Fe3 O4 nanoparticles, synthesis of 344

777