Current Developments in Biotechnology and Bioengineering: Smart Solutions for Wastewater: Road-mapping the Transition to Circular Economy 0323999204, 9780323999205

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Current Developments in Biotechnology and Bioengineering Smart Solutions for Wastewater: Road-mapping the Transition to Circular Economy Edited by Giorgio Mannina Engineering Department, Palermo University, Palermo, Italy Ashok Pandey

Centre for Innovation and Translational Research, CSIR-Indian Institute of Toxicology Research, Lucknow, Uttar Pradesh, India; Sustainability Cluster, School of Engineering, University of Petroleum and Energy Studies, Dehradun, Uttarakhand, India; Centre for Energy and Environmental Sustainability, Lucknow, Uttar Pradesh, India

Ranjna Sirohi Department of Food Technology, School of Health Sciences, University of Petroleum and Energy Studies, Dehradun, Uttarakhand, India

Current Developments in Biotechnology and Bioengineering

Series Editor Ashok Pandey Centre for Innovation and Translational Research, CSIRIndian Institute of Toxicology Research, Lucknow, Uttar Pradesh, India; Sustainability Cluster, School of Engineering, University of Petroleum and Energy Studies, Dehradun, Uttarakhand, India; Centre for Energy and Environmental Sustainability, Lucknow, Uttar Pradesh, India

Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, Netherlands The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States Copyright © 2023 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher's permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/ permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. ISBN: 978-0-323-99920-5 For Information on all Elsevier publications visit our website at https://www.elsevier.com/books-and-journals Publisher: Joe Hayton Editorial Project Manager: Helena Beauchamp Production Project Manager: Sujatha Thirugnana Sambandam Cover Designer: Matthew Limbert Typeset by Aptara, New Delhi, India

Contents Contributors  xiii Preface  xvii 1.

Introduction to smart solutions for wastewater: Road-mapping the transition to circular economy

1

Giorgio Mannina, Dario Presti, Ashok Pandey, Herman Helness, Ranjna Sirohi, Jacek Mąkinia

1.1 Introduction

1

1.2  Water-smart solutions to enhance the transition to circular economy 5 1.3  Conclusions and perspectives

7

Acknowledgments 8 References 8

2.

Treatment and disposal of sewage sludge from wastewater in a circular economy perspective

11

Giorgio Mannina, Lorenzo Barbara, Alida Cosenza, Zhiwei Wang

2.1 Introduction

11

2.2  European laws

13

2.3  SS management

17

2.4  SS reuse

20

2.5  Conclusions and perspectives

26

Acknowledgements 26 References 26

3.

Integration of polyhydroxyalkanoates (PHAs) production into urban wastewater treatment plants

31

Dario Presti, María Eugenia Suárez-Ojeda, Giorgio Mannina

3.1 Introduction

31 v

vi  Contents

3.2  PHAs: biobased and biodegradable alternative to plastics

32

3.3  A circular economy approach: PHA production integrated into WWTPs 36 3.4  A detailed view of the independent PEs for PHA production by using MMCs

37

3.5  PHAs extraction from microbial cells (PE4)

45

3.6  Economic sustainability of PHAs production process

50

3.7  Conclusions and perspectives

51

Acknowledgments 52 References 53

4.

Production of volatile fatty acids from sewage sludge fermentation 61 Dario Presti, Bing-Jie Ni, Giorgio Mannina

4.1 Introduction

61

4.2  Biological mechanism and strategies for VFA production from sewage sludge

62

4.3  Trends and innovations in VFA production from sewage sludge

74

4.4  Final applications of sludge-derived VFA and economic evaluation 82 4.5  Conclusions and perspectives

85

Acknowledgments 86 References 86

5.

Zeolites for the nutrient recovery from wastewater

95

Sofia Maria Muscarella, Luigi Badalucco, Vito Armando Laudicina, Giorgio Mannina

5.1 Introduction

95

5.2  Structure and chemical composition of zeolites

96

5.3  Natural zeolites and synthetic zeolites

99

Contents vii

5.4  Applications of zeolites

102

5.5  Use of zeolite for nutrients recovery

104

5.6  Conclusions and perspectives

109

Acknowledgments 110 References 110

6.

Wastewater treatment sludge composting

115

Sofia Maria Muscarella, Luigi Badalucco, Vito Armando Laudicina, Zhiwei Wang, Giorgio Mannina

6.1 Introduction

115

6.2  Legislation about sewage sludge

116

6.3  Sewage sludge composting

122

6.4  Conclusions and perspectives

131

Acknowledgments 132 References 132

7.

Advances in technologies for sewage sludge management 137 Giorgio Mannina, Lorenzo Barbara, Alida Cosenza, Bing-Jie Ni

7.1 Introduction

137

7.2  Technologies in water treatment line

139

7.3  Technologies in sludge treatment line

143

7.4  Evaluation and maturity of technologies for reducing sludge production 149 7.5  Sludge characterization to optimize the dewatering process 150 7.6  Conclusions and perspectives

151

Acknowledgements 151 References 151

viii  Contents

8.

Energy and valuable organic products recovery from anaerobic processes

157

Ewa Zaborowska, Mojtaba Maktabifard, Xiang Li, Xianbao Xu, Jacek Mąkinia

8.1 Introduction

157

8.2  Energy balance in wastewater treatment plants and potential energy recovery

158

8.3  Potential valuable products recovery

161

8.4  Anaerobic processes focused on liquid products recovery

162

8.5  Anaerobic digestion (AD) processes focused on gaseous products recovery

164

8.6  Processes enhancing energy and valuable organic products recovery 169 8.7  Conclusions and perspectives

176

References 176

9.

Life-cycle assessment for resource recovery facilities in the wastewater sector

183

Sofía Estévez, María Teresa Moreira, Gumersindo Feijoo

9.1 Introduction

183

9.2  Life-cycle analysis (LCA) as an environmental impact assessment methodology

185

9.3  Environmental diagnosis of the different alternatives based on the environmental outcomes

207

9.4  Conclusions and perspectives

214

Acknowledgements 215 References 215

10. Water reuse in the frame of circular economy

221

Jiří Wanner, Martin Srb, Ondřej Beneš

10.1 Introduction

221

10.2  Legal framework of water reuse

223

Contents ix

10.3  Worldwide used national water reuse guidelines and regulations 225 10.4  National water reuse guidelines and regulations in selected EU countries

226

10.5  Drivers for water reuse: water resources scarcity and climate change; increasing quality and prize of drinking water; water reuse as a natural part of circular economy

236

10.6  Circular economy and water resources

242

10.7  Barriers of water reuse

242

10.8  Processes of recycled water production from effluents of municipal WWTPS

251

10.9  Examples of successful water reuse projects in Europe

256

10.10  Conclusions and perspectives

261

Acknowledgments 262 References 262

11. Governance factors influencing the scope for circular water solutions 267 Sigrid Damman, Henrik Brynthe Lund, Tuukka Mäkitie, Giorgio Mannina, Gordon Akon-Yamga, Jiří Wanner

11.1 Introduction

267

11.2  Toward a new paradigm

268

11.3  Perceived governance challenges

269

11.4  A multilevel approach

270

11.5  Main drivers and barriers in the studied cases

273

11.6  Contextual interactions and need for new governance perspectives 279 11.7  Conclusions and perspectives

286

Acknowledgments 287 Appendix: List of abbreviations

287

References 288

x  Contents

12. Advances in environmental bioprocess technology for an effective transition to a green circular economy

291

Merve Atasoy

12.1 Introduction

291

12.2  Promising biobased products for resource recovery at WWTPs 293 12.3  Manipulation of microbial community performance for resource recovery

302

12.4  Conclusions and perspectives

305

References 306

13. Advanced technologies for a smart and integrated control of odour emissions from wastewater treatment plant

315

Giuseppina Oliva, Mark Gino Galang, Tiziano Zarra, Vincenzo Belgiorno, Vincenzo Naddeo

13.1 Introduction

315

13.2  Full-scale smart solutions for odour control with treatment and abatement solutions

317

13.3  Implementation of smart technologies

323

13.4  Conclusions and perspectives

325

References 326

14. Microbial biotechnology for wastewater treatment into circular economy 333 Giuseppe Gallo, Walter Arancio, Emilia Palazzotto, Fanny Claire Capri, Rosa Alduina

14.1 Introduction

333

14.2 Metagenomics

337

14.3 Metatranscriptomics

343

14.4 Metaproteomics

344

14.5  Resource recovery and energy production by microbial communities: from WWTPs to biorefineries

345

Contents xi

14.6  Conclusions and perspectives

349

References 350

15. Biological nutrient recovery from wastewater for circular economy 355 Shihai Deng, Huu Hao Ngo, Wenshan Guo, Na You, Shuai Peng

15.1 Introduction

355

15.2  Anaerobic processes for nutrients recovery

357

15.3  Photo-bioprocesses for nutrients recovery

369

15.4  Microbial electrochemical technologies for nutrients recovery

390

15.5  Conclusions and perspectives

396

References 397

16. Stakeholder engagement: A strategy to support the transition toward circular economy business models

413

Antonino Galati, Nino Adamashvili

16.1 Introduction

413

16.2  From a linear to a circular model for greater sustainable development 415 16.3  Stakeholder engagement and management of sustainable business models

417

16.4  How do stakeholders and their engagement affect the transition toward circular business models?

420

16.5  Conclusions and perspectives

424

References 425

Index  431

Contributors Nino Adamashvili  Department of Economics, University of Foggia, Foggia, FG, Italy Gordon Akon-Yamga  Council for Scientific and Industrial Research (CSIR), Accra, Ghana Rosa Alduina  Dipartimento di Scienze e Tecnologie Biologiche Chimiche e Farmaceutiche, Università di Palermo, Palermo, Italy Walter Arancio  Advanced Data Analysis Group, Fondazione Ri.MED, Palermo, Italy Merve Atasoy  Department of Chemical Engineering, KTH Royal Institute of Technology, Stockholm, Sweden; UNLOCK, Wageningen University and Research, Wageningen, the Netherlands Luigi Badalucco  Department of Agricultural, Food and Forest Sciences, University of Palermo, Palermo, Italy Lorenzo Barbara  Engineering Department, Palermo University, Palermo, Italy Vincenzo Belgiorno  Sanitary Environmental Engineering Division (SEED), Department of Civil Engineering, University of Salerno, Fisciano, SA, Italy Ondřej Beneš  VEOLIA ČESKÁ REPUBLIKA, Prague, Czech Republic Fanny Claire Capri  Dipartimento di Scienze e Tecnologie Biologiche Chimiche e Farmaceutiche, Università di Palermo, Palermo, Italy Alida Cosenza  Engineering Department, Palermo University, Palermo, Italy Sigrid Damman  SINTEF Digital, Trondheim, Norway Shihai Deng  School of Human Settlements and Civil Engineering, Xi’an Jiaotong University, Xi’an, China Sofía Estévez  Department of Chemical Engineering, CRETUS. Universidade de Santiago de Compostela, Santiago de Compostela, Spain María Eugenia Suárez-Ojeda  GENOCOV Research Group, Department of Chemical, Biological and Environmental Engineering, School of Engineering, Universitat Autònoma de Barcelona, Escola d’Enginyeria, Barcelona, Spain Gumersindo Feijoo  Department of Chemical Engineering, CRETUS. Universidade de Santiago de Compostela, Santiago de Compostela, Spain

xiii

xiv  Contributors

Mark Gino Galang  Sanitary Environmental Engineering Division (SEED), Department of Civil Engineering, University of Salerno, Fisciano, SA, Italy Antonino Galati  Department of Agricultural, Food and Forest Sciences, University of Palermo, Palermo, PA, Italy Giuseppe Gallo  Dipartimento di Scienze e Tecnologie Biologiche Chimiche e Farmaceutiche, Università di Palermo, Palermo, Italy Wenshan Guo  Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, Australia Herman Helness  SINTEF Digital, Trondheim, Norway Vito Armando Laudicina  Department of Agricultural, Food and Forest Sciences, University of Palermo, Palermo, Italy Xiang Li  College of Environmental Science and Engineering, Donghua University, Shanghai, China; Shanghai Institute of Pollution Control and Ecological Security, Shanghai, China Henrik Brynthe Lund  SINTEF Digital, Trondheim, Norway Jacek Mąkinia  Faculty of Civil and Environmental Engineering, Gdańsk University of Technology, Gdańsk, Poland Mojtaba Maktabifard  Faculty of Civil and Environmental Engineering, Gdańsk University of Technology, Gdańsk, Poland Giorgio Mannina  Engineering Department, Palermo University, Palermo, Italy María Teresa Moreira  Department of Chemical Engineering, CRETUS. Universidade de Santiago de Compostela, Santiago de Compostela, Spain Sofia Maria Muscarella  Department of Agricultural, Food and Forest Sciences, University of Palermo, Palermo, Italy Tuukka Mäkitie  SINTEF Digital, Trondheim, Norway Vincenzo Naddeo  Sanitary Environmental Engineering Division (SEED), Department of Civil Engineering, University of Salerno, Fisciano, SA, Italy Huu Hao Ngo  Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, Australia Bing-Jie Ni  Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia Giuseppina Oliva  Sanitary Environmental Engineering Division (SEED), Department of Civil Engineering, University of Salerno, Fisciano, SA, Italy Emilia Palazzotto  Dipartimento di Scienze per la Promozione della Salute e Materno infantile “G. D’Alessandro”, Università di Palermo, Via del Vespro, Italy

Contributors xv

Ashok Pandey  Centre for Innovation and Translational Research, CSIR-Indian Institute of Toxicology Research, Lucknow, Uttar Pradesh, India; Sustainability Cluster, School of Engineering, University of Petroleum and Energy Studies, Dehradun, Uttarakhand, India; Centre for Energy and Environmental Sustainability, Lucknow, Uttar Pradesh, India Shuai Peng  College of Environmental Science and Engineering, Tongji University, Shanghai, Singapore Dario Presti  Engineering Department, Palermo University, Palermo, Italy Ranjna Sirohi  Department of Food Technology, School of Health Sciences, University of Petroleum and Energy Studies, Dehradun, Uttarakhand, India Martin Srb  Prague Water Supply and Sewerage Company, Prague, Czech Republic Zhiwei Wang  Tongji University, Shanghai Institute of Pollution Control and Ecological Security, State Key Laboratory of Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shanghai, China Jiří Wanner  Department of Water Technology and Environmental Engineering, Faculty of Environmental Technology, University of Chemistry and Technology, Prague, Czech Republic Xianbao Xu  College of Environmental Science and Engineering, Donghua University, Shanghai, China; Shanghai Institute of Pollution Control and Ecological Security, Shanghai, China Na You  Department of Civil and Environmental Engineering, National University of Singapore, Singapore, Singapore Ewa Zaborowska  Faculty of Civil and Environmental Engineering, Gdańsk University of Technology, Gdańsk, Poland Tiziano Zarra  Sanitary Environmental Engineering Division (SEED), Department of Civil Engineering, University of Salerno, Fisciano, SA, Italy

Preface The book entitled “Smart Solutions for Wastewater: Road-mapping the Transition to Circular Economy” is a part of the comprehensive series on Current Developments in Biotechnology and Bioengineering (Editor-in-Chief: Ashok Pandey). In recent decades, technological solutions to recover water, energy, fertilizers, and other products from wastewater have been proposed. Drivers for this work range from low resource recovery potential and cost effectiveness, to the high energy demands and large environmental footprints of current treatment-plant designs. However, only a few technologies have been implemented and a shift from wastewater treatment plants toward water resource recovery facilities still seems far away. The book is aimed at fostering an effective transition of wastewater management strategies from a linear to a circular economy approach. Specifically, the goal is to propose practical tools to support the sustainable design, operation, and optimization of wastewater treatment plants to foster resource recovery while minimizing the overall environmental footprint and costs. The shift from a linear to a circular economy scheme is hampered by barriers which are not only technological but also of organizational, regulatory, social, and economic character. This book provides a holistic overview of the current status of the recent progress in resource recovery from wastewater treatment, including the fundamental aspects and development of processes, as well as the needs of further research and industrialization. The book also gives special emphasis and discussion on green technologies for resource recovery from wastewater and bottlenecks regarding their applications. As a future green bioprocess, biogas production and waste minimization, opportunities, future perspectives, and research needs are also discussed. The book provides evidence about how the wastewater can be turned into a resource. With the transition from linear to circular economy, wastewater treatment plants need to be converted into resource recovery facilities. Therefore, there is an urgent need to provide comprehensive review regarding the knowledge and technical advances on the latest development of resource recovery technologies from wastewater treatment. Thus, this book provides state-of-the-art information and the perspectives for the future developments and gathers contributions of the EU project “Achieving wider uptake of water-smart solutions—WIDER UPTAKE.” WiderUptake has the aim to provide measures to overcome the existing barriers hampering the application of circular economy concepts to the wastewater sector through the creation of a roadmap for fostering new industrial symbioses based on water-smart solutions that will link water and wastewater treatment, resource extraction, energy supply and product development for the agriculture, building & manufacturing materials industries, and energy supply. The book aims to have an impact in the water and wastewater sector. The expected impact will rely on providing a roadmap for fostering new industrial symbioses based on water-smart solutions that will link water and wastewater treatment, resource extraction, energy supply and product development for the agriculture, building & manufacturing materials industries, and energy supply. It is expected that the book provides measures to overcome the existing barriers (social, economic, and political) hampering the application of circular economy concepts to the water and wastewater sector. The book discusses how it is possible to achieve the recovery of resources from water and wastewater in symbiotic circular economy relationships between utilities and industries.

xvii

xviii  Preface

The book will be a potential resource for environmental and chemical engineers, scientists, educators, students, and general public to understand the current developments and future prospects in the field of resource recovery from wastewater. It provides an excellent, concise, interdisciplinary, and updated overview of technologies in terms of potential recovered yields, pollutants removal, nutrients recovery, and energy production, as well as the achievement of energy efficiency of the process itself. We would like to express our deepest appreciation to the authors and reviewers who have made valuable contributions to this book. We thank Elsevier team for their consistent hard work in the publication of this book.

Editors Giorgio Mannina Ashok Pandey Ranjna Sirohi

1 Introduction to smart solutions for wastewater: Road-mapping the transition to circular economy Giorgio Manninaa, Dario Prestia, Ashok Pandeyb,c,d, Herman Helnesse, Ranjna Sirohif, Jacek Mąkiniag a ENGINEERING DEPARTMENT, PALERMO UNIVERSITY, PALERMO, ITALY bCENTRE FOR INNOVATION AND TRANSLATIONAL RESEARCH, CSIR-INDIAN INSTITUTE OF TOXICOLOGY RESEARCH, LUCKNOW, UTTAR PRADESH, INDIA cSUSTAINABILITY CLUSTER, SCHOOL OF ENGINEERING, UNIVERSITY OF PETROLEUM AND ENERGY STUDIES, DEHRADUN, UTTARAKHAND, INDIA dCENTRE FOR ENERGY AND ENVIRONMENTAL SUSTAINABILITY, LUCKNOW, UTTAR PRADESH, INDIA eSINTEF DIGITAL, TRONDHEIM, NORWAY fDEPARTMENT OF FOOD TECHNOLOGY, SCHOOL OF HEALTH SCIENCES, UNIVERSITY OF PETROLEUM AND ENERGY STUDIES, DEHRADUN, UTTARAKHAND, INDIA gFACULTY OF CIVIL AND ENVIRONMENTAL ENGINEERING, GDAŃSK UNIVERSITY OF TECHNOLOGY, GDAŃSK, POLAND

1.1 Introduction From its origins in the late 19th century, the aim of wastewater treatment has been focused on removal of contaminants in order to protect public health and water environment. In response to those concerns, more stringent regulations have gradually been enforced and adequate methods of wastewater treatment have been developed. However, today scarcity of resources and environmental challenges are driving major global changes also in the wastewater industry. Sustainability issues in wastewater treatment plants (WWTPs), including the energy efficiency, greenhouse gas (GHG) emissions, usage of chemicals, carbon footprint and fate of emerging pollutants, have highly been prioritized in wastewater management [1]. Moreover, wastewater treatment becomes a critical component of a new circular economy (CE) concept. The current production system, based on raw materials extraction and industrial transformation into products, has long-term sustainability issues due to the use of nonrenewable resources. The linear economy that has been employed in our society for years is not possible anymore. Indeed, the “take-make/use-dispose” model resulted in scarcity of resources and degradation of the environment, because it does not take into account the environmental sustainability of the technological innovations. In order to overcome these issues, the CE concept has to be adopted to face the current situation, decoupling economic growth from environmental degradation, and thus thriving respectfully and sustainably with the environment [2,3]. To promote the CE concept among water professionals, a “Water and circular economy” white paper has been released [4], focusing on such issues as common characteristics, ideas, Current Developments in Biotechnology and Bioengineering. DOI: https://doi.org/10.1016/B978-0-323-99920-5.00015-9 Copyright © 2023 Elsevier Inc. All rights reserved.

1

2  Current Developments in Biotechnology and Bioengineering

and approaches between CE principles and sustainable water management in urban areas. The International Water Association (IWA) prepared a framework “Water Utility Pathways in a Circular Economy” [5], which identified CE opportunities within three interrelated pathways (water, material, energy) with the central role of a WWTP in each pathway. CE concept demands that wastes are acknowledged (from industrial and agricultural sources to those derived from direct human consumption) as a renewable resource from which water, materials (e.g., nutrients, metals, biopolymers), and energy can be recovered [6]. A sustainable transformation in the global production system is needed. It has to become an auto-regenerative process in which the waste, generated in technical and biological cycles, is converted into raw matter [7]. In such a scenario, wastewater treatment is becoming a key platform to pave the way for technological development that focuses on the transformation of global production systems from a linear to a circular economic model. Indeed, more than 50% of lost waste resources are embedded in wastewater, and its treatment is established worldwide with a long technological history [7–9]. The application of the CE approach to the wastewater treatment sector will reduce the energy and resource consumption of traditional wastewater treatment plants (WWTPs). Further, CE will allow for production of energy and materials recovered from wastewater, thus valorizing residues from waste to raw feedstock for the production of a variety of new products, closing loops and boosting the implementation of CE [10–12]. Coupling the traditional wastewater treatment technologies, which are focused on the degradation of contaminants embedded in wastewater to suitable effluent concentrations, with the new water-smart paradigm focusing on resource recovery leads to view WWTPs as Water Resource Recovery Facilities (WRRFs) [13–15]. In the United States, the Water Environment Federation formally adopted the term WRRF in place of WWTP to reflect the changing paradigm in the wastewater sector. A WRRF (also referred to as “water resource factory” or “wastewater biorefinery”) is assumed to apply innovative technologies to operate at the intersection of wastewater treatment and bioprocessing [13]. Indeed, it mostly utilizes biological agents to extract valuable products, while treating wastewater at the same time [16,17]. This conceptual transformation will lead to the next generation of WWTP that will target energy neutrality and the recovery of nutrients and other materials, thus allowing for realization of the CE concept in wastewater treatment. The closure of the cycle and the valorization of waste resources will also reduce the economic impact of wastewater treatment on communities. The challenge is to slip away from the paradigm of merely operating conventional WWTPs as endof-pipe environmental protection infrastructures and start perceiving wastewater as a resource requiring different levels of management in order to successfully reintroduce resources to the market for social consumption, thus transforming wastewater management utilities in market actors extracting raw materials rather than just managing treatment plant operations for costeffective pollutant removal [13]. This also requires a link and cooperation with other sectors, such as agriculture, energy, and construction [18]. Despite the urgent need to secure natural water resources, the transformation of WWTPs in WRRFs is still limited [19]. The barriers that hinder this transition are not only technological

Chapter 1 • Introduction to smart solutions for wastewater  3

but also of economic, social and educational, regulatory, administrative, and logistic/organizational character [12,17,20]. New and better models for intersectoral collaboration are needed, to avoid the situation where regulations, policies, and incentives in other sectors (e.g., agriculture, health, energy) may prohibit or hinder development of industry based on wastewater resources. There is therefore the need to further develop and demonstrate water-smart solutions, which are not only technological but may also encompass other types of innovation, such as innovative governance and stakeholder engagement, business models and educational programs targeting both citizens and wastewater professionals. In order to overcome the barriers hindering the transition to the WRRF paradigm, both technical innovations and non-technical innovations are needed, therefore a robust roadmap for the transition should be established (Fig. 1.1). The development of the CE concept requires support by research and innovation and indeed the CE has received a special attention in the European Union funding program “Horizon 2020”. Recently completed and ongoing projects, for example, NextGen, Project Ô, Hydrousa, SMART-Plant, Water-Mining, B-WaterSmart, WIDER UPTAKE, REwaise, and Ultimate address specific technical solutions related to water, energy and materials recovery, and the framework conditions for their success. Among them, the project “Achieving wider uptake of water-smart solutions—WIDER UPTAKE” has the aim to provide measures to overcome the existing barriers hampering the application of circular economy concepts to the wastewater sector trough the creation of a roadmap for fostering new industrial symbioses based on water-smart solutions that will link water and wastewater treatment, resource extraction, energy supply and product development for the agriculture, building and manufacturing materials industries, and energy

Technological barriers

Social barriers

ap for transition Roadm Administrative barriers

Economic barriers Educational barriers Regulatory barriers

Wastewater

Logistic/ organizational barriers

Wastewater

Linerar economy:

Circular economy:

WWTP

WRRF Energy

Fertilizers

Treated water

Sewage sludge

Others Bioplastics

Water for reuse

FIG. 1.1  Roadmap and barriers for transition from the concept of wastewater treatment plant to the paradigm of water resource recovery facility.

4  Current Developments in Biotechnology and Bioengineering

GOVERNANCE

TECHNICAL PERFORMANCE

Water-Smart solutions

ECONOMY

Water-Smart society

ENVIRONMENT

SOCIAL FIG. 1.2  Water smart solutions according to WiderUptake EU project.

supply [21]. It is expected that this roadmap will provide measures and recommendations to achieve optimized water-smart systems enhancing the barriers overcome (Fig. 1.2). Bearing in mind the considerations above, the aim of this book is to promote an effective transition of wastewater management strategies from a linear to a circular economy model. The goal is to propose practical tools to support the sustainable design, operation, and optimization of wastewater treatment plants to foster resource recovery in symbiotic CE relationships between utilities and industries, while minimizing the overall environmental footprint and costs. In the following chapters, a holistic overview of the current status of the recent progress in resource recovery from wastewater treatment, including the fundamental aspects, development of processes and bottlenecks regarding their applications, as well as the needs of further research and industrialization is provided. In Chapters 2 and 7, the fundamental aspects of sewage sludge production from wastewater treatment, the advances in technologies for sludge management and the new trends are discussed, focusing on the CE perspectives. In Chapters 3 and 4, volatile fatty acids and polyhydroxyalkanoates production from sewage sludge and the related treatment schemes are analyzed focusing on advances and perspectives. In Chapters 5 and 15, nutrients recovery from wastewater and advances in zeolite medium for nitrogen adsorption and desorption are presented in view of resource recovery. In Chapters 6 and 8 anaerobic digestion processes and sewage sludge composting in wastewater sector are discussed in view of an optimization for CE applications. In Chapter 9, the main advances in life cycle assessment and application to resource recovery facilities in the wastewater sector are reported. In Chapter 10, water reuse from wastewater treatment is discussed reporting advantages, disadvantages and the limits that hamper a wider application. In Chapters 11 and 16, the main drivers and bottlenecks for the transition from linear to

Chapter 1 • Introduction to smart solutions for wastewater  5

circular economy model are highlighted, and governance models for engaging stakeholders for fostering CE in the water sector are presented. In Chapters 12 and 14, advances in bioprocess technology and microbial biotechnology in wastewater treatment for an effective transition to a green CE are discussed. In Chapter 13, smart water systems to minimize odor emissions from wastewater treatment plants are presented.

1.2  Water-smart solutions to enhance the transition to circular economy According to Water Europe, the European Technology Platform for water that represents the whole value-chain of water and aims to achieve a European Water-Smart Society, the definition of Water-Smart Society is: “a society in which the true value of water is recognised and realised, and all available water sources are managed in such a way that water scarcity and pollution of groundwater are avoided. Water and resource loops are largely closed to foster a circular economy and optimal resource efficiency, while the water system is resilient against the impact of climate change events” [22]. In 2016 vision document, Water Europe defined four new concepts that are fundamental to drive the transition toward a water-smart society, namely [22]: • “Multiple Waters” dealing with the need to increase the reuse and recycling of water, as well as the use of alternative water resources, such as brackish water, saline water, rainwater, etc., in addition to groundwater and surface water. • “Digital Water” referring to the use of capillary networks and sensors, meters, modelling, and analysis of the water system all the way along to the individual user, as such generating large amounts of valuable data (big data) for innovative decision support and governance systems. • “Value in Water” concerns capturing and exploiting the potential value in wastewater. This includes using it for energy generation, recovering resources (e.g., nutrients, metals, bioplastics) embedded in the wastewater stream, and developing business models for resource recovery and regeneration. • “Hybrid Grey and Green Infrastructure” as a combination of grey engineered infrastructure, green engineered infrastructure and natural systems, part of the water system that will be used for water extraction, treatment, distribution, reuse, and resilience against the effects of climate change. In a follow-up document [18], Water Europe emphasized that WWTPs should implement a holistic management considering water reuse, energy savings, and exploitation of energy embedded in wastewater. Such an approach meets the EU Green Deal objectives and leads to carbon-neutral wastewater management. In order to achieve a Water-Smart Society, in the WWTP of the future, namely a WRRF, other added value products such as nutrients, metals, cellulose fibers, proteins, enzymes, biofuels, industrial chemicals (like carboxylic acids, alcohols and hydrogen) biopolymers, and the water itself, need to be recovered and reintroduced to the market for social consumption, thus closing the loop and boosting the implementation of CE [10,11]. In this perspective, the

6  Current Developments in Biotechnology and Bioengineering

adoption of Water-Smart solutions is essential: innovative water technologies, digital tools and economic, governance and business models that contribute to solving water-related challenges are needed [17]. Similarly, to Water Europe, the IWA defined three new concepts to inspire the international water sector to adopt a smarter approach to water management. Indeed, in a recent white paper on Digital Water, presenting the perspectives that need to be pursued to achieve a smarter water management, it is stated that water should be [23]: • Smart by design: adaptive ‘off-grid’, distributed systems that provide diversity, and modularity, characteristics critical for resiliency; • Smart use: combining concepts of water fit for purpose (different grades for different uses), and resource recovery and reuse (of water, energy, and nutrients from wastewater); • Smart (digital) control: IoT supporting data-driven models that can help integrate and optimize smart pumps, valves, sensors and actuators, and enabling each device to “talk” to each other and send real-time information to be accessed and shared via the cloud. Despite the different terminology used by different institutions to describe the transformation currently taking place in the water sector, it is important to define the water-smart solutions needed to achieve the common recognized objectives of the water sector transformation: lowenergy and resource recovery, resilience against the effects of climate change and exploitation of the value of data through the adoption of digital technologies, automation, and artificial intelligence [22,23]. As far the authors are aware, there is not a univocal definition in the field of wastewater treatment for “water-smart solutions.” Terminologies, such as “smart,” “intelligent” and other analogous ones, have been attributed to different kind of technologies and practices linked to the water and wastewater sector. In a recent book “Smart Water Utilities, complexity made simple,” Ingildsen and Olsson [24] envisage the “smartening up” of water management systems through a combination of control and process technologies. They define the emerging paradigm of Smart Water utility as a utility that “ensures a systematic and intelligent decision-making process at all levels, based on online water quality and quantity sensors, taking into account the full water cycle from water intake to water effluent, with the aim of ensuring adequate water quality and quantity, with a minimum consumption of energy and materials” [24]. In the water supply sector, the attribute “smart” has often been used in relation to “smart water metering,” “water smart grids,” and “smart water quality monitoring systems.” All these applications use information and communication technology (ICT) tools for sensing and system monitoring, communication, data storing and analysis to improve the management of water supply systems and ensure water conservation by tempestive identification of leakage and water quality alterations, thus reducing water loss and prevent any health hazards [25–29]. Similarly, in the wastewater sector, the attribute “smart” is often used referring to tools that allow wastewater utilities to achieve low-cost wastewater treatment, reuse, and resource recovery [30–33]. Such tools comprise online measurements, analytics platform and control and automation supported by Internet of Things (IoT) and technologies providing real-time, actionable insights with predictive capabilities and operational intelligence, These technologies, used

Chapter 1 • Introduction to smart solutions for wastewater  7

to accomplish the needs of smart wastewater management, are embedded in the emerging framework of cyber-physical systems, that uses digital tools (such as sensors) based on information and communication technologies to control physical systems [34,35]. “Water-smart solutions,” therefore, still lack a univocal and clear definition, however, bearing in mind the considerations above, it is clear that such solutions should contribute to multiple objectives in order to achieve a water-smart society. Within the H2020 project WIDER UPTAKE, water-smart solutions are assessed according to technical, environmental, social, economic and governance aspects needed to allow the transition from a linear to a circular economy model in the water sector. All tools, technologies and practices discussed in the following chapters of this book, range from innovative technologies for wastewater resource recovery, to those for sludge production minimization and greenhouse gases emission reduction, as well as innovative business and governance models can be regarded as solutions to achieve a water-smart society. Further, tools presented in this book are aimed to support the sustainable design, operation, and optimization of wastewater treatment plants to foster resource recovery in symbiotic CE relationships between utilities and industries, while minimizing the overall environmental footprint and costs

1.3  Conclusions and perspectives The transition from a linear to a circular economy model in the wastewater sector involves the widespread application of water-smart solutions. This transition is needed to achieve a watersmart society but this is a challenging and long-term process. Indeed, there are still several barriers (economic, technological, social, and educational, regulatory, administrative, and logistic/organizational) that need to be addressed in the complex process of the transition from the traditional concept of WWTP to the new paradigm of WRRF. The role of water-management utilities will be crucial in this change of the paradigm. Indeed, they should go from just managing treatment plant operations for cost-effective pollutant removal to be real market actors extracting raw materials to be reintroduced to the market, and thus closing the loop. In order to achieve this transformation, in the near future, the newly designed and existing WWTPs should be equipped with advanced water-smart technologies that are able to reduce the amount of energy and resources to be used, allow recovery of resources (such as nutrients, biopolymers, fertilizers, etc.) and energy, while at the same time improve the effluent quality for water reuse and reduce the amount of waste (e.g., sewage sludge) produced. In order to overcome the existing barriers and allow water management utilities and policy makers to shift toward more sustainable and resource recovery-oriented decisions, the creation of a robust roadmap is an essential step. It should be an innovative tool based on a decisionsupport system, able to guide managers, operators, designer, and researchers in the water sector to foster the transition toward a circular economy. The roadmap should address not only wastewater technicalities, but should include innovative governance, business and stakeholder engagement models and encourage actions aimed at boosting the social acceptability of using wastewater recovered resources through public information and education.

8  Current Developments in Biotechnology and Bioengineering

Acknowledgments This work was funded by the project “Achieving wider uptake of water-smart solutions—WIDER UPTAKE” (grant agreement number: 869283) financed by the European Union’s Horizon 2020 Research and Innovation Programme, in which the first author of this chapter, Giorgio Mannina, is the principal investigator for the University of Palermo. The Unipa project website can be found at: https://wideruptake.unipa.it/.

References [1] Metcalf & Eddy Inc. Wastewater engineering: Treatment and reuse. New York: McGraw-Hill Professional; 2014. [2] Maina S, Kachrimanidou V, Koutinas A. A roadmap towards a circular and sustainable bioeconomy through waste valorization. Curr Opin Green Sustain Chem 2017;8:18–23. [3] Ghisellini P, Cialani C, Ulgiati S. A review on circular economy: the expected transition to a balanced interplay of environmental and economic systems. J Cleaner Prod 2016;114:11–32. [4] Antea Group. Water and circular economy. A White Paper ver. 1.2. AnteaGroup, the Ellen MacArthur Foundation and ARUP; 2019. [5] IWA. Water utility pathways in a circular economy. International Water Association; 2016. Available from: www.iwa-connect.org. [6] Norouzi M, Chàfer M, Cabeza LF, Jiménez L, Boer D. Circular economy in the building and construction sector: a scientific evolution analysis. J Building Eng 2021;44:102704. [7] Puyol D, Batstone DJ, Hülsen T, Astals S, Peces M, Krömer JO. Resource recovery from wastewater by biological technologies: opportunities, challenges, and prospects. Front Microbiol 2017;7(Jan):1–23. [8] Guest JS, Skerlos SJ, Barnard JL, Beck MB, Daigger GT, Hilger H, et al. A new planning and design paradigm to achieve sustainable resource recovery from wastewater. Environ Sci Technol 2009;43(16):6126–30. [9] Van Loosdrecht MCM, Brdjanovic D. Anticipating the next century of wastewater treatment. Science 2014;344(6191):1452–3. [10] Nielsen PH. Microbial biotechnology and circular economy in wastewater treatment. Microb Biotechnol 2017;10(5):1102–5. [11] Nizami AS, Rehan M, Waqas M, Naqvi M, Ouda OKM, Shahzad K, et al. Waste biorefineries: enabling circular economies in developing countries. Bioresour Technol 2017;241:1101–17. [12] Neczaj E, Grosser A. Circular economy in wastewater treatment plant: challenges and barriers. Proc AMIA Annu Fall Symp 2018;2(11):614. [13] Kehrein P, Van Loosdrecht M, Osseweijer P, Garfí M, Dewulf J, Posada J. A critical review of resource recovery from municipal wastewater treatment plants-market supply potentials, technologies and bottlenecks. Environ Sci Water Res Technol 2020;6(4):877–910. [14] Regmi P, Maere T, Stewart H, Amerlinck Y, Samstag R, Rieger L, et al. The future of WRRF modelling: outlook and challenges. Water Sci Technol 2018;79(1):1–12. [15] Pott R, Johnstone-Robertson M, Verster B, Rumjeet S, Nkadimeng L, Raper T, et al. Wastewater biorefineries: integrating water treatment and value recovery. In: Filho WL, Surroop D, editors. The nexus: Energy, environment and climate change. Green Energy and Technology. Cham: Springer; 2018. p. 289–302. [16] Verster B, Minnaar S, Cohen B. Introducing the wastewater biorefinery concept: A scoping study of polyglutamic acid production from a Bacillus-rich mixed culture using municipal wastewater. Water Research Commission Report TT587/13, Pretoria; 2014. [17] Coats ER, Wilson PI. Toward nucleating the concept of the water resource recovery facility (WRRF): perspective from the principal actors. Environ Sci Technol 2017;51(8):4158–64.

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[18] Water Europe. For a green, circular & smart urban wastewater treatment directive (Position paper); 2020. [19] Papa M, Foladori P, Guglielmi L, Bertanza G. How far are we from closing the loop of sewage resource recovery? A real picture of municipal wastewater treatment plants in Italy. J Environ Manage 2017;198:9–15. [20] Alcalde Sanza L, Gawlik BM. Water reuse in Europe: relevant guidelines, needs for and barriers to innovation. JRC Sci Policy Rep 2014:1–18. [21] Mannina G, Badalucco L, Barbara L, Cosenza A, Di Trapani D, Gallo G, et al. Enhancing a transition to a circular economy in the water sector: the EU project wider uptake. Water (Switzerland) 2021;13(7):1–18. [22] Water Europe. Water Europe Water Vision 2030; 2017. Available from: https://watereurope.eu/wp-content/ uploads/2020/04/WE-Water-Vision-english_online.pdf. [23] Sarni W, White C, Webb R, Cross K, Glotzbach R. Digital Water: Industry leaders chart the transformation journey. London, IWA Publishing; 2020. [December]. p. 165–85. [24] Ingildsen P, Olsson G. Smart water utilities: Complexity made simple. London: IWA Publishing; 2016. [25] Hajebi S, Song H, Barrett S, Clarke A, Clarke S. Towards a reference model for water smart grid. Int J Adv Eng Sci Technol 2013;2(3):310–17. [26] Britton TC, Stewart RA, O’Halloran KR. Smart metering: enabler for rapid and effective post meter leakage identification and water loss management. J Cleaner Prod 2013;54:166–76. [27] Cominola A, Giuliani M, Piga D, Castelletti A, Rizzoli AE. Benefits and challenges of using smart meters for advancing residential water demand modeling and management: a review. Environ Modell Softw 2015;72:198–214. [28] Boyle T, Giurco D, Mukheibir P, Liu A, Moy C, White S, et al. Intelligent metering for urban water: a review. Water 2013;5(3):1052–1081. [29] Prasad AN, Mamun KA, Islam FR, Haqva H. Smart water quality monitoring system. 2015 2nd Asia-Pacific World Congress on Computer Science and Engineering: APWC on CSE; 2015. p. 1–6. [30] Salam, A. Internet of Things in water management and treatment BT: Internet of Things for sustainable community development: Wireless communications, sensing, and systems. in: A. Salam (Ed.), Springer International Publishing, Cham. 2020;273–98. [31] Martínez R, Vela N, el Aatik A, Murray E, Roche P, Navarro JM. On the use of an IoT integrated system for water quality monitoring and management in wastewater treatment plants. Water 2020;12(4):1096. [32] Boulos PF, Walker AT. Fixing the future of wastewater systems with smart water network modeling. J Am Water Works Assoc 2015;107(4):72–80. [33] Dong J, Wang G, Yan H, Xu J, Zhang X. A survey of smart water quality monitoring system. Environ Sci Pollut Res Int 2015;22(7):4893–906. [34] Radini S, Marinelli E, Akyol Ç, Eusebi AL, Vasilaki V, Mancini A, et al. Urban water-energy-food-climate nexus in integrated wastewater and reuse systems: Cyber-physical framework and innovations. Appl Energy 2021;298:117268. [35] Sun C, Puig V, Cembrano G. Real-time control of urban water cycle under cyber-physical systems framework. Water 2020;12(2):406.

2 Treatment and disposal of sewage sludge from wastewater in a circular economy perspective Giorgio Manninaa, Lorenzo Barbaraa, Alida Cosenzaa, Zhiwei Wangb a ENGI NEERI NG DEPARTM ENT, PA L E R MO U N I V E R S I T Y, PA L E R MO , I TA LY b T O N G JI UNIVERSIT Y, S HANGHAI I NS TI TUTE O F P O L L U T I O N C O N T R O L A N D E C O L O G I C A L S E C U R I T Y, STAT E KEY L ABO RATO RY O F PO L L UT I O N C O N T R O L A N D R E S O U R C E R E U S E , S C H O O L O F ENVI RO NM ENTA L S C I E N C E A N D E N G I N E E R I N G , S H A N G H A I , C H I N A

2.1 Introduction Climate change and environmental degradation represent threats to Europe and the whole world. The recent European Green Deal aims to turn Europe into a “climate-neutral” continent by 2050 [1]. With this regard, the European Green Deal proposes several actions to support the transition towards a sustainable and competitive economy in terms of exploitation of environmental resources: (1) invest in environment-friendly technologies, (2) support industry in innovation, (3) introduce cleaner and cheaper forms of private and public transport, and (4) improve the energy sector and collaborate with international partners to improve global environmental standards. The current European trend provides for a transition towards the concept of circular economy. Unlike the linear economy, which considers resources exploitation with the final inevitable waste production, the circular economy provides the reintroduction of waste materials within the economic cycle. Consequently, the material life cycle is extended, thus minimizing the waste production [1]. The new European “Circular Economy Action Plan” announces measures (such as product design, promotion of circular economy processes, promotion of sustainable consumption) to ensure that the resources used are kept in the European economy as long as possible [2]. The water sector and its modernization fall within the objectives previously mentioned of the recent European Directives. With this regard, the wastewater treatment plants (WWTPs) could have a key role for the resource recovery towards the circular economy [2,3]. Indeed, during the last decade the historical aim of WWTPs management to provide effluent water quality under law constraints has been widened towards a more sustainable perspective [4]. WWTPs can be considered as a biorefinery from which recover valuable resources, reduce the energy consumption, minimize the impact of emerging pollutants, thus contributing to the European sustainable goals [5]. Current Developments in Biotechnology and Bioengineering. DOI: https://doi.org/10.1016/B978-0-323-99920-5.00011-1 Copyright © 2023 Elsevier Inc. All rights reserved.

11

12  Current Developments in Biotechnology and Bioengineering

Specifically, the wastewater has to be recognized as a resource rather than a waste stream from which recover water, sewage sludge (SS) and nutrients (nitrogen (N) and phosphorus (P)) (Fig. 2.1) [6]. The recovered SS can further be used for land applications (fertilizer and/or soil amendment for agricultural land), energy and construction materials production (Fig. 2.1) [7]. In this chapter the attention is mainly focused on the SS fate in a sustainable perspective. The SS is a complex mixture of biodegradable organic matter (present in percentage from 59% to 88%), inorganic matter (nitrogen (N), phosphorus (P), potassium (K), calcium (Ca), sulfur (S), magnesium (Mg)), cellular matter (bacteria, protozoa, pathogens and their products) and pollutants (heavy metals, carcinogenic substances) having a huge energetic potential [8]. Treated SS is becoming increasingly recognized as a multifunctional resource, creating marketable use opportunities [9]. SS can find various applications, as a substrate for fertilization and soil remediation, as an energy resource to produce heat and electricity and as an additive for building materials [9] (Fig. 2.1). Therefore, the SS reuse allows to face the environmental and economic problems dealing with the sludge management such as the land occupation and the high disposal costs [10]. The SS management represents approximately 50% of the total WWTPs operating costs [11]. Moreover, the sludge disposal entails a considerable environmental impact due to the soil occupation and the risk of surface and groundwater pollution [12]. Thus, the final SS landfill disposal should be avoided as much as possible [12]. Despite the potential of SS, currently several WWTPs in Europe are not able to produce sludge suitable for the final reuse [13]. This is mainly due to the inappropriate applied technology that does not guarantee to meet the current new regulations for sludge reuse [13]. Although academia has proposed several technological solutions to recover material (such as phosphorus, water or sludge to be reused) to establish a circular economy in WWTPs, their full scale applications are still very limited since investments are required [14]. This circumstance coupled with the increase of sludge production (mainly due to the more stringent limit for treated wastewater and the increase of the amount of treated wastewater) makes the SS a bottleneck in Sludge reuse

Urban area

Wastewater treatment plant

Water

Sewage sludge

Nutrient

Land applications

Energy

Construction materials

FIG. 2.1  Resources recovery from wastewater treatment plants and sewage sludge fate as recovered resource.

Chapter 2 • Treatment and disposal of sewage sludge from wastewater  13

the WWTPs operation [15–17]. Therefore, the adoption of a virtuous process aimed to establish recovery resource oriented decisions (from plant designers, managers, policy makers, etc.) is mandatory to support the sludge reuse [18–23]. Within this context, the role of policy makers, WWTP managers and population is crucial. Policy makers have to focus all their efforts in making the resource recovery and their reuse feasible and applicable. WWTPs managers have to change their paradigm in operating plants. Managers have to become actors of the resource recovery process by applying all the available technologies and strategies to trigger a market of the recovered resources [14]. Population has to be properly informed in view of trusting on the value of the recovered resources in a circular economy perspective. This chapter aims to provide an overview of the issues surrounding the SS and its reuse. Specifically, regulatory aspects, potential and drawbacks of SS reuse will be discussed.

2.2  European laws Environmental protection and public health are priority themes in the European Union. However, the current fragmentary disciplines sometimes limit the sustainable development to which European Union focuses on mainly in case these disciplines are not supported by the economic and technological investments [14]. In addition, each Member State has issued a series of laws which included more stringent limit values and introduced more restrictions [22]. In this section an overview on the existing European directives on the SS and wastewater treatments field will be provided. Furthermore, an analysis on the Directive 86/278/EEC revision is also performed.

2.2.1  European directives on SS and wastewater treatments The European legislation contains several Directives regulating the wastewater and SS management. In Table 2.1 the key European Directives are summarized with respect to their application sector and main objective.

Table 2.1  Directives and regulations on sewage sludge. Directive/Regulation

Application sector

Objectives

Directive 86/278/EEC [24] Directive 91/271/EEC [16]

Agriculture and soil application Environment protection

Directive 91/676/EEC [25]

Environment protection

Directive 2008/98/EC [28] and Directive 2018/851/EC [20] EU regulation 2019/1009 [29]

Waste management

Regulate the agriculture sludge application. Environment protection from the negative effects of wastewater discharge. Reducing water pollution caused or induced by nitrates from agricultural sources and preventing further such pollution. Environment protection and human health by preventing or reducing the waste generation. Regulate fertilizer production and re-evaluate sludge stabilization processes considering current risks.

Fertilizing products

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Directive 86/278/EEC [24] known as Sewage Sludge Directive (SSD) encourages the sludge adoption in agriculture, regulating its use to avoid harmful effects on soil, vegetation, animals and humans. According to the SSD [24] the untreated sludge cannot be used for agricultural purposes (fertilizer and/or soil amendment). The Directive 86/278/EEC [24] imposes that SS can be used (after treatment) according to the plant’s nutritional needs and avoiding risk for soil, surface water and groundwater quality. Directive 86/278/EEC [24] imposes the limit values for: (1) heavy metal concentrations in sludge for use in agriculture, (2) heavy metal concentrations in soil, (3) amount of heavy metal which may be added annually to agricultural land based on 10-year average. Based on the Directive 86/278/EEC [24] member states have introduced regulations with more restrictive limits also including limits for other compounds not included in the Directive. Specifically, the European states member have chosen two different approaches to apply the SSD at national level: adopt more stringent requirements (Austria, Belgium, Croatia, Czech Republic, Denmark, Finland, France, Germany, Hungary, Lithuania, Luxembourg, Malta, Netherlands, Poland, Romania, Slovenia and Sweden) or follow similar requirements (Bulgaria, Cyprus, Estonia, Greece, Ireland, Italy, Latvia, Portugal, Slovakia, Spain and United Kingdom) [22]. Regarding the limitations for heavy metals in sludge, Directive 86/278/EEC [24] imposes the following limits: cadmium (Cd) 20–40 mg/kgDM, copper (Cu) 1000–1750 mg/kgDM, mercury (Hg) 16–25 mg/kgDM, nickel (Ni) 300–400 mg/kgDM, lead (Pb) 750–1200 mg/kgDM and zinc (Zn) 2500–4000 mg/kgDM. In addition to these values imposed by the SSD, almost all member countries (excepting for Cyprus and United Kingdom) impose limits on chromium (Cr) with values ranging between 200 mg/kgDM and 1500 mg/kgDM. Belgium, The Netherlands, Denmark, Finland, Austria, Italy, United Kingdom have also imposed limit values for the arsenic (As) with values ranging between 15 mg/kgDM and 150 mg/kgDM. Although SSD does not provide any limits or requirements for organic trace pollutants in sludge, some national regulations on the use of sludge have added these limits. It is worth noting that the most inspected organic compounds in different states member are the polychlorinated biphenyls (PCB) (the limit is 0.4 mg/kg dry matter in Sweden, and 0.2 mg/kg dry matter in Low Austria and Germany), the adsorbable organic halogens (AOX) (the limit is 500 mg/kg dry matter in Low Austria and Germany), and the polycyclic aromatic hydrocarbons (PAH) (the limit is 3 mg/kg dry matter in Denmark and Sweden). The SSD does not impose limits on the content of pathogens in sludge, but many European countries impose controls on the content of salmonella (exception for Lithuania, Luxembourg and Slovakia) and other pathogens, for which the limit values are different. For example, France and Italy imposes the limit for Salmonella at 8 MPN/10 g of dry matter and 1000 MPN/g of dry matter, respectively. The SSD also prescribes that the treated sludge must contain a significant amount of organic matter, N and P to be considered a good fertilizer. Many of the member states prescribe analyzes for agronomic parameters. Furthermore, Latvia and Sweden give indications on the maximum annual rate of N and P that can be spread on soils, while only Italy imposes limit values for application on soils. Specifically, Italy imposes the following limits: organic matter >20% of dry matter (DM), total N >1.5% of DM, and total P >0.4% of DM. Despite the Directive 86/278 / EEC does not impose limits on the maximum quantity of sludge that can be spread on the soil, some member states have therefore adopted an annual limit. Specifically, Denmark allows spreading

Chapter 2 • Treatment and disposal of sewage sludge from wastewater  15

10 MgDM/ha year, with stricter requirements on sludge quality than European guidelines. Ireland, on the other hand, limits sludge spreading to 2MgDM/ha year, but the pollutant limit values remain unchanged compared to the SSD. The Directive 91/271/EEC [16] deals with wastewater treatment. In particular, the Directive 91/271/EEC imposes limits on the pollutants concentration in the treated wastewater. Since these limits are sometimes very stringent, the application of Directive 91/271/EEC [16] has led to an increase in the amount of sludge produced in Europe. The Directive 91/676/EEC [25], or Nitrates Directive establishes the maximum acceptable annual addition of total N in the soil (250 kgN/ha year). This value represents the limiting factor determining the biosolids application rate as fertilizers. The maximum acceptable annual addition of total N reduces to 175 kgN/ha year for the nitrate vulnerable zones. For the materials for which the nitrogen content is low (e.g., sludge compost, dewatered sludge cake) the maximum acceptable addition of total N is 500 kgN/ha every two years [26]. Currently, soils in different European areas (especially in the Mediterranean area) suffer from nutrients depletion [27]. Therefore, the application of biosolids rich in nutrients could provide an opportunity to solve this problem. Directive 2008/98/EC [28], updated by Directive 2018/851/EC [20], introduces the waste hierarchies and the end-of-waste criteria. Specifically, the end-of-waste criteria establish when a waste ceases to be waste and obtains a status of a product [28]. Directives 2008/98/EC [28] and 2018/851/EC [20] lay down measures to protect the environment and human health by preventing or reducing the waste generation and the adverse impacts of waste by promoting the transition from a linear to circular economy. Regarding the SS, Directives 2008/98/EC [28] and 2018/851/EC [20] support the adoption of prevention techniques (i.e., technologies aimed at reducing the amount of produced SS), promote the sludge reuse (for land applications or energy production). Despite the aforementioned Directives push towards the sludge reuse, in the recent European regulation 2019/1009 [29] on fertilizing products, SS is excluded from the categories of component material category (CMC) for compost (CMC3) and digestates (CMC5). Thus, introducing a new barrier to the development of the SS management in a more sustainable perspective.

2.2.2  Revision of the SSD In 2020 the European Commission decided to revise the Directive 86/278/EEC, creating a shared initiative to assess the risks and opportunities of using sludge in agriculture. Specifically, between 16th June and 25th August 2020, citizens and stakeholders were invited to give feedback on the roadmap created for the revision of the Directive 86/278/EEC. In order to provide here an overview of the key needs in the context of SS management and reuse, public data have been analyzed. Specifically, from the discussions raised by the authors feedback, 14 main issues type have been identified according to the key issues reported in the scientific literature [30–33]: (1) directive revision approval (DRV), (2) pollutant beware (PB), (3) favorable to agriculture sludge reuse (FASR), (4) environmental protection (EP), (5) nutrient recovery (NR), (6) circular Economy (CE), (7) coordination of SSD with other directives (CSSD), (8) hierarchy for SS reuse (HR), (9) pollution prevention at source (PPS), (10) review of end-of-waste policies (REWP),

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100% DRA = Directive revision approval PB = Pollutant beware FASR = Favorable to agriculture sludge reuse EP = Environmental protection NR = Nutrient recovery CE = Circular Economy CSSD = Coordination of SSD with other directives

90%

Percentage of response

80% 70%

HR = Hierarchy for SS reuse PPS = Pollution prevention at source REWP = Review of end-of-waste policies CEU = Coordination with all EU countries LEP = Local economy promotion SI = Sludge incineration USSU = Unfavorable to SS use

60% 50% 40% 30% 20% 10% 0%

DRA

PB FASR EP

NR

CE CSSD HR PPS REWP CEU LEP Issue type

SI USSU

FIG. 2.2  Percentage of response for each issue type.

(11) coordination with all EU countries (CEU), (12) local economy promotion (LEP), (13) sludge incineration (SI), and (14) unfavorable to SS use (USSU). In Fig. 2.2 the result of the analysis is shown. By analyzing data reported in Fig. 2.2 one can observe that: • The current Directive must be revised in view of satisfying the existing need to recover resources in a sustainable and circular economy perspective. • Most feedback underlines the need to encourage agricultural sludge reuse. With this regard, the water management utilities perspective needs to be changed. The WWTP has to become a market actor producing resources to insert in the market and favoring their circularity [14]. • The application of new technologies at full scale must be considered. This can allow the creation of sludge hierarchy reuse depending on specific contexts. • The community and social aspects require to be included in the sludge circular economy issues. Indeed, the poor social acceptance towards the resources recovered from wastewater may represent a barrier for triggering and spreading the circular economy in the water sector. Therefore, population must be opportunely informed about the potential, risks and environmental beneficial effects of the recovered resources. • Some feedback suggests to increase the control policies to identify the responsible of the illegal discharges into the WWTPs. Indeed, illegal discharges (for example, of inorganic pollutants) may seriously compromise the quality of the recovered resources (and obviously of sludge).

Chapter 2 • Treatment and disposal of sewage sludge from wastewater  17

• A coordination among the various directives (SSD, Urban Wastewater Treatment Directive, etc.) is required to avoid legislative contradictions. • People agree on improving communication among the European Member States, which in some cases have completely different policies on sludge recovery in order to avoid legislative fragmentation. From 20th November 2015, citizens and stakeholders are invited to express their opinion on the revision of the directive in a public consultation carried out via a questionnaire available online. The questionnaire document consists of two parts. The first introduces the reasons that led to the creation of the public consultation, with specific references to the SSD and to the European environmental protection objectives, such as the new action plan for the circular economy; following is a guidance on the questionnaire. The second part of the document is the actual questionnaire, comprising 20 questions divided into six parts. The questions range from general aspects, which anyone can answer, to more specific questions (such as technological solutions to implement for improving circular economy) which only experts in the field can answer [34]. After the deadline set for 5th March 2021, the European Commission, having received the feedback from the previous consultations, will finalize the proposal to revise the directive [35].

2.3  SS management SS management represents a critical issue for WWTPs managers and governments. The increase of sludge production is due to the growth of wastewater to be treated and to stringent effluent quality limits. This circumstance leads to an increase in the WWTPs operating costs. In this section some issues related to the sludge management will be summarized and discussed: (1) increase in the SS production, (2) SS disposal, and (3) environmental impact of sludge management.

2.3.1  Increase in the SS production In 2020 the production of SS in Europe was around 13 million tons of dry matter [17], an increase compared to the 9.7 million tons of dry matter produced in 2015 [36]. The increase in SS production is linked to the increase of the civil wastewater discharged to the WWTPs. Indeed, during the last 20 years the amount of wastewater discharged to the WWTPs has gradually increased. Before the year 2000, more than 80% of the population in 16 of the European member states, was already connected to a WWTP [37]. In 2018, the population connected to a WWTP was more than 95% in Germany, Latvia, Luxembourg, Netherlands, Austria, Sweden, Switzerland and United Kingdom. In the other countries, such as Cyprus, Malta, Romania, Croatia, Iceland, Albania, Serbia, Bosnia and Herzegovina and Kosovo this percentage reduces to 50% [37]. From 2000 to 2018, several countries considerably increased the population connected to WWTPs, for example, Hungary (from 29.8% to 80.4%) and Slovenia (from 12.3% to 68.9%) [37]. Furthermore, as mentioned above, the stringent quality limits of the treated wastewater have also contributed to the increase in sludge production. Indeed, in some cases the

18  Current Developments in Biotechnology and Bioengineering

adoption of advanced biological/chemical/physical processes (such as temperature-phased anaerobic digestion ozonation, and lysis-thickening) had the consequence of increase the amount of SS produced. In Europe, the largest producers of SS are Germany (1.8 million Mg of dry matter, 2016), Spain (1.2 million Mg, 2016), and France (1.2 million Mg, 2017) [36].

2.3.2  SS disposal and costs The nature and composition of SS, rich in nutrients, but also very often with high concentrations of pollutants such as heavy metals, has led the European countries to adopt different solutions for its disposal. According to Eurostat 2020 data [33] SS disposal can be performed by means of four paths: agricultural use (treated SS), compost production, incineration and landfill disposal. In particular, in Spain (2016), Ireland (2017) and Norway, more than half of the produced SS is used as fertilizer for agricultural use (80%, 79% and 59% of the total produced quantity, respectively). Alternatively, the SS can be used to produce compost; this solution is adopted in several states such as Estonia (84%, 2016), Hungary (72%), Cyprus (58%), Luxembourg (53%, 2017) and Slovakia (46%). In order to reduce the spread of pollutants in soils, some Countries adopt different disposal solutions, such as incineration or landfill. Incineration is the method most chosen by many countries, not only to avoid landfill disposal which would pose a significant environmental problem (such as leachate formation that require to be further treated, space occupation), but also for energy recovery through waste-to-energy plants; in this regard, the Countries that prefer this methodology are Switzerland (100%), Netherlands (87%), Belgium (82%), Germany (65%), Austria (54%) and Greece (32%). Landfill is instead used as the main disposal route in Malta (100%), Bosnia and Herzegovina (100%) and Romania (52%) [37]. The choice of the SS disposal influences the SS management cost. In Table 2.2 are summarized the costs in Europe related to the different types of disposal, including the transport [11,36,38]. Approximately 50% of the total SS management cost is due to the transport [11]. Indeed, the SS produced inside the WWTP is transported outside the plant to reach the sites for disposal; in Europe the transport costs vary from € 50 to € 160 MgDM−1. For the reuse of sludge in agriculture the disposal cost varies from € 150 to € 400 MgDM−1. Higher prices are for landfill disposal, from € 200 to € 600 MgDM−1, and for composting, from € 250 to € 600 MgDM−1. The highest disposal costs are related to the SS incineration, ranging between € 450 and € 800 MgDM−1. Considering that landfill disposal and incineration do not allow for credit recovery from the

Table 2.2  Costs in Europe for each type of disposal [17]. Type of disposal

min

Transport Agriculture Compost Incineration Landfill

50 150 250 450 200

Disposal costs (€ MgDM−1)

max 160 400 600 800 600

Chapter 2 • Treatment and disposal of sewage sludge from wastewater  19

reuse of resources, except for incineration in waste-to-energy plants, the most economically convenient alternatives for the SS disposal are agriculture and composting.

2.3.3  Environmental impact The environmental impact of the SS management depends on technologies adopted during its stabilization process and on its final disposal. Greenhouse gases (GHG) production represents a critical issue in the SS management. In particular, methane (CH4) emissions are strictly related to the processes occurring in the sludge line of a WWTP. About 72% of CH4 emitted from a WWTP derives from the sludge treatment and digestion units, while the rest are attributable to the CH4 dissolved in the wastewater. Regarding carbon dioxide (CO2) emission, its production is due the biological treatment processes occurring both in the water and sludge lines of the WWTP (aerobic metabolism, endogenous respiration) [39]. Tarpani et al. [40] have assessed the environmental impact due to the SS treatment and disposal processes by using the ReCiPe and USEtox methods [41,42]. Tarpani et al. have assessed the GHG emission (in terms of equivalent CO2, kgCO2/1000kgDM) for different technologies applied during the SS management process (e.g., anaerobic digestion, incineration, agricultural use, etc.). Tarpani et al. have found that the anaerobic digestion represents the SS treatment technology having the lowest impact in terms of GHG emissions. Indeed, authors have found that the GHG emissions of the anaerobic digestion is negative (−174 kg CO2 eq./1000kg DM) due to the energy production. Incineration is the second option with a lower environmental impact, producing 39.8 kg CO2 eq./1000kg DM. This value is given by the emissions produced by the fuel oil used in the incinerator, which are counterbalanced by energy recovery to produce electricity and heat. Wet air oxidation represents the third option the lowest impact in terms of GHG emissions, producing 86.4 kg of CO2 eq./1000kg DM, followed by composting, with 197 kg of CO2 eq./1000kg DM produced. In both cases, the environmental impact is due to the use of electricity to power the processes. Pyrolysis, on the other hand, is the worst option analyzed by Tarpani et al. producing 315 kg CO2 eq./1000kg DM caused by the use of natural gas for drying and consumption of electricity. Heavy metals represent the main problem for sludge reuse in agriculture. The anaerobically digested sludge constitutes a much higher average ecotoxicity of fresh water than composted sludge. Tarpani et al. found that Zn (95% of the total) and Ni are the main heavy metals that can be found in sludge [40]. Hg, Pb, Cr, and Zn are the most dangerous metal elements that have the greatest impact on the population [43]. Hg, Pb, and Cr are the metal carcinogens that pose the greatest risk to humans [44]. Excess Zn may cause taste alterations [44]. The data on heavy metals confirm doubts about the application potential of sludge in agriculture [45]. However, as there is no clear assessment of the life cycle of heavy metals [44]. Tarpani et al. suggest that future studies could focus more on assessing the actual risk that metals can cause to the environment [40]. Other environmental impacts are attributed to the presence of newly discovered microplastics [46] and pharmaceutical products [47]. Sun et al. [46] have assessed that the microplastics are removed during treatments within a WWTP and can be found in the sludge 2 × 106 particle/ day, corresponding to an average annual efflux of 5 × 107m3/year. Therefore, the application of

20  Current Developments in Biotechnology and Bioengineering

the sludge on soils would lead to an increase in the diffusion of microplastics in the environment. The combustion processes are suitable to avoid the release of microplastics into the environment. Pharmaceutical products such as bisphenol A, triclosan, triclocarban and the antibiotics ciprofloxacin, ofloxacin and norfloxacin can be found at concentration higher than 1 mg kg−1 [47]. Lombardi et al. [48] assert that the composting process of SS can be useful in limiting the environmental impact when applied to the soil. Zhao et al. [49] have assessed that the use of composted sludge can lead to environmental risks related to eutrophication and acidification of ecosystems due to the release of nitrogen and the diffusion of heavy metals. Therefore, before applying the composted sludge on the soils, it is necessary to know the amount of nutrients and pollutants that can be absorbed by the plants and spread into the soil and water [50].

2.4  SS reuse The SS recovery paths are: (1) land applications, (2) energy recovery, and (3) manufacturing of construction materials. These three possible uses will be discussed and analyzed with respect to the existing literature on this topic. With this regard, in Table 2.3 the studies found in literature are summarized based on the type of use and the treatment method or the adopted technology.

2.4.1  Land applications Land application involves the dispersion, spraying, injection or incorporation of the SS above or below the soil surface (USEPA, 1995). In addition to treated sludge, derived materials such as composted and pelletized SS can be used in these applications. As mentioned in Section 2.2, in some European countries SS reuse as a soil improver or fertilizer in cultivated fields is the most common method of disposal [36]. SS must be treated through processes that include biological (aerobic or anaerobic digestion), thermal and/or chemical stabilization before use for land applications [51]. Indeed, the pathogen stability and reduction degree are the basic requirements for applications on soils [52,53]. Land applications of SS can represent an interesting strategy for increasing land productivity by increasing fertility and soil organic matter content. In addition, SS can also improve the physical properties of soil, especially when used on soils with heavy texture and poor structure [54]. Castán et al. [54] conducted a study with compost of different origins (sewage sludge, organic solid waste, manure) applied to sandy soil. The compost produced from the sludge would lead to an increase of carbon (C) and N in the soil, while it showed a lower P release than other compost and a continuous decrease over time in the field, despite this type of compost has a P content higher. The use of sludge allows a greater availability of nutrients for plants, thus increasing agricultural productivity. Petersen et al. [55], by applying anaerobic digested sludge to an oat crop, have found the increase of N and P quantity in the treated soils thus allowing the improvement of the crop yield. Positive effects on forest productivity were found with the application of sludge on forest soils. In fact, an increase in tree growth has been found, with minimal negative effects on the ecosystem. Wang et al. [56] applied the sludge to a Pinus radiata plantation for several years. The application was conducted by using two types of sludge with different N levels in two

Chapter 2 • Treatment and disposal of sewage sludge from wastewater  21

Table 2.3  Main type of sludge reuse and treatment methods found in literature. Reference

Type of use

Castán et al. [54]

Land applications

Petersen et al. [55]

Land applications

Wang et al. [56]

Land applications

Valdecantos et al. [57] Land applications

Ojeda et al. [58]

Land applications

Li et al. [59]

Land applications

Devlin et al. [62]

Energy recovery

Aryal et al. [63]

Energy recovery

Martín et al. [64]

Energy recovery

Houtmeyers et al. [65] Energy recovery

Nielsen et al. [66]

Energy recovery

Lucena et al. [71]

Construction materials

Arulrajah et al. [72]

Construction materials

Ukwatta et al. [68]

Construction materials

Valderrama et al. [75]

Construction materials

Treatment method or adopted technology

Main results 2

SS compost applied on 6 m plot of land at 40Mg/ha. 20 months of test. Anaerobic digested sludge applied at 120 kg/ha on oat crop. SS applied to a 1000 ha of Pinus radiata silviculture for 6 years. SS applied to Mediterranean degraded drylands with Pinus halepensis. SS applied at 10 Mg/ha to a burnt Mediterranean forest area. SS applied at 4.5 kg/m2 in an abandoned mine site. Acid or alkaline sludge pre-treatment to improve biomethane extraction. Water scrubber pretreatment to extract biomethane. Sonication pretreatment at 100–400 W for 5–6 min. Microwaves pretreatment 800–1250 W at 2.45 GHz for 1–6 min. Thermal and alkaline-thermal pretreatment

Increment on soil: – 100% for C (from 3 to 6 g/kg). – 125% for N (from 0.2 to 0.45 g/kg). – 140% increase of N on soil (from 25 to 60 kg/ha). – Increase in crop yield. 40% increase in tree volume.

Restoration of soil basal area from 15% to 300%. – Vegetation cover up to 58%. – Soil erosion decreased to 26% for the sandy soil and 32% for the loam soil. Increase in herbage grew by 18 times compared with negative control. 54–88% increase in methane generated. Biomethane produced with 97.55% in CH4 content. – 77% increase in biogas production. – 95% increase in methane yield. 20–207% increase in biogas yield

Increase in volatile solid fraction: – 27% (80°C) – 57% (170°C) – 74% (170°C and KOH) Using 10% of sewage sludge The use of SS mixed with soil proved by weight in pavement base to be good stabilizers for mechanical layers. stresses of pavement base layers for roads and highways. Geotechnical tests to evaluate Possibility of using sludge as a filling sewage sludge usage material for road embankments. as a fill material in road Indispensable addition of additives to embankments. improve geotechnical properties. Made of clay–biosolids bricks Compressive strength reduced by with 25% by weight of 28–55% compared to traditional firedsludge clay bricks. Sludge mixed with lime in a Stabilization of sludge with reduction screw-based plug reactor of organic matter and water content by 10% compared to traditional treatments.

22  Current Developments in Biotechnology and Bioengineering

different areas, comparing them with an untreated area. The data collected showed a notable increase in the diameter and volume of trees. In particular, 40% of the increase was noted in the area treated with low N content sludge, and 46% in the area with high N content, compared to the untreated control. Although a slight reduction in wood density was noted, this is counterbalanced by the large increase in wood volume. Even for sites highly degraded the use of SS benefiting reforestation. Valdecantos et al. [57] studied the application of sludge on arid Mediterranean soils that have been degraded due to fires and intensive grazing. Three types of applications were carried out with air-dried sludge, fresh sludge and composted sludge. Due to the arid climate, the application of sludge has improved the fertility of the soil, with improvements from 15% to 300%. Valdecantos et al. [57] obtained the highest soil fertility improvement (300%) by adopting composted sludge. Composted sludge, fresh sludge, and thermally dried sludge were also used in the study by Ojeda et al. [58] for applications in a Mediterranean forest area degraded by a fire. In addition to recording an increase in soil fertility, the application of all three types of sludge has led to an improvement in the conditions of the soil, with a decrease in erosion and runoff compared to untreated controls. Runoff decreased by 32% for loam soils and by 26% for sandy soils Another opportunity for SS reuse is the rehabilitation of a mine site to restore contaminated soil for future agricultural use. For this purpose, Li et al. [59] studied the application of SS and nitrogen fertilizers by evaluating the growth of native herbaceous species. The application of the sludge, after a hundred days of growth, led to a growth of the herbaceous species 18 times higher than the untreated control, with higher yields than nitrogen fertilizers. In this way it is possible to increase the total population of microorganisms, organic matter, total nitrogen, available nitrogen, phosphorus, and potassium, increasing soil fertility. The application of SS in this case is also useful for preventing soil erosion. Improved soil structure, reduced bulk density, increased soil porosity, increased soil moisture retention and hydraulic conductivity rate are the main advantages associated with the application of SS on soil [30].

2.4.2  Energy recovery Anaerobic digestion is a widely used technology that allows the use of sludge with a high moisture content to produce biogas (mixture of 60–70% of CH4 and 30–40% of CO2), with a calorific value of about 30 MJ/Nm3 [60]. The biogas is used directly to produce electricity through thermal generators and internal combustion engines. Electricity generation is very often done inside the WWTPs. In this way, a plant can reduce the consumption of energy from the electricity distribution grid for the management of the various processes, or even in some cases, when production exceeds needs, the energy surpluses are resold to the electricity companies [61]. Alternatively, the biogas can be refined to produce biomethane. To improve the extraction process, Devlin et al. [62] studied an acid pre-treatment of the sludge using HCl at various dosages. The decrease in pH, from 6 to 1, led to an increase in methane yield compared to nonpretreated sludge and a reduction of about 8 days in the digestion process. With an alkaline pre-treatment, using KOH and NaOH with increasing pH up to values ​​of 10–12, it is possible to increase the methane fraction up to 88% [62]. Indeed, since acid and alkaline pre-treatments speed hydrolysis step the global digestion time is reduced (reduction of retention time) with

Chapter 2 • Treatment and disposal of sewage sludge from wastewater  23

the increasing in the volatile solid destruction and consequently in methane yield. Aryal et al. [63] conducted a study using water scrubber to extract biomethane with a 97.55% CH4 content. In this way the calorific value of the extracted gas is higher than that of the biogas, with values​​ of 51.31 MJ/Nm3. Martín et al. [64] used sonication pre-treatment for sludge using batch reactors at 100–400 W for 5–60 min, improving the solubility of organic compounds and achieving a 77% increase in biogas and an increase in methane yield up to 95%. Houtmeyers et al. [65] used microwaves of 800–1250 W at 2.45 GHz for 1–6 min as a pre-treatment for sludge. There was an increase in the biodegradability of the sludge by 47–50%, an increase in soluble organic fractions of 117% and in the biogas yield of 20–207%. Nielsen et al. [66] have used thermal pre-treatments. Compared to traditional pre-treatments, at 80°C the volatile solid increased by 2–27% after each treatment step and at 170°C by 44–57%. With the addition of KOH at 170°C, this type of thermochemical pre-treatment led to an increase of 68–74% of the volatile solid. These pre-treatments led to an increase of the CH4 fraction in the biogas of 20%, 28%, and 45% respectively compared to the untreated controls. Another method for extracting energy from sludge concerns the combustion process. An incinerator, which is normally used to reduce the volume of sludge and elements harmful to the environment, can be converted into a waste-to-energy plant. The heat extracted from the combustion process at temperatures between 800°C and 1150°C, is used to heat water which can be used as a direct source of heating or into a steam turbine to generate electricity. Unlike anaerobic digestion, which can directly use dewatered sludge, combustion requires further drying of dewatered sludge to reduce the moisture content to C14) chain-length PHAs chain length PHAs depending on the number of carbon atoms in the monomers [23,24]. Their physicochemical and mechanical properties range from brittle and stiff plastics to those of elastomeric materials depending on PHAs chemical structure and, given the existence of a large number of different types of monomers, different homopolymers or copolymers can be synthetized, making possible to obtain tunable properties [25]. However, the short-chain-length poly(3-hydroxybutyrate) (P(3HB)) homopolymer and poly(3-hydroxybutyrate-co-3-hydroxyvalerate) (P(3HB-co-3HV)) copolymer are the PHAs most commonly synthesized by MMCs. These PHAs are generally semi-crystalline thermoplastic polymers, with physical and mechanical characteristics similar to those of other widely used fossil-fuel based plastics, such as propylene (PP) and low-density polyethylene (LDPE) [26]. A comparison of the properties of these PHAs with conventional petrochemical based plastics such as PP, LDPE and polystyrene is shown in Table 3.1. Poly(3-hydroxybutyrate) (P(3HB)) has some thermal and mechanical properties comparable to, for example, PP, therefore it has a high melting point and good tensile strength. However, P(3HB) is not very interesting from the application point of view, due to its stiffness and Table 3.1  Comparison of the properties of specific PHAs and commonly used synthetic polymers. P(2HB): poly(3-hydroxybutyrate), P(3HB-co-XX mol%3HV): poly(3-hydroxybutyrate-co-XX% mol 3-hydroxyvalerate).

Polymer

Melting Glass transition Young’s Tensile Elongation temperature temperature modulus strength at break (°C) (°C) (GPa) (MPa) (%) References

P(3HB) (average properties) P(3HB-co-8 mol% 3HV) P(3HB-co-20 mol% 3HV) P(3HB-co-50 mol% 3HV) Polypropylene (PP) Low-density polyethylene (LDPE) Polystyrene (PS)

176 161 145 162 176 130 240

5 – –1 0 –10 –30 100

2.9 0.34 0.8 0.41 1.7 0.2 3.1

37 13.2 20 13.4 38 10 50

4–5 – 50 230 400 620 –

[21] [27] [25] [28] [25] [25] [22]

34  Current Developments in Biotechnology and Bioengineering

brittleness (Young’s number above unity and low elongation at break); this is due to the formation of large crystals during a re-crystallization that occurs over time, even at room temperature [29]. On the contrary, copolymers properties of 3-hydroxybutyrate (3HB) and 3-hydroxyvalerate (3HV) (reported in Table 3.1) have lower melting temperatures than that of P(3HB) which improves their processing, furthermore the elongation at break of the copolymers is significantly higher than that of the P(3HB), as well as their elasticity. Given the wide range of properties, PHAs are a class of biopolymers that are well suited to multiple applications. Among the most useful characteristics of PHAs, biodegradability, biocompatibility, and low permeability to water vapor, make them very interesting for various sectors such as packaging, medical and agricultural. The first field of use of PHAs was in the area of packaging (shampoo bottles, containers for cosmetics) and in the food sector (film for covering cartons, milk bottles) [30]. PHAs are used extensively in the medical field, due to the non-toxicity of its biodegradation compounds, in areas such as tissue engineering (reconstruction of blood vessels, prosthesis and regeneration of bone tissue), pharmaceuticals (drug delivery systems), surgery (suture thread, scaffolds) and cardiology (stents, heart valves) [31]. In agriculture, biodegradable polymers are finding more and more areas of application, for example: biodegradable mulch sheets, which are degraded directly in the soil instead of being removed [32], pesticides integrated into PHA pellets, that are released while the polymer is degraded [33] and PHA-based controlled release fertilizers that are able to release nutrients to the soil in pace with plant growth requirements, reducing the usage and loss of chemical fertilizers [34]. Another emerging application for PHA is self-healing concrete, where PHA is mixed with nutrients and spores of specific limestone-producing bacteria that are used as healing agent. The spores activate when moisture and oxygen enter the concrete due to crack formation causing the growth of bacteria that consume the PHA and generate mineral precipitates that seals the cracks decreasing the water permeability of the concrete, thus prolonging its service life-time [35]. Niche applications like slow-release fertilizer/pesticides and self-healing concrete are of great interest because they could stimulate the development of waste-derived PHA supply chain infrastructure and market demand, thus fostering the scale up of processes and technologies that use waste streams, such as sewage sludges, as feedstock for PHA production. Indeed, small variations in polymer characteristics and minor fractions of impurities are not regarded as problematic for these applications. Furthermore, these niche applications could allow to a more efficient exploitation of the unique PHA properties such as biodegradability and, at the same time, allow to avoid the polymer quality requirements and the complexity of the conventional plastic industry in which the competition of the petrochemical-based plastics and their production scale hamper the PHA commercial development [35,36].

3.2.2  Current industrial production methods and developments The first patent concerning PHB, as a biodegradable thermoplastic, dates back to 1962 [37]. In 1983 the Imperial Chemical Industries marketed the P(3HB-co-3HV) for the first time, under the name of Biopol [25,38]. Subsequently, numerous companies began the production of

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  35

different PHAs and up to date, there are 10 with a high production potential [39]. Currently, all the commercialized PHA products are obtained exclusively from pure cultures of natural or engineered microbial strains which are able to accumulate high polymer concentrations intracellularly (>0.9 g PHA g−1 VSS [40]. The bacteria most commonly used for PHA production are Ralstonia eutropha and Alcaligenes latus for the production of P(3HB) and P (3HB-co3HV); Pseudomonas spp. are used for MCL PHA production; recombinant Escerichia coli and Pseudomonas oleovorans, which can produce different types of PHA depending on the substrate used [41]. In the current industrial biotechnology for PHA production, three main process operation modes are used: batch, feed-batch and continuous. The batch culture is simple in its operation but has an intrinsically low productivity due to the restriction on the maximum allowed carbon source and nutrient concentrations at the beginning of the batch, to avoid inhibitory effects. Moreover, high cell concentrations are not reached [42]. The feed-batch culture foresees a first phase identical to the batch culture, up to the last part of the exponential growth phase, where fresh substrate is introduced in order to guarantee a desired growth rate and avoid formation of by-products. This technique is used when PHA production is not associated with the growth phase (for example, in Ralstonia eutropha) and has better cell density values than that of the batch cultures [43]. Continuous production processes are characterized by continuously feeding the culture medium at a constant flow, while keeping constant the work volume by continuously removing the fermentation broth. The continuous fermentation systems are of considerable commercial interest as they have higher productivity than of the other systems, especially for strains with high maximum specific growth rates [42]. Despite the successful industrial scale production, one of the biggest drawbacks in pure culture technology is the necessity of a sterile environment with complicated contamination prevention procedures that make the process complex and expensive. In addition, the high cost of specific carbon sources, such as glucose, with a low substrate-to-PHA conversion yield, the large amount wastewater generated and the process low productivity are factors that keeps the production costs high and limit PHA market success [41,44]. Over the years, research has focused on lowering the cost of PHAs in various ways: the selection of different overproductive bacterial strains, the use of carbon sources deriving from waste and the use of more efficient fermentation techniques. The use of waste feedstocks, pre-treated to make them a suitable substrate for PHA accumulation, in the place of refined carbon sources, coupled with MMC platform, is nowadays one of the most promising cost-reduction strategies [45]. It not only allows a lower energy consumption (unsterile condition) but also reduces the complexity of the production process (simpler equipment, with less control of temperature and/or pH) and consequently the final cost of PHAs [46]. This process cannot be too complex and expensive, otherwise the economic advantage would be lost [40]. The main wastes and by-products used as carbon source for the production of PHA by MMC can be derived from different sources such as molasses from sugar production, starch-containing material, material containing cellulose and hemicellulose, whey from the dairy industry, glycerol, oils and fatty acids and wastewaters [47].

36  Current Developments in Biotechnology and Bioengineering

Another possibility for PHA cost reduction that has recently started to attract researcher’s attention are extremophilic bacteria. Microorganisms such as halophilic Halomonas spp., have been increasingly studied for cost-reduction and mass production of PHAs. Indeed, extremophiles open fermentation processes that can reduce the complexity of the sterile fermentation process and cut the cost associated with current industrial production methods [48,49].

3.3  A circular economy approach: PHA production integrated into WWTPs The synthesis of PHA in WWTPs was observed for the first time in activated sludge associated to the biological phosphorous removal process, where aerobic and anaerobic operational conditions are combined [50]. Indeed, dynamic conditions, such as those caused by alternate redox conditions (anaerobic/anoxic/aerobic) and/or discontinuous feeding regimes favor PHA-producing microorganisms over non-PHA accumulating microorganisms and can be used to promote the enrichment of PHA-storing microorganisms in activated sludge biomass [15]. The use of such MMC for PHA production is of great interest due to the lower costs associated in comparison with pure cultures technologies. Indeed, sterile conditions for limiting the growth of the non-PHA-storing phenotype are not needed, and the control requirements are lower [39]. The PHA production process by MMCs usually is composed of four independent process elements (PE): PE1, in which the raw complex organic substrate is fermented to obtain a volatile fatty acids (VFA) rich stream; PE2, in which a selection pressure is applied in order to produce biomass with a PHA accumulation potential (PAP) as high as possible; PE3, in which the enriched biomass is fed with the VFA stream produced by PE1 to maximize the PHA accumulated by the microorganisms. When cells reach the maximum PHA content, they are harvested and sent to downstream extraction processes (PE4) [51]. Following this path, the MMC platform can be integrated in WWTPs following two different approaches: direct accumulation and enrichment accumulation [36]. Both approaches follow the general scheme with the four process elements (Fig. 3.2), but there is a difference in how the functional biomass with PHA storing capacities is produced. In the enrichment accumulation approach a specific PE2 enrichment/selection step is operated to produce biomass with a PAP as high as possible, through the application of optimum selective pressures prior to the PHA accumulation phase. In the direct accumulation approach, there is no enrichment of the biomass and the polymer production rely on the PHA-storage capacity of the waste activated sludges [36]. In this case the secondary treatment of a full-scale municipal WWTPs would serve as PE2 and all the surplus activated sludge produced can be harvested and used to accumulate PHA, while the fermentation of primary sludge from the same plant can provide up to 40% of the VFA rich feedstock needed to maximize the PHA content of the produced biomass [52]. Therefore, the integration of MMC PHA production to WWTP through the direct accumulation approach could be a direct solution to turn what is actually considered a waste stream (waste activated sludge) into a useful resource to recover carbon from wastewater by biopolymer synthesis.

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  37

(A)

(B)

Enrichment accumulation

Direct accumulation Municipal wastewater treatment plant (PE2)

Wastewater Municipal wastewater treatment plant Sewage sludge

Surplus activated sludge

Acidogenic fermentation (PE1) VFA

Selected biomass

Sewage sludge Acidogenic fermentation (PE1)

VFA

Biomass enrichment/ selection (PE2)

Wastewater

PHA accumulation (PE3)

PHA accumulation (PE3)

PHA in biomass

VFA

PHA in biomass

PHA extraction (PE4) FIG. 3.2  Differences between the enrichment accumulation (A) and the direct accumulation (B) approaches in PHA production from municipal wastewater.

3.4  A detailed view of the independent PEs for PHA production by using MMCs 3.4.1  Substrate acidogenic fermentation (PE1) Volatile fatty acids are platform chemicals for PHA production by MMC since they are readily made available and efficiently converted into PHA [14], while other substrates, such as carbohydrates or glycerol tend to form glycogen instead of PHA [15,53]. Hence, when complex substrates such as sewage sludges are used in the process, their transformation in VFA through a fermentation step (PE1) is required [51]. Acidogenic fermentation is the initial phase of the biogas production by anaerobic digestion of organic matter and, it involves the bioconversion of the monomers produced from hydrolysis into hydrogen (H2), a mixture of VFA and also CO2 [54]. When mixed cultures are used for the anaerobic production of VFA from waste substrates, lot of microorganisms are involved and so, several biochemical reactions take place and the formation of different intermediates and byproducts such as ethanol and butanol can be expected [55]. Despite acetate and butyrate are the most commonly reported products of acidogenic fermentation, the production of other kind of VFA, as well as their yields, depends on the type of substrate used, the operational parameters

38  Current Developments in Biotechnology and Bioengineering

such as loading rate, pH and temperature, the inocula and the environmental conditions [52,53]. In general and depending of the waste stream used, to achieve high levels of VFA: (1) hydrolysis must be improved to produce enough soluble products for their further fermentation, (2) acidogenesis process must be promoted, and (3) acidogens inhibition should be avoided by minimizing, for example, methanogens activity [55,56]. The optimization of process parameters such as temperature, pH, retention time and organic loading rate is important to ensure biogas formation inhibition, allowing, therefore, the recovery of most of organic matter in form of VFAs [55, Vázquez-Fernández, 2021]. The VFAs obtained from the PE1 can be used as feedstock in the subsequent two PEs, that is, both, in culture selection/enrichment (PE2) and PHA accumulation (PE3). Furthermore, this fermentation step makes possible to use a huge range of complex organic substrates as feedstock for the PHA production process, including WAS generated in municipal WWTPs thus lowering the PHAs production costs and facilitating sludge management and disposal in WWTPs. Sewage-sludge-derived VFA have been successfully used for PHA production in many studies, including the first demonstration project integrated in a full scale WWTP in which biomass was able to accumulate up to almost 0.5 gPHA g−1 VSS (XX gCODPHA g−1 VSS) [18]. The results of pilot scale PE1 that performed the acidogenic fermentation of urban wastewater derived sludges are reported in Table 3.2. We should discuss a little bit the reported results of Table 3.2. In the feedstock fermentation step (PE1), attention should be paid to the quantity, but also to the quality of the produced VFA-rich stream. Indeed, it might be necessary to adjust its nutrients content since phosphorus and ammonium excess could promote microorganisms growth and reduce PHA yield [57,58]. Another factor that could hamper PHA production from waste-sludge-derived VFA, is the difficult culture selection due to the presence of non-VFAs organic compounds, such as soluble carbohydrate and protein in the fermentation liquid [59]. Indeed, the slow biodegradable non-VFA fraction keeps the microorganisms always exposed to substrate availability, thus hindering the establishment of a strong selection pressure on the MMC [58]. Consequently, strong sludge pre-treatments yielding high levels of readily available nutrients and non-VFA organics, such as high-pressure thermal hydrolysis, should be avoided. Fermentation strategies such as bioaugmentation could be instead very advantageous since enhancing the production of specific acids type allows the tailored synthesis of a biopolyester with desired properties and functionality. Indeed, the composition of PHA produced by MMC is mostly determined by the VFA stream composition, that is, the odd-to-even ratio with propionic and valeric acids (odd number of carbon atoms in its chemical structure) favoring the synthesis of 3-hydroxyvalerate and acetic and butyric acids (even number of carbon atoms in its chemical structure) promoting the production of 3-hydroxybutyrate [60]. For example, the production of a copolymer, like PHB-co-PHV would need a VFA stream where the even and odd numbered C-atom carboxylic acids are present.

3.4.2  Culture selection (PE2) The second process element is devoted to produce biomass with a high PAP, by applying the classical ecological selection principles in which, to enforce a desired metabolism, the

3–5% w/w TS

28 gVS /L

45 gTS/L

44 gVS/kg

Feedstock

Waste activated sludge

Waste activated sludge

Primary sludge

Mixture of sewage sludge 65–70% and OFMSW8 30–35%

0.86

HBu 30–40% HCa6 25–35% HAc 24–34% HPr 10–20%

0.77–0.86 HPr 51–53% HAc 25–29% HBu 12–13% HVa 7–10% 0.73 –

0.86

HAc 38% HBu2 18% HPr3 14% HIVa4 13% HVa5 11% HAc 28–38% HBu 15–26% HPr 13–23% HIVa 12–18% HVa 4–11% HPr 33–47% HAc 25–41% HBu 14–24% HVa 7–16% HAc 33% HBu 25% HCa6 14% HHe7 11% –

1

–

–

0.65

0.4

–→

References

gNH4+-N/L gPO43−-P/L gNH4+-N/L gPO43−-P/L

[78]

[122]

[19]

0.29–0.33 gNH4+N/L 0.07 gPO43−-P/L 0.67 0.11 0.69 0.22

[20]

[121]

[52]

27 g N/kg TS 2.8 g P/kg TS

sCOD:N:P ratio of 100/4.4/0.5

COD:N:P ratio of 100:5:0.1

[17]

average sCOD:N:P [77] ratio* of 100:12:7

0.21–0.33 0.9 gNH4+-N/L 0.49 gPO43−-P/L

0.27

Main acids in VFA yield the VFA mix gCODVFA Nutrients (COD basis) g−1 VS in VFA mix

Acetic acid, 2butyric acid, 3propionic acid, 4isovaleric acid, 5valeric acid, 6caproic acid, 7heptanoic acid, 8organic fraction of municipal solid waste. *Soluble COD:N:P ratio. **Total COD concentration.

1

72 (Thermal pre- 32 treatment) 37 (Fermentation)

–

–

–

Mixture of waste activated sludge 46 gVS/kg (70%) and OFMSW (30%) Mixture of OFMSW (30%) and 51 gVS/kg waste activated sludge (70%)

Cellulosic primary sludge

72 (Thermal pre- 30 treatment) 37 (Fermentation) 37 8.3–10

0.75

0.8–0.92

7 gCOD/L**

19

0.9

0.9

8.3

7

VFA concentration CODVFA gCODVFA L−1 sCOD−1

Mixture of sewage sludge (70%) 51 gVS/kg and OFMSW8 (30%)

42

37

35–55

Operating Solids temperature concentration °C

Table 3.2  PE1 performances in pilot scale studies that used urban WWTPs derived feedstocks.

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  39

40  Current Developments in Biotechnology and Bioengineering

ecosystem is engineered instead of the microorganisms [7]. When the adopted approach is that of enrichment accumulation part of the VFAs produced in the PE1 are used to select PHAstoring bacteria with a PAP as high as possible. An effective biomass selection can be obtained by varying the operational conditions of the biological system such as applying an intermittent availability of electron acceptors or an intermittent feeding regime (feast/famine), thus causing a stress to the microorganisms. Two main enrichment methods have been used: (1) anaerobic/ aerobic enrichment (AN/AE) and (2) aerobic dynamic feeding (ADF) [15]. In the first approach (AN/AE), the biomass enrichment system alternate anaerobic/aerobic cycles and the PHA accumulation occurs during the anaerobic phase. Different microorganisms, such as the polyphosphate-accumulating organisms (PAOs) and the glycogen-accumulating organisms (GAOs), take up the carbon substrate under anaerobic conditions and use it towards PHA synthesis as a consequence of the electron acceptors limitation. The energy required for this process is provided by the internally stored polyphosphate (PAOs) or glycogen (GAOs). Then, under aerobic conditions, the stored PHA is consumed for growth and maintenance while glycogen/polyphosphate is replenished using the available oxygen [61,62]. AN/AE enriched biomass has generally a low PHA accumulation potential, with biomass PHA contents not exceeding 20% g PHA g−1 VSS [15]. In order to improve the accumulation potential, the establishment of microaerophilic–aerobic condition supplying a limited amount of oxygen to the anaerobic zone been proposed, this strategy allow to enrich a biomass with a higher PAP but PHA production proved to be unstable [63,64]. In the second approach, the ADF is reported to be more efficient than AN/AE enrichment method indeed, despite its energy costs due to aeration requirements compared to AN/AE, PHA-storing MMC is typically produced applying ADF strategy [7,65]. In this strategy, short periods of substrate availability (feast) are followed by periods of unavailability (famine), under aerobic conditions. When the famine period is relatively long compared to the feast, microorganisms reduce the expression of enzymes that are essential for cell growth which will limit the maximum growth rates in the feast phase. Once a carbon source become suddenly available the feast phase starts, the carbon source fed to the MMCs is then accumulated as PHA [66]. Therefore, the bacterial population able to use the carbon source most quickly will have a competitive advantage over other populations, since they take up the substrate during the feast phase and store it as intracellular PHA granules, that will be used as carbon and energy source to withstand the famine phase, once the external carbon source is depleted. The selection of PHA-storing microorganisms occurs by the competitive advantage to accumulate PHA they possess over the other microorganisms without this ability. Indeed the bacteria expressing the PHA-storing phenotype may grow during the famine phase while non-storing PHA microorganisms will starve, so those populations of bacteria exhibiting higher PAP become enriched over time in dynamically fed bioprocesses [7,14,67]. Recently, aerobic-anoxic dynamic feeding (anoxic famine) has been applied to simultaneously selecting MMC with high PAP while also treating nitrogen-rich wastewaters, with the concurrent benefit of saving energy for aeration [19,68,69]. Whit this regime and using fermented thermal-hydrolysed sewage-sludge as substrate, Tu et al. [68] obtained a storage yield of 0.47 g

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  41

CODPHA g−1 CODVFA and a nitrate removal efficiency of 98%, with a 79% lower energy input for aeration, please also include the percent of PHA accumulated (if possible). Conca and colleagues [19] demonstrated the feasibility of PHAs production and concurrent nitrogen removal at pilot scale, using cellulosic primary sludge as a novel carbon source. By coupling a nitritation SBR with the selection SBR, they achieved an ammonia removing efficiency of 80% and a PHA yield of 0.58–0.61 gCODPHA g−1CODVFA, please also include the % of PHA accumulated (if possible). The accomplishment of aerobic-feast and anoxic-famine conditions coupled with a low OLR (0.7–1.5 kg COD/m3•d) translated into a higher selective pressure, increasing the PHA yields under feast conditions compared with an aerobic ADF process. Another aspect that has to be considered in the dynamic feeding strategy for biomass enrichment is the presence of non-VFAs organic in the carbon source used. Indeed, the presence of slow biodegradable non-VFAs fractions in the substrates make the MMC to be always exposed to a substrate-available environment and hinder a real famine period, thus significantly influencing the selection efficiency resulting in low PHA yields. In order to improve the enrichment of PHA-storing MMC using the ADF strategy, when high concentration of non-VFAs organics are involved, Tu et al. [70] introduced a sedimentation and discharge step after VFA consumption. This resulted in a reduction of non-VFAs concentrations during the famine phase and increased the culture maximum PHA storage capacity from 0.43 to 0.56 g PHA g−1 VSS. The ADF strategy for MMC selection has been usually investigated at laboratory and pilot scale in sequencing batch reactors (SBRs), due to ease of process control and simplicity of operation [14]. Continuous flow configurations are likely be more applicable at full scale, as they are more commonly used in the direct accumulation approach. The aim of such approach is to exploit the PAP built in the surplus activated sludges of the main stream of WWTPs. In this case the MMC is enriched in PHA-storing bacteria by using the COD readily biodegradable (CODRB) contained in the raw wastewater instead of the VFAs produced in the PE1 as discussed earlier in the enrichment accumulation approach. In the direct accumulation MMCs are enriched based on the bioprocess conditions of the WWTP that may tend to select PHA accumulating microorganisms over time if a selective pressure for internal carbon storage is established in the biological treatment process. Under favorable conditions of alternate substrate availability, the relatively low CODRB concentrations normally found in municipal wastewater are sufficient to produce biomass with a high PHA accumulation potential [17] indeed high PAP in activated sludge at existing WWTPs is not characteristic for one specific plant, but it has been found to be a relatively wide spread feature [18]. Bengtsson and colleagues [52] assessed the PHA accumulation potential of biomass samples from 15 different municipal WWTPs. On the one hand, PAP levels were generally higher in activated sludge from processes with only predenitrification and nitrification than that from processes with anaerobic and anoxic tanks in series. Predenitrification on the incoming wastewater under anoxic conditions may promote a stronger feast stimulation when compared to anaerobic conditions. Regardless the process configuration, the most critical factor for driving PAP enrichment is how the influent wastewater CODRB is brought in contact with the biomass [18]. Indeed, its periodic exposition to sufficient CODRB concentration is critical for

42  Current Developments in Biotechnology and Bioengineering

a feast stimulation that is a crucial factor controlling the level of MMC enrichment with respect to PAP. On the other hand, no correlation seem to exist between operating conditions such as loading rates, influent concentrations and sludge age with the biomass PHA accumulation potential [52]. Although all WWTPs may not have configurations that are favorable to the creation of feaststimulating conditions, simple and relatively minor process modifications can achieve the same effect and provide significant improvements in MMC enrichment performance without changing the plant configuration or effluent quality. Various methods to establish feast-stimulating zones have been proposed that is,: (1) a small tank up-stream of the main treatment reactor? where influent wastewater and recirculated activated sludge (RAS), or a fraction of the RAS, are brought in contact, (2) a side-stream feast SBR in which RAS is mixed with influent wastewater batch-wise at given intervals, (3) optimization of the confluence point between influent wastewater and recirculated biomass, so that the contact zone is well defined and small enough to expose the biomass to a sufficient CODRB concentration, (4) generation of a localized feast zone, for at least some of the biomass, strategically placing a separated COD-rich stream (if available) [52]. Under optimized condition, a municipal WWTP could possibly achieve up to 60% PAP (g PHA g−1 VSS) [18]. When the PHA-accumulating MMC is enriched with the direct accumulation approach, an acclimation process, consisting in repeated feast-famine cycles using the same feedstock as for the subsequent polymer accumulation, prior to the PHA accumulation phase can enhance the biomass PHA accumulation potential [52,71] Both direct accumulation and enrichment accumulation approaches have been extensively investigated and consistent results in terms of biomass PHA accumulation potential and yield have been obtained. The results and operational parameters of pilot scale studies that used urban wastewater derived substrates are reported in Table 3.3.

3.4.3  PHA accumulation (PE3) The third process element has the aim of maximizing the biomass intracellular PHA accumulation of the PHA-storing biomass selected in PE2, thus ensuring a high and stable biopolymer content and quality in MMC cells. Its efficiency strongly depends on the PHA accumulation potential of the MMC selected in the PE2. The final biopolymer content and quality of the biomass produced in PE3 is a pivotal parameter, in order to achieve the whole economic viability for the downstream processing (PE4), the biomass PHA content should be at least 0.40 gPHA g−1 VSS [72,73]. Other important parameters that should be optimized in PE3 for a commercially viable PHA production process are those related to productivity: PHA storage yield (gPHA produced per gCOD substrate consumed), PHA storage rate (gPHA per gram of active biomass per hour) and PHA productivity (gPHA L−1 h−1 or gPHA g−1 VSS h−1) [14]. In PE3, the excess sludge produced in the PE2 and the VFAs produced in the PE1 are used as microbial and carbon sources, respectively. The biomass is subjected to aerobic conditions of extended feast that stimulates precursors production and polymerisation, while, ideally, little

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  43

Table 3.3  PE2 operational parameters and performance in pilot scale studies that used urban WWTPs derived substrates. Carbon source for MMC enrichment Screened municipal wastewater Screened municipal wastewater Municipal wastewater Fermented mixture of sewage sludge and OFMSW Fermented mixture of sewage sludge and OFMSW Fermented cellulosic primary sludge Fermented mixture of sewage sludge and OFMSW Fermented mixture of sewage sludge and OFMSW

Biomass enrichment approach

SRT** d

Cycle length d

Enriched biomass PAP gPHA g−1VSS

References

0.21

1–2

0.08

0.19–0.34

[77]

3 ± 0.8

0.125

1.7 ± 1.2 8.4–22.8

0.27–0.38

[17]

–

–

17

–

0.28–0.42

[52]

4

1

1

0.25

0.46

[121]

Enrichment accumulation

3.1–5.9

1

1

0.25

0.52

[20]

Enrichment accumulation Enrichment accumulation

0.89–1.58

1.7–2.3

6–7

0.36–0.40 0.44

[19]

2.2–4.4

1

1

0.25

0.36–0.48

[122]

Enrichment accumulation

4.0

1–2

1–2

0.25–0.5

0.40–0.59

[78]

OLR gCOD (L d)−1 HRT* d

Direct accumulation Direct accumulation Direct accumulation Enrichment accumulation

3

*Hydraulic retention time. **Sludge retention time.

or no PHA is simultaneously consumed. The results of PE3 units operated in pilot scale studies that used urban wastewater derived substrates are reported in Table 3.4. Operational parameters such as temperature and pH can have an impact on PHA accumulation but they are highly dependent on other factors like the type feedstock and the type of selection strategy used [74]. Nevertheless, the most critical aspect in PE3 operation is the cultivation strategy adopted [15]. The accumulation can be achieved through a series of strategies for substrate input and/or substrate stream quality adjustment. PHA accumulation is generally conducted in batch or fedbatch mode with pulsed or continuous substrate feeding [7]. Batch mode is the less effective cultivation strategy for PHA accumulation since high substrate concentrations can cause inhibition to the bacterial population limiting the PHA productivity, and therefore it should be avoided [75,76]. The pulsed fed-batch mode, consisting substrate addition via pulses [42], is the most commonly used in MMC PHA production since it allows to obtain high PHA productivities. The pulsed fed batch mode is often coupled with a feed-on-demand control, meaning that the substrate is feed when its concentration drops below a fixed threshold in order to maintain high respiration rates while continuously repeating

44  Current Developments in Biotechnology and Bioengineering

Table 3.4  PE3 performances in pilot scale studies that used urban WWTPs derived substrates. Biomass Carbon source for enrichment MMC enrichment approach

Carbon source for PHA Accumulation accumulation strategy

Screened municipal Direct Acetic acid wastewater accumulation Screened municipal Direct Fermented wastewater accumulation sewage sludge centrate Municipal Direct Fermented wastewater accumulation primary sludge centrate Fermented mixture Enrichment Same as of sewage sludge accumulation enrichment and OFMSW Fermented mixture Enrichment Same as of sewage sludge accumulation enrichment and OFMSW Fermented cellulosic Enrichment Same as primary sludge accumulation enrichment Fermented mixture of sewage sludge and OFMSW Fermented mixture of sewage sludge and OFMSW

Enrichment Same as accumulation enrichment Enrichment Same as accumulation enrichment

PHA yield PHA content gCODPHA gPHA g−1VSS g−1CODVFA References

Fed-batch, feedon-demand (DO as control parameter) Fed-batch, feedon-demand (DO as control parameter) Fed-batch, feedon-demand (DO as control parameter) Fed-batch

0.19–0.34

0.20–0.38

[77]

0.27–0.38

0.25–0.37

[17]

0.28–0.42

0.28–0.55

[52]

0.46

0.44

[121]

Fed-batch

0.52

0.59

[20]

Fed-batch, feedon-demand (DO as control parameter) Fed-batch

0.44

0.61

[19]

0.36–0.48

0.33–0.47

[122]

Fed-batch, feedon-demand (DO as control parameter)

0.40–0.59

0.47–0.59

[78]

conditions of carbon limitation in the mixed liquor [17,19,52,77,78]. Feed-on-demand control can be triggered by changes in the biomass respiration rate, using dissolved oxygen (DO) as indicator [6,79]. Nevertheless, when fermented feedstocks with relatively low VFAs concentration are used pulsed fed-batch strategy can cause problems, since a considerable volume of feed is supplied for each new pulse, requiring to decant the biomass and discharge the exhaust supernatant. In this case the microorganisms could begin to degrade the accumulated PHA before the addition of the next pulse, therefore decreasing the process productivity [7,80]. Another disadvantage of the pulse fed-batch strategy is the change in productivity and substrate consumption rate throughout the process, which initially have higher values to then, decrease over time [74]. By supplying the substrate in a continuous manner PHA accumulation can be maximised, also obtaining a steadier PHA storing activity. Indeed, with this feeding regimen biopolymer losses between pulses is avoided and a more stable substrate concentration can be maintained through the entire accumulation step [81,82]. The pH value can be used as an indicator for continuous substrate addition, since it increases with VFAs consumption [80,83].

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  45

Another important aspect of the PHA accumulation step is the nutrients content of the fermented feedstock used. N and P limiting or deficiency conditions have been successfully applied in order to maximize PHA accumulation, since these conditions avoid a growth response in the biomass and promote the carbon conversion towards PHA production [57,84,85]. However, other studies have reported that nitrogen concentration has little effect on PHA accumulation step [86,87]. Moreover, some studies suggest that nutrient availability during accumulation can sustain PHA storage activity by promoting new biomass growth [79,88]. These facts show that the requirement for nutrients limitation, is also related to the type of substrate used and on conditions of the selection step that greatly influence the composition of the enriched culture [15]. The effect of DO limitation on the PE3 was also studied. It was observed that under microaerobic conditions, a higher proportion of the substrate accumulates as PHA, while the total accumulation capacity is not affected. On the other hand, low DO concentrations resulted in decreased PHA production rates [89]. When dealing with nutrient-rich substrates, such as fermented sewage sludge, low dissolved oxygen accumulation rather than nutrient limitation may be an advantage, since it would avoid costs related to nutrients content adjustment and reduce also those for aeration.

3.5  PHAs extraction from microbial cells (PE4) The fourth process element of the PHA production process, namely polymer extraction from microbial cell, is pivotal for the economy of the whole process since it can account ca. 50% or more of the final polymer cost [90]. After the accumulation phase, the PHA-rich biomass produced in PE3 is separated from the aqueous matrix by normal methods, such as filtration, centrifugation or sedimentation. Subsequently the PHA is recovered from the non-PHA cell mass (NPCM), which includes phospholipids, polypeptides, RNA, DNA and peptidoglycans [91]. In order to recover the intracellular PHA granules, many methods are described in the literature. All of these methods basically work on selectively dissolving the PHA in a suitable solvent that is able to pass through the cell membrane, or by rupturing bacterial cells and removing the protein layer that coats the intracellular PHA inclusion bodies [7]. Numerous factors have to be taken into consideration when adopting a specific PHA recovery method, such as: PHA biomass content (high PHA contents can increase the cells fragility); the microbial producer (the fragility of the cell wall is different for pure cultures and MMC); intracellular PHA type and composition; extracted polymer purity requirements and impact on the PHA properties (depending on the final application of the biopolymer), cost and environmental sustainability [91]. In general, since the extraction method of PHA from microbial cells is central to the process economics, the optimal method should allow to obtain the highest purity along with the highest recovery yield while maintaining the polymer properties and at the minimal cost [92]. However, in addition to these requirements, the use of environmentally friendly chemicals and simple processes with few separation steps are also essential for achieving an environmentally and economically sustainable process [93].

46  Current Developments in Biotechnology and Bioengineering

Table 3.5  MMC PHA extraction methods developed for PE4.

Method

Extraction agent

Solvent Chloroform extraction Dichloromethane

Pretreatment HCl, 37 °C

Extraction conditions

Lyophilized biomass, 3 days at 37°C Acetone, 3 h Thermal dried (50°C) biomass, 30 min 2-Butanol – Thermal dried and ground (0.71–3.15 mm) biomass, 2 h at 125°C Dimethylcarbonate NaClO, Freeze-dried biomass 1 h (DMC) 5 min–1 h at 90°C at room temperature –100°C Acetone – Thermal dried (70°C) and grounded (99% R: 59% P: 97%

0.3–0.5

[109]

–

[112]

0.1

[113]

0.1

[113]

The PHA extraction methods used on MMC derives from the ones already used for pure cultures, although MMCs PHA recovery from the biomass is more difficult due to several factors: • MMCs are claimed to be more resistant to cell hydrolysis than pure cultures, in which genetic manipulation and high content of PHA granules increase cellular fragility [94]; • MMCs produce a stronger and more complex extracellular biomass matrix that contain the PHA accumulating cells [7]; • MMCs have a stronger NPCM and a lower starting PHA levels compared to pure cultures, that result in a lower cell constrains that decrease cellular fragility [95]. So far there are no existing industrial-scale recovery processes for MMC PHA, but among the many studied extraction procedures at laboratory or pilot scale, the most used and consolidated are solvent extraction and NPCM chemical digestion. Table 3.5 summarized the procedures that have been recently applied for MMC PHA recovery.

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  47

3.5.1  Solvent extraction Following the established extraction methods applied on pure cultures, chlorinated solvents [96–99], or a mixture of chlorinated and non-chlorinated compounds [95,100] have been tested in the extraction of PHA from MMC. After the extraction, the PHA is usually recovered by solvent evaporation or polymer precipitation with water methanol or ethanol. Dichloromethane and chloroform are the solvents most applied so far to extract PHAs from MMCs. Chlorinated solvents allow to obtain a polymer with high molecular weight (MW) however, the recovery yield is reported to be very low (18–52%) [94,95]. Such result, in particular if compared to the recoveries around 100% obtainable with pure cultures, could be due to the presence of a stronger extracellular matrix in which the PHA containing biomass is embedded [95]. In this sense, a pre-treatment with acetone can be used to enhance cell breakage and improve the PHA recovery [101]. Non-chlorinated solvents have been successfully used for MMC PHA extraction, in particular ketones and alcohols, the so-called PHA poor solvents, like 2-butanol and acetone allowed to recover PHAs with a purity up to 98% but the MW dropped to 0.2–0.6 MDa [102]. Werker et al. [103] patented an extraction method specifically using PHA poor solvents on dried granulated PHA-rich biomass with particle size distribution in the range of 0.71–3.15 mm. The PHA extraction is performed within a temperature range of 120°C–160°C, achieving a PHA concentration in solvent up to 100 gPHA L−1. After extraction, the biomass is separated from the solvent and the PHA-rich solvent becomes a gel upon cooling. After gelation, the solvent is recovered from the polymer by mechanical separation leaving a polymer cake that can be further processed. In view of environmental concerns regarding the use of solvents, recyclable and environmentally friendly green solvents that have been successfully used for MMC PHA extraction are organic carbonates like dimethyl carbonate (DMC) [104]. On the one hand, Samorì et al. [94] applied a double extraction with DMC that provided a polymer with a high purity level (98%) and a MW above 1 MDa, but the recovery percentage was low, around 61% over two cycles. De Souza Reis et al. [105], on the other hand, reported a very high polymer recovery with a purity of 91%, suggesting that the efficiency of the DMC as a solvent for PHA extraction could be influenced by the MMC selection strategy used. Since it is necessary to reach higher temperatures when using non-chlorinated solvents for PHA extraction, pre-treatment methods to protect PHA quality in the biomass before the extraction are reported to be very important for polymer quality [92]. A pre-treatment widely used to increase the thermal and chemical stability of PHA contained in biomass consists in lowering the pH of mixed liquor between 2 and 5, thus reducing the level of inorganic matter associated with the biomass since it is negatively correlated with thermal stability of the PHA [106].

3.5.2  NPCM digestion Dissolving the non-PHA cell mass while maintaining the PHA granules intact is the primary objective in extraction methods based on NPCM digestion, which are suggested to be more competitive compared to solvent extraction [91]. Similarly to what is usually used on single

48  Current Developments in Biotechnology and Bioengineering

strains, the chemicals mainly applied for NPCM digestion in MMC PHA extraction are oxidants, surfactants and alkali, while enzymatic digestion on MMC is not reported, probably due to its mildness that is not suitable for MMC high cellular resistance [101]. NPCM chemical digestion by means of NaClO is a common approach for PHA recovery from MMC. NaClO is able to oxidize, thus making them water-soluble, most of the cellular material surrounding PHA granules which can be subsequently separated by precipitation [7,107]. However, PHA is not completely insoluble in the NaClO solution therefore reaction time and temperature of the treatment affect the quality of the recovered polymer due to random chain scission that can strongly decrease its MW, thus compromising the obtained polymer quality and applications [101,108]. Villano et al. [109] treated PHA-rich MMC biomass, from a lab-scale three step process fed with synthetic VFA mixture, with NaClO solution at 5% Cl2 for 3 and 24 h and reported that the reaction time has little impact on the purity and the MW of the extracted PHA. Moreover, the temperature of the NaClO treatment was inversely correlated to both the recovery and the polydispersity index. Recently, Lorini et al. [110] investigated the impact of two pre-treatment methods applied at the end of the accumulation phase to protect PHA quality before NaClO digestion, namely wet acidification with H2SO4 at pH 2 and thermal drying at 145°C for 30 min followed by 70°C overnight. The wet-acidification pre-treatment is able to better preserve the amount and the length of PHA polymer chains produced by MMC, providing an extracted polymer with three-times higher MWs (0.4 MDa) than that of thermal pre-treated biomass, a recovery of 96%, and a purity of 90%. Others commonly used chemicals for NPCM digestion are alkalies such as sodium and ammonium hydroxides. Alkaline treatments act by destabilizing the cell wall since alkaline compounds promote a saponification reaction with the lipid layer thus increasing the cell membrane permeability, that helps release the nonpolymeric cellular material [111]. Alkaline treatment on PHA rich MMC has obtained high recovery and purification percentage [109,112,113], however it causes PHA to hydrolyze, resulting in a decrease in MW (usually between 0.3 MDa and 0.5 MDa) and an increase in polydispersity index [101]. Among the alkali tested on MMC, NaOH allow a higher recovery rate, purity, and MW of the extracted PHA than NH4OH, but the polydispersity index value is also higher (4–10 vs. 2–3) [112,113]. However, NH4OH may be a better extraction agent than NaOH because it is said to be easier to recycle [104,113]. Villano et al. [109], reported that in a direct comparison between NaClO and NaOH, the former works better than the latter under the same conditions and the PHA recovered using NaClO has higher thermal stability (up to 200°C) than PHA recovered through NaOH (between room temperature? and 100°C). Another class of chemicals used in NPCM digestion extraction methods are surfactants and between them the most used is sodium dodecyl sulphate (SDS). In PHA extraction surfactants work disrupting microbial cells by incorporating themselves into the membrane lipid bilayer thus increasing the volume of the cell envelope until the membrane breaks and produce large micelles keeping membrane phospholipids suspended in the water and allowing their removal. This leads PHA to be released into the solution surrounded by non-PHA cellular materials debris that are at the same time to solubilized thus facilitating the disruption NPCM [7,114].

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  49

The recovery rate and purity degree obtained by SDS treatment are comparable to other NPCM digestion techniques, and this extraction agent obtained similar recovery with and without biomass pre-treatment. SDS was also used in association with NaOH alkaline treatment, this combination allowed to obtain better purity levels, due to an improved removal of hydrophobic impurities [112]. In order to improve the sustainability of surfactants digestion PHA recovery methods, the use of switchable surfactants, which can be removed from the liquid phase and recovered for being reused after PHA extraction, has been first proposed for pure culture and then applied to MMC biomass [104,113]. Mannina et al. [113], successfully applied the switchable surfactant ammonium laurate for PHA extraction on MMC biomass with a previous NaClO pre-treatment obtaining recovery rates up to 90% and a purity of 100%, however the MW of the extracted polymer was low (0.1 MDa).

3.5.3  Influence of the type of extraction on polymer properties Along to the polymer composition, that is determined by the feedstock and operational conditions used in biomass selection and PHA accumulation steps, the MW of PHAs accumulated by microorganisms is an important characteristic which can determine the final biopolymer applications. Indeed, polymer chain size greatly influences material properties; therefore, the MW provides key information regarding these properties and thus, the practical uses of PHAs biopolymers [92]. As an example, in order to be processed through injection molding or extrusion, PHB requires to have a MW over 0.5 MDa [115]. The MW of PHAs varies with the carbon feedstock and microbial cultures used for polymer synthesis [95,116] but different extraction and purification methods bring MW modifications in respect to the original intracellular polymers. Therefore, the choice of the extraction method has a large impact on material properties [92,112], and often the poor PHA MW is due to the polymer degradation caused by extraction processes [107]. The MW of polymers extracted with both carbonates and chlorinated solvents are reported to be the highest among the ones achieved from MMCs, ranging between 0.8 and 0.9 MDa and, even if there is a large variability between the results of different studies, this class of solvents seems to be suitable for solubilizing high MW PHA [101]. Ketones and alcohols seem more suitable for recovering lower MW biopolymers, since extractions carried out on MMC with these PHA poor solvents resulted in polymers with a lower average MW, around 0.5 MDa, although in literature few reports focusing on using PHA poor solvents for extraction from MMCs can be found [101]. The MWs of PHA extracted through NPCM digestion are largely variable since the different chemicals and extraction conditions applied have different effects on polymer degradation [7]. PHA extracted from MMC by NPCM digestion have, on average, significantly lower MW than that of pure strains, especially in the case of surfactants and alkali [101], while for solvent extraction such trend is not observed and biopolymers with comparable MW are obtained from MMC and pure microbial cultures [92]. Another factor that has an influence on PHA quality in terms of MW are the drying conditions. Bengtsson et al. [52], observed that higher initial drying temperatures resulted in a higher PHA MW. The polymer degradation kinetics are temperature dependent but the higher is the

50  Current Developments in Biotechnology and Bioengineering

initial temperature, the sooner the biomass biological activity, and thus PHA degradation, is stopped. The results of this study showed the need to engineer the drying process in order to quickly limit any biological activity (higher temperatures) without resulting in any thermal degradation of the polymer (time at higher temperatures) during drying. Freeze-drying (lyophilization) is suggested as another way to minimize polymer chain scission during the drying process, although capital and operational costs are significantly higher if compared to thermal drying [91].

3.6  Economic sustainability of PHAs production process Techno-economic assessments of the whole MMC PHA production process interrelated to the wastewater treatment process, are rarely reported in literature, and not easily comparable between them due to the difference in the feedstock used, plant schemes and parameters (such as capital costs, impacts on the other wastewater treatment process units and by-products handling and disposal costs) considered in different studies [117]. Furthermore, due to the variable quality and purity of the recovered polymers, and the lack of real application in production processes, their market price is often very volatile or even unknown. In the first techno-economic evaluation of PHA production from activated sludge, Mudliar et al. [118] considered a process that used acetate as carbon substrate and an extraction protocol consisting in a NaClO pre-treatment followed by chloroform extraction. They calculated the PHB production cost for a plant with a fermentation capacity of 100 m3 d−1 and 44% gPHA g−1CODVFA? PHB yield to be about US$ 11.8 kg−1, which reduces to US$ 5.38 kg−1 for 1,000 m3 d−1 plant capacity and 70% UNITS? PHB yield, identifying the major cost contribution in the acetic acid used as carbon substrate and chloroform used for polymer extraction. Therefore, using a waste feedstock, such as fermented sewage sludge, as a carbon source and non-solvent-based extraction methods should be able to further reduce the cost of PHA significantly. In a more recent study, Moretto et al. [20] performed a preliminary economic evaluation of an urban biorefinery technology chain for the production of PHA and biogas from sewage sludge and the organic fraction of municipal solid waste (OFMSW), where the biogas production was implemented in order to give value to the residual solids, rich in organic matter, generated by the VFA production unit. Based on the data obtained at pilot scale (PHA overall yield of 76 g PHA for each kg of volatile solids used as feedstock, and specific gas production of 0.69 m3biogas kg−1 VS) and considering a WWTP with a potential of 70,000 habitants equivalent, they estimated a total annual revenue of 552,031 € y−1 (324,543 € y−1 given by PHA production and 324,543 € y−1 given by net electric energy production from biogas) which is 23% higher than revenues from simply co-digesting sewage sludge and OFMSW for biogas production. Bengtsson et al. [18] conducted a more detailed economic analysis in the context of PHARIO, a PHA production demonstration project centered on processing surplus activated sludge biomass from a full-scale municipal wastewater to produce PHA rich biomass using VFA from fermented rich liquors from primary sludge or from industrial wastewater, in which a routine kilogram scale production was established over 10 months and the polymer material

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  51

properties were evaluated. They estimated the production cost based on a PHA production system with a capacity of 5000 tons PHA y−1 (reported as the minimum capacity to reach the economic sustainability of the process) and concluded that with the current state of the art, the PHARIO concept would be able to produce PHA at a cost price of 3.4 € kg−1 PHA, that could be competitive on the market considering a current market price of 4.5 € kg−1. The higher share of both capital and operational costs is detained by the PHA downstream processing (PE4) that account for about 50% of capital investment and 53% of operational costs. The authors of that study, however, highlighted how a cost reduction is expected by the process optimization and upscaling, since a number of opportunities for cost minimization has been already identified at pilot scale [52]. The economic impact of downstream processing was found to be the more relevant of the whole process also in the techno-economic analysis conducted on a scenario of PHB largescale production from MMC fermentation of industrial wastewater and a downstream processing based on NPCM digestion with surfactant and NaClO [119]. Pagliano et al. provided a comparison of different PHA extraction methods and draw some general rules to estimate the costs of an extraction process. They reported that NPCM digestion can be considered as a low-cost extraction process that can be applied to produce low-quality PHA, while solvents extraction methods have higher costs due to solvent recovery but allow to obtain a higher quality biopolymer and become competitive at large scale [101,120]. Depending on the time we still have, it could be interesting to make an overview of the studies accounting for a LCA of PHA production….

3.7  Conclusions and perspectives The production of MMC-based PHA from urban wastewater has not yet been developed at industrial scale, but at pilot scale it has been shown to be a ready technology with a real possibility of implementation in WWTP, offering a meaningful contribution to resource recovery from wastewater treatment. The concepts of biorefinery and circular economy are promoting the establishment of this process. However, more research is still needed to make the PHA production process more efficient, sustainable and cost-effective in order to reach the level of commercialization and to make wastewater-derived-PHA competitive in the market. The level of development of the process decreases progressively from upstream to downstream to application (Fig. 3.3). In general, the upstream bioprocess technology is well established and MMC highly enriched in PHA producing biomass can be consistently produced and used to achieve significant PHA contents of copolymer blends with PHA yields on substrate that can be close to theoretical maximum levels. The major knowledge gaps remain in the downstream process. In fact, there is a lack of research reports on pilot experiences of the downstream processes for PHA recovery, polymer properties characterization and testing of its processability in industrial equipment for specific products manufacture. Based on the current development level of PHA production process there are still many technical, economical, and biological challenges to overcome in the optimization of the process

52  Current Developments in Biotechnology and Bioengineering

Le

ss

de

ve l

op

ed

Prototyping ...

Application Knowledge at lab-scale Choice between of DSP for scale-up Limited information about properties before and after DSP Desired properties for extracted polymer

Downstream process (DSP)

M

or e

de ve lo pe d

Feasibility of using different feedstock Experience at pilot-scale Constant yield close to theoretical maximum Economical viable PHA content Evolution of PHA properties

Up stream process

FIG. 3.3  Summarized current level of development for PHA production process components. Adapted from EstevezAlonso et al. 2021 [36]. Cost minimisation PHA yield Process scale-up

Downstream processing

Challenges in MMC PHA production from waste feedstocks

Consistency of the material quality

FIG. 3.4  Main challenges in MMC PHA production from waste feedstocks.

(Fig. 3.4). The yields of PHA on waste substrate should be maximized by developing suitable pre-treatment for the acidogenic fermentation of waste feedstock and optimizing operational parameters regarding both fermentation and feeding strategies for PHA accumulation. Methods to achieve consistency of product quality in MMC PHA production, despite wastewater feedstocks variability, needs to be fully established in order to promote commercial scale productions. There is thus an urgent need of feasibility studies, at a pilot-scale, and most important at pre-commercial scale, focusing not only on PHA rich biomass production, but especially on the downstream processing and application engineering, as well as attracting stakeholder and creating the relations needed for an effective commercial development of PHA biosynthesis and extraction from wastewater feedstocks.

Acknowledgments This work was funded by the Project “Achieving wider uptake of water-smart solutions—WIDER UPTAKE” (Grant Agreement number: 869283) financed by the European Union’s Horizon 2020 Research and Innovation Programme, in which the first author, Prof. Giorgio Mannina, is Principal Investigator for the University of Palermo; local project website: https://wideruptake.unipa.it/.

Chapter 3 • Integration of polyhydroxyalkanoates (PHAs) production  53

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[61] Koller M, Gasser I, Schmid F, Berg G. Linking ecology with economy: insights into polyhydroxyalkanoateproducing microorganisms. Eng Life Sci 2011;11(3):222–37. [62] Paul E, Neuhauser E, Liu Y. Biodegradable bioplastics from fermented sludge, wastes, and effluents. In: Biological sludge minimization and biomaterials/bioenergy recovery technologies. Hoboken, NJ John Wiley & Sons, Inc.; 2012. p. 465–98. [63] Satoh H, Iwamoto Y, Mino T, Matsuo T. Activated sludge as a possible source of biodegradable plastic. Water Sci Technol 1998;38(2):103–9. [64] Takabatake H, Satoh H, Mino T, Matsuo T. Recovery of biodegradable plastics from activated sludge process. Water Sci Technol 2000;42(3–4):351–6. [65] Li M, Wilkins MR. Recent advances in polyhydroxyalkanoate production: feedstocks, strains and process developments. Int J Biol Macromol 2020;156:691–703. [66] Majone M, Massanisso P, Carucci A, Lindrea K, Tandoi V. Influence of storage on kinetic selection to control aerobic filamentous bulking. Water Sci Technol 1996;34(5):223–32. [67] Salehizadeh H, Van Loosdrecht MCM. Production of polyhydroxyalkanoates by mixed culture: recent trends and biotechnological importance. Biotechnol Adv 2004;22(3):261–79. [68] Tu W, Zhang D, Wang H, Lin Z. Polyhydroxyalkanoates (PHA) production from fermented thermal-hydrolyzed sludge by PHA-storing denitrifiers integrating PHA accumulation with nitrate removal. Bioresour Technol 2019 July;292:121895. [69] Basset N, Katsou E, Frison N, Malamis S, Dosta J, Fatone F. Integrating the selection of PHA storing biomass and nitrogen removal via nitrite in the main wastewater treatment line. Bioresour Technol 2016;200:820–9. [70] Tu W, Zou Y, Wu M, Wang H. Reducing the effect of non-volatile fatty acids (non-VFAs) on polyhydroxyalkanoates (PHA) production from fermented thermal-hydrolyzed sludge. Int J Biol Macromol 2020;155:1317–24. [71] Werker A, Bengtsson S, Hjort M, Morgan-Sagastume F, Majone M. Valentino, F. Process for enhancing polyhydroxyalkanoate accumulation in activated sludge biomass., Veolia Water Solutions and Technologies WO/2016/020884; 2016. [72] Morgan-Sagastume F, Bengtsson S, Arcos-Hernández MV, Alexandersson T, Anterrieu S, Hjort M, et al. Organic carbon recovery as biopolymers from residuals and wastewater treatment: steps from technology development to demonstration. 18th European Biosolids & Organic Resources Conference Exhibition; 2013. [73] Serafim LS, Lemos PC, Albuquerque MGE, Reis MAM. Strategies for PHA production by mixed cultures and renewable waste materials. Appl Microbiol Biotechnol 2008;81(4):615–28. [74] Serafim LS, Pereira JSM, Lemos PC. Polyhydroxyalkanoates by mixed microbial cultures: the journey so far and challenges ahead. In: Koller M, editor. Handbook of polyhydroxyalkanoates: Kinetics, bioengineering, and industrial aspects, Vol. 2. Oxon, UK and Boca Raton, FL: CRC Press; 2020. [75] Albuquerque MGE, Eiroa M, Torres C, Nunes BR, Reis MAM. Strategies for the development of a side stream process for polyhydroxyalkanoate (PHA) production from sugar cane molasses. J Biotechnol 2007;130(4):411–21. [76] Serafim LS, Lemos PC, Oliveira R, Reis MAM. Optimization of polyhydroxybutyrate production by mixed cultures submitted to aerobic dynamic feeding conditions. Biotechnol Bioeng 2004;87(2):145–60. [77] Morgan-Sagastume F, Valentino F, Hjort M, Cirne D, Karabegovic L, Gerardin F, et al. Polyhydroxyalkanoate (PHA) production from sludge and municipal wastewater treatment. Water Sci Technol 2014;69(1):177–84. [78] Moretto G, Lorini L, Pavan P, Crognale S, Tonanzi B, Rossetti S, et al. Biopolymers from urban organic waste: Influence of the solid retention time to cycle length ratio in the enrichment of a mixed microbial culture (MMC). ACS Sustain Chem Eng 2020;8(38). [79] Valentino F, Karabegovic L, Majone M, Morgan-Sagastume F, Werker A. Polyhydroxyalkanoate (PHA) storage within a mixed-culture biomass with simultaneous growth as a function of accumulation substrate nitrogen and phosphorus levels. Water Res 2015;77:49–63.

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[80] Albuquerque MGE, Martino V, Pollet E, Avérous L, Reis MAM. Mixed culture polyhydroxyalkanoate (PHA) production from volatile fatty acid (VFA)-rich streams: Effect of substrate composition and feeding regime on PHA productivity, composition and properties. J Biotechnol 2011;151(1):66–76. [81] Chen Z, Huang L, Wen Q, Guo Z. Efficient polyhydroxyalkanoate (PHA) accumulation by a new continuous feeding mode in three-stage mixed microbial culture (MMC) PHA production process. J Biotechnol 2015;209:68–75. [82] Chen H, Meng H, Nie Z, Zhang M. Polyhydroxyalkanoate production from fermented volatile fatty acids: effect of pH and feeding regimes. Bioresour Technol 2013;128:533–8. [83] Johnson K, Jiang Y, Kleerebezem R, Muyzer G, Van Loosdrecht MCM. Enrichment of a mixed bacterial culture with a high polyhydroxyalkanoate storage capacity. Biomacromolecules 2009;10(4):670–6. [84] Johnson K, Kleerebezem R, van Loosdrecht MCM. Influence of the C/N ratio on the performance of polyhydroxybutyrate (PHB) producing sequencing batch reactors at short SRTs. Water Res 2010;44(7):2141–52. [85] Venkateswar Reddy M, Venkata Mohan S. Effect of substrate load and nutrients concentration on the polyhydroxyalkanoates (PHA) production using mixed consortia through wastewater treatment. Bioresour Technol 2012;114:573–82. [86] Dionisi D, Majone M, Vallini G, Di Gregorio S, Beccari M. Effect of the applied organic load rate on biodegradable polymer production by mixed microbial cultures in a sequencing batch reactor. Biotechnol Bioeng 2006;93(1):76–88. [87] Jiang Y, Marang L, Tamis J, van Loosdrecht MCM, Dijkman H, Kleerebezem R. Waste to resource: converting paper mill wastewater to bioplastic. Water Res 2012;46(17):5517–30. [88] Dionisi D, Beccari M, Gregorio SD, Majone M, Papini MP, Vallini G. Storage of biodegradable polymers by an enriched microbial community in a sequencing batch reactor operated at high organic load rate. J Chem Technol Biotechnol 2005;80(11):1306–18. [89] Pratt S, Werker A, Morgan-Sagastume F, Lant P. Microaerophilic conditions support elevated mixed culture polyhydroxyalkanoate (PHA) yields, but result in decreased PHA production rates. Water Sci Technol 2012;65(2):243–6. [90] Macagnan KL, Alves MI, Moreira A, da S. Approaches for enhancing extraction of bacterial polyhydroxyalkanoates for industrial applications. In: Kalia VC, editor. Biotechnological applications of polyhydroxyalkanoates. Singapore: Springer; 2019. p. 389–408. [91] Koller M, Niebelschütz H, Braunegg G. Strategies for recovery and purification of poly[(R)-3-hydroxyalkanoates] (PHA) biopolyesters from surrounding biomass. Eng Life Sci 2013;13(6):549–62. [92] Majone M, Chronopoulou L, Lorini L, Martinelli A, Palocci C, Rossetti S, et al. PHA copolymers from microbial mixed cultures: synthesis, extraction and related properties. In: Koller M, editor. Current advances in biopolymer processing and characterization. New York: Nova Science Publishers; 2017. p. 223–76. [93] Heinrich D, Madkour MH, Al-Ghamdi MA, Shabbaj II, Steinbüchel A. Large scale extraction of poly(3hydroxybutyrate) from Ralstonia eutropha H16 using sodium hypochlorite. AMB Express 2012;2(1):59. [94] Samorì C, Abbondanzi F, Galletti P, Giorgini L, Mazzocchetti L, Torri C, et al. Extraction of polyhydroxyalkanoates from mixed microbial cultures: impact on polymer quality and recovery. Bioresour Technol 2015;189:195–202. [95] Patel M, Gapes DJ, Newman RH, Dare PH. Physico-chemical properties of polyhydroxyalkanoate produced by mixed-culture nitrogen-fixing bacteria. Appl Microbiol Biotechnol 2009;82(3):545–55. [96] Lemos PC, Viana C, Salgueiro EN, Ramos AM, Crespo JPS, Reiszcorr MAM. Effect of carbon source on the formation of polyhydroxyalkanoates (PHA) by a phosphate-accumulating mixed culture. Enzyme Microb Technol 1998;22(8):662–71. [97] Serafim LS, Lemos PC, Torres C, Reis MAM, Ramos AM. The influence of process parameters on the characteristics of polyhydroxyalkanoates produced by mixed cultures. Macromol Biosci 2008;8(4):355–66.

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[98] Bengtsson S, Pisco AR, Johansson P, Lemos PC, Reis MAM. Molecular weight and thermal properties of polyhydroxyalkanoates produced from fermented sugar molasses by open mixed cultures. J Biotechnol 2010;147(3):172–9. [99] Duque AF, Oliveira CSS, Carmo ITD, Gouveia AR, Pardelha F, Ramos AM, et al. Response of a three-stage process for PHA production by mixed microbial cultures to feedstock shift: impact on polymer composition. New Biotechnol 2014;31(4):276–88. [100] Hu S, McDonald AG, Coats ER. Characterization of polyhydroxybutyrate biosynthesized from crude glycerol waste using mixed microbial consortia. J Appl Polym Sci 2013;129(3):1314–21. [101] Pagliano G, Galletti P, Samorì C, Zaghini A, Torri C. Recovery of polyhydroxyalkanoates from single and mixed microbial cultures: a review. Front Bioeng Biotechnol 2021;9(February):1–28. [102] Laycock B, Arcos-Hernandez MV, Langford A, Pratt S, Werker A, Halley PJ, et al. Crystallisation and fractionation of selected polyhydroxyalkanoates produced from mixed cultures. New Biotechnol 2014;31(4):345–56. [103] Werker AG, Johansson PST, Magnusson POG. Process for the extraction of polyhydroxyalkanoates from biomass., Veolia Water Solutions and Technologies US 2015/0368393 A1; 2015. [104] Samorì C, Basaglia M, Casella S, Favaro L, Galletti P, Giorgini L, et al. Dimethyl carbonate and switchable anionic surfactants: two effective tools for the extraction of polyhydroxyalkanoates from microbial biomass. Green Chem 2015;17(2):1047–56. [105] de Souza Reis GA, Michels MHA, Fajardo GL, Lamot I, de Best JH. Optimization of green extraction and purification of PHA produced by mixed microbial cultures from sludge. Water (Switzerland) 2020;12(4). [106] Werker A, Johansson P, Magnusson P, Maurer, F, Jannasch, P. Method for recovery of stabilized polyhydroxyalkanoates from biomass that has been used to treat organic waste, International Publication Number WO 2012/022998 A1, 23.02.2012; 2012. [107] Madkour MH, Heinrich D, Alghamdi MA, Shabbaj II, Steinbüchel A. PHA recovery from biomass. Biomacromolecules 2013;14(9):2963–72. [108] Kosseva M, Rusbandi E. Recovery of polyhydroxyalkanoates from microbial biomass. In: Koller M, editor. Handbook of polyhydroxyalkanoates: Kinetics, bioengineering, and industrial aspects, Vol. 2. Oxon, UK and Boca Raton, FL: CRC Press; 2020. [109] Villano M, Valentino F, Barbetta A, Martino L, Scandola M, Majone M. Polyhydroxyalkanoates production with mixed microbial cultures: from culture selection to polymer recovery in a high-rate continuous process. New Biotechnol 2014;31(4):289–96. [110] Lorini L, Martinelli A, Pavan P, Majone M, Valentino F. Downstream processing and characterization of polyhydroxyalkanoates (PHAs) produced by mixed microbial culture (MMC) and organic urban waste as substrate. Biomass Conv Bioref 2020;11:693–703, https://doi.org/10.1007/s13399-020-00788-w. [111] Anis SNS, Md Iqbal N, Kumar S, Amirul AA. Effect of different recovery strategies of P(3HB-co-3HHx) copolymer from Cupriavidus necator recombinant harboring the PHA synthase of Chromobacterium sp. USM2. Sep Purif Technol 2013;102:111–7. [112] Jiang Y, Mikova G, Kleerebezem R, van der Wielen LAM, Cuellar MC. Feasibility study of an alkaline-based chemical treatment for the purification of polyhydroxybutyrate produced by a mixed enriched culture. AMB Express 2015;5(1). [113] Mannina G, Presti D, Montiel-Jarillo G, Suárez-Ojeda ME. Bioplastic recovery from wastewater: a new protocol for polyhydroxyalkanoates (PHA) extraction from mixed microbial cultures. Bioresour Technol 2019;282:361–9. [114] Chen Y, Chen J, Yu C, Du G, Lun S. Recovery of poly-3-hydroxybutyrate from Alcaligenes alcaligenes eutrophus by surfactant–chelate aqueous system. Process Biochem 1999;34(2):153–7. [115] Mothes G, Schnorpfeil C, Ackermann J-U. Production of PHB from crude glycerol. Eng Life Sci 2007;7(5):475–9.

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[116] van der Walle GAM, de Koning GJM, Weusthuis RA, Eggink G. Properties, modifications and applications of biopolyesters. In: Babel W, Steinbüchel A, editors. Biopolyesters. Berlin Heidelberg: Springer; 2001. p. 263–91. [117] Cabrera F, Torres Á, Campos JL, Jeison D. Effect of operational conditions on the behaviour and associated costs of mixed microbial cultures for PHA production. Polymers 2019;11(2):191. [118] Mudliar SN, Vaidya AN, Suresh Kumar M, Dahikar S, Chakrabarti T. Techno-economic evaluation of PHB production from activated sludge. Clean Technol Environ Policy 2008;10(3):255–62. [119] Fernández Dacosta C, Posada JA, Ramirez A. Large scale production of polyhydroxyalkanoates (PHAs) from wastewater: a study of techno- economics, energy use and greenhouse gas emissions. Int J Sci Res Innov 2015;9(5):6. [120] Saavedra del Oso M, Mauricio-Iglesias M, Hospido A. Evaluation and optimization of the environmental performance of PHA downstream processing. Chem Eng J 2021;412(July 2020):127687. [121] Valentino F, Moretto G, Lorini L, Bolzonella D, Pavan P, Majone M. Pilot-scale polyhydroxyalkanoate production from combined treatment of organic fraction of municipal solid waste and sewage sludge. Ind Eng Chem Res 2019;58(27):12149–58. [122] Valentino F, Lorini L, Gottardo M, Pavan P, Majone M. Effect of the temperature in a mixed culture pilot scale aerobic process for food waste and sewage sludge conversion into polyhydroxyalkanoates. J Biotechnol 2020;323(July):54–61. [123] Chan CM, Johansson P, Magnusson P, Vandi LJ, Arcos-Hernandez M, Halley P, et al. Mixed culture polyhydroxyalkanoate-rich biomass assessment and quality control using thermogravimetric measurement methods. Polym Degrad Stab 2017;144:110–20.

4 Production of volatile fatty acids from sewage sludge fermentation Dario Prestia, Bing-Jie Nib, Giorgio Manninaa a ENG I NEERI NG DEPARTM ENT, PAL E R MO U N I V E R S I T Y, PA L E R MO , I TA LY b C E N T R E F O R TE CHN O L O GY I N WATER AND WAS T E WAT E R , S C H O O L O F C I V I L A N D E N V I R O N ME N TA L EN GI NEERI NG, UNI VERS I TY O F T E C H N O L O G Y S Y D N E Y, S Y D N E Y, N S W, A U S T R A L I A

4.1 Introduction The growth of global population, together with the increasing urbanization and the always more stringent requirements for wastewater treatment plants (WWTPs) effluent quality, are fostering a fast expansion in WWTPs number and treatment capacity. As a consequence of that growth, a rapid increase in the production of sludge derived from wastewater treatment is ongoing. For instance, in 2015 the European sewage sludge production was 9.7 million tons as dry solids, while 13 million tons are estimated to be produced in 2020 [1]. Due to the generation of this large amount of sludge, the correct management of this waste stream has become one of the most challenging problems related to wastewater treatment both from an environmental and economic perspective. An improper management can cause pollution, affecting ecosystems and human health and contributing to climate change [2]. Indeed, sewage sludges are a potential source of contamination as they might contain a great number of harmful substances such as heavy metals, pathogens, and organic pollutants. Regarding the economic side, the cost of sludge management can account up to 60% of the total operation cost of the wastewater treatment plants [3]. A proper management of sludges and the adoption of effective measures can not only minimize their environmental impact but also convert sewage sludges into resources, thus contributing to save raw materials, conserve natural resources and sustain circular economy. Indeed, sewage sludges are a source of valuable elements, having nitrogen and phosphorus contents of about respectively 3–4% and 1–10% of the sludge dry weight and an organic content often higher than 45% in dry weight [4]. In recent decades, the implementation of practices for the recovery of material and energy from sewage sludge has been promoted also from European environmental legislation [5]. Consequently, anaerobic digestion for biogas production has been widely implemented as the most sustainable alternative for on-site management of sludge. In spite of its advantages, such as generation of energy rich biogas, organic pollution reduction, sludge stabilization and volume reduction, there are some major drawbacks that weaken the advantages of anaerobic digestion in sewage sludge treatment, namely: the Current Developments in Biotechnology and Bioengineering. DOI: https://doi.org/10.1016/B978-0-323-99920-5.00006-8 Copyright © 2023 Elsevier Inc. All rights reserved.

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62  Current Developments in Biotechnology and Bioengineering

low biogas yield, the low value of the produced biogas due to the low methane percentage, the potential leaking of methane as greenhouse gas, and the low carbon utilization rate of the process [6]. Anaerobic fermentation for volatile fatty acid (VFA) production is considered to be a more promising method for an efficient resource utilization in sewage sludge treatment [7]. Indeed, by interrupting anaerobic digestion in the acidogenic stage, useful high value-added VFA are recovered, instead of low-value biogas. In last decades, volatile fatty acid production from sewage sludge fermentation has drawn increasing attentions from scientists due to their wide potential of applications ranging from carbon source for biological nutrient removal process to the use as bioenergy resource for hydrogen, methane and liquid biofuels generation, or biopolymers precursor in the bioplastic industry [8]. Volatile fatty acids are short-chain organic compounds consisting of a number of carbon atoms equal to or lower than six [9]. At present, VFAs continue to be generated from the classic oil refinery scheme, but the high costs of key petroleum-derived raw materials and their non-recyclable aspects are limiting the sustainability of this production process [10]. Therefore, it is important to develop a robust fermentation process to produce VFAs biologically, from organic waste with high organic content such as sewage sludges. In this chapter, a critical illustration of the state of the art of sewage sludge fermentation for VFA production is presented, as well as, a discussion of the recent developments in the field. In this sense, the chapter is organized as follows. Firstly, the biological mechanism and controlling strategies for VFA production from sewage sludge are described. Secondly, technologies are discussed, sludge pre-treatments methods are reviewed, and the recent enhancements of the process are highlighted as well as the potential applications of sludge-derived VFAs. Finally, the economic side of the process is analyzed and challenges and future perspectives are presented.

4.2  Biological mechanism and strategies for VFA production from sewage sludge 4.2.1  Biological mechanism of sludge anaerobic fermentation The anaerobic digestion process can be considered as an ecosystem where several groups of microorganisms work interactively, in the absence of oxygen, in the conversion of complex organic compounds into final products, such as CH4, CO2, H2S, and H2O, in addition to generating new cellular material [11]. This process makes it possible to convert large amounts of organic waste into useful by-products. The anaerobic degradation route of organic matter is a multi-stage process of several serial and parallel reactions, which consists of four successive steps, namely: hydrolysis, acidogenesis, acetogenesis and methanogenesis [12]. A diagram of the process is shown in Fig. 4.1. The hydrolytic process involves the breakdown of particulate substrates that cannot be used directly by microorganisms, to soluble bioavailable substrates [13]. Hydrolysis is promoted by the action of exoenzymes excreted by hydrolytic-acidogenic bacteria. Most of them are strict anaerobes such as Clostridia, Bacteriocides, and Bifidobacteria, but some facultative anaerobes such as Enterobacteriaceae and Streptococci take part as well [14].

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  63

Complex organic matter (proteins, carbohydrates, lipids)

Soluble organic molecules

Hydrolysis

(sugars, amino acids, long chain fatty acids)

Volatile fatty acids + Alcohols

Acidogenesis

(Butirate, proprionate, valerate etc.)

Acetate

H2 + CO2 H 2S

CH4 + CO2

H2S

Acetogenesis Methanogenesis

FIG. 4.1  Schematic representation of anaerobic fermentation process.

Organic matter can be composed of three basic types of macromolecules: carbohydrates, proteins and lipids that can be converted into sugars, amino acids, and long chain fatty acids respectively. The hydrolysis of each type of compound is carried out by different enzymatic groups [15]. The versatility and the ability of the bacteria present in the fermentation reactors to use more than one substrate is a selective advantage due to the heterogeneous composition that usually characterizes wastes when used as a substrate [16]. Therefore, hydrolysis is an essential stage, and it generally constitutes the limiting step of the anaerobic fermentation process with solid or particulate material. This is because, together with the readily biodegradable organic matter, a refractory fraction that limits the substrate biodegradability also exists in heterogeneous wastes such as sewage sludge. The speed of hydrolysis is conditioned by factors such as alkalinity, temperature, particles size, etc. and can be improved with the application of pre-treatments [4]. After the hydrolysis stage, during acidogenic fermentation, anaerobic or facultative hydrolytic-acidogenic bacteria metabolize the hydrolysis products inside the cell, transforming dissolved organic matter into a great variety of fermentation products [12]. The end products are mainly volatile fatty acids (VFA) such as acetate, propionate, butyrate, and succinate, in addition to small amounts of other important by-products for later stages such as lactic acid, ethanol, ammonia, carbon dioxide and hydrogen. The specific concentrations of components in the end products mix depends on various environmental conditions, such as temperature, pH, gas-phase composition, and biomass retention time [17]. The kinetics of the acidogenesis stage is relatively fast compared to the other stages of anaerobic fermentation and while VFAs are precursors for methane formation, VFA accumulation is widely reported to be an inhibition factor of the final methanogenesis stage [18]. The final products of the acidogenic phase are transformed in this stage to acetic acid, CO2 and H2, which serve as a substrate for methanogenic archaea. Although modest quantities of acetic acid are formed during the acidogenic phase, most of it is produced from the fermentation products by two classes of bacteria: syntrophic acetogens, which degrade fatty acids to

64  Current Developments in Biotechnology and Bioengineering

acetate, H2 and CO2, and homoacetogens, which produce acetic acid using H2 and CO2 and compete with hydrogenotrophic methanogenic microorganisms [17,19]. High organic loads lead to a high metabolic activity by acid-forming bacteria and a consequent increase in hydrogen and acetic acid concentrations: in this case the acetogenic bacteria undergo product inhibition, and the lowering of pH leads to the accumulation of propionic acid and other fatty acids, as well as the pH inhibition of methanogenic bacteria, compromising the syntropy between the various microorganisms involved in the anaerobic digestion process [14]. The methanogenic phase is the final stage of anaerobic digestion, in which most of the energy contained in the substrate is converted into methane by the action of obligate anaerobic methanogenic archaea. This non-bacterial group requires stricter environmental conditions for its development than those necessary for acidogenic microorganisms [20]. The production of methane can essentially take place through two different types of reactions: one way involves methanogenesis by hydrogenotrophic archaea, which operate the anaerobic oxidation of hydrogen and the CO2 reduction to CH4, while the second way, the so-called aceticlastic pathway, involves the anaerobic dismutation of acetic acid with the formation of methane and carbon dioxide [21]. Although most of the methane production occurs through this second mechanism, the CO2 reduction pathway through H2 oxidation is essential since acetogenic hydrogen-producing bacteria grow in syntrophic association with methanogenic hydrogenotrophic organisms. Indeed, the latter, through their consumption, keep the concentration of H2 low enough so that acetogenesis is favored [22].

4.2.2  Influence of key operational conditions on VFA production Several species of microorganisms take part in the anaerobic fermentation process, each with its own metabolism, therefore different critical factors affect the subsequent stages of the digestion process. A crucial aspect for VFA production from the fermentation process is the relationship between the structure of the mixed microbial community and the different products generated. The main operation parameters of the process that influence this relationship and consequently have a great effect on the concentration, yield and composition of VFA produced from residues are: temperature, pH, retention time (RT), organic loading rate (OLR), as well as the substrate composition and some other factors [20,23].

4.2.2.1  Influence of temperature Temperature affects the production of volatile fatty acids. In particular, it acts on the activities of enzymes with consequences on the hydrolysis of particulate organic substances, and on the growth of microorganisms [24]. Hydrolytic-acidogenic bacteria have been found to exhibit better temperature adaptability and heat shocks tolerance than methanogens [25]. High temperatures are therefore more conducive to the dissolution of organic substances and the formation of intermediate VFAs during hydrolysis and fermentation [26], however excessive temperature can lead to inhibition of acid production. As shown in Table 4.1, many authors observed that an adequate increase in fermentation temperature could enhance VFA concentration, production rate and yield. As an example Yuan et al. [27] reported a 2.2-fold increase in VFA production

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  65

Table 4.1  Effect of temperature on volatile fatty acid (VFA) production and composition. Type of sludge a

WAS (VSS = 6.6 g/L) WAS (VSS = 8.3 g/L) WAS and PSb mixture (VSS = 45.3 g/L) Thermal-pretreated WAS and PS mixture (VSS = 31.6 g/L) WAS (VSS = 7.1 g/L) Ultrasonic-pretreated WAS (VSS = 7.5 g/L) WAS (VSS = 23.8 g/L)

Temperature Optimal VFA yield Main VFA composition range studied temperature (mgCOD/gVSS) at optimal temperature

References

15–30°C 10–35°C 35–55°C

30°C 35°C 55°C

261 304 170

[27] [28] [31]

35–55°C

35°C

220

Acetic and iso-valeric acids Acetic and iso-valeric acids Acetic, valeric, and butyric acids Butyric, acetic, and valeric acids

4–24.6°C 10–55°C

24.6°C 37°C

355 620

Acetic and propionic acids Acetic and iso-valeric acids

[32] [33]

40–60°C

50°C

240

Acetic and propionic acids

[69]

[31]

a

WAS, waste activated sludge. PS, primary sludge.

b

from sewage sludge alkaline fermentation when temperature is raised from 15°C to 30°C. Comparable results have been reported from Zhang et al. [28] that obtained a 4-fold increase in VFA concentration from the fermentation of sewage sludge in presence of surfactant, with a temperature enhancement from 10°C to 35°C. The most efficient and economical favorable temperatures for VFA production are reported to be in the higher end of the mesophilic range (25–40°C) [29,30]. However, the optimal temperature depends also on the other operational conditions of the fermentation reactor and on the substrate composition [20]. The application of pre-treatments on the same substrate can also have an influence on the optimal process temperature. Indeed, while studying the effect of thermal hydrolysis pre-treatment on sewage sludge fermentation for VFA production, Zhang et al. [31] found that the optimum fermentation temperature was 35°C for pre-treated sludges while it was 55°C for raw sludges. The temperature has an effect also on the composition of VFA mix produced in the fermentation process. Possible reasons for the VFA distribution change are shifts in yield and the reaction pathways due to changes in thermodynamic yields and microbial population caused by the temperature changes [32]. A clear correlation between temperature and VFA mix composition produced from sewage sludge fermentation is hard to establish, probably due to the diversity in sludges composition, with different studies often presenting contradictory results. Yuan et al. [32] reported that the increase in temperature from 4°C to 14°C in sewage sludge fermentation led to a decrease in the percentage of acetate from 55% to 43% and to a slight increase in the percentage of propionate and butyrate from 20% to 29% and 11% to 16% respectively. By further increasing the temperature up to 24.6°C, no major alteration of the composition of the volatile fatty acids was observed. In contrast, Yuan et al. [27] demonstrated that mesophilic fermentation of biological sludge at 30°C resulted in higher acetate production as a major component of VFAs compared to psychrophilic conditions at 15°C. On the other hand, the composition of volatile fatty acids during the fermentation of ultrasonically pre-treated

66  Current Developments in Biotechnology and Bioengineering

biological sludge from 10°C to 55°C [33], was not subject to significant variation. Similarly, Zhang et al. [31] reported that the composition of VFAs mix did not vary significantly when operating at 35 or 55°C. Higher temperature slightly decreased the proportion of longer chain VFAs (butyric and valeric acid from 27% to 25% and 24% respectively), and improved that of shorter chain VFAs (acetic and propionic acid from 30% and 16% to 33% and 18% respectively).

4.2.2.2  Influence of pH The pH is one of the factors that most influence the production and composition of volatile fatty acids, as it significantly affects hydrolysis, acidogenesis, methanogenesis, and the composition of the microbial community [34,35]. To optimize the production of volatile fatty acids there is a need to ensure an optimal pH, which should be favorable for both hydrolysis and the acidogenic process, while at the same time hindering methanogenesis. The methanogenic archaea have an optimum pH range of 6.6–7.5, therefore, it follows that if the fermentation takes place with different values, the production of methane will be limited. For pH values ​​above 8.0 or below 6.0, however, methane production is no longer observed with a consequent accumulation of volatile fatty acids [20]. At the same time, most acid-forming bacteria are less sensitive to pH change, they can undergo anaerobic fermentation even in acidic or alkaline conditions, in a range of pH between 3 and 12 [9]. The optimal pH values for VFA production usually depend on the type of substrate used and in general they are in the range of 5.25–11.0 [36]. However, as shown in Table 4.2, alkaline pH in the range 8.0–10.0 is reported to be optimal when sewage sludges are used as substrate [23]. Beyond inhibiting methanogenesis, the alkaline condition improves hydrolysis of sludge by breaking the sludge matrix through ionization and further dissociation of the charged acidic groups of the sludge extracellular polymeric substances [34]. This causes the release of soluble organic matters to the environment thus increasing the soluble substrate availability and promoting VFA production [37]. Regarding the influence of pH on the acidogenesis process, it was observed that acidogenesis showed poor performance in alkaline conditions compared to the condition of neutral pH [38]. The alkaline environment is not suitable for acidogenic microorganism’s metabolism, Table 4.2  Effect of pH on volatile fatty acid (VFA) production and composition. Type of sludge Carbohydrate enriched WASa (VSS = 18.3 g/L) WAS (VSS = 45 g/L) WAS (VSS = 8.3 g/L) Heat-alkali pretreated sewage sludge PSb and WAS mixture (VSS = 5 g/L) a

pH range Optimal studied pH VFA yield

References

4–11

8

520 mgCOD/gVSS Propionic and acetic acids

[36]

7–10 4–11 3–12

10 10 9

235 mgCOD/gVS 256 mg/gVSS 600 mgCOD/gVS

Acetic, propionic, and butyric acids Acetic, propionic, and iso-valeric acids Acetic and propionic acids

[38] [39] [40]

4–12

10

620 mg/gVSS

Acetic and propionic acids

[44]

WAS, waste activated sludge. PS, primary sludge.

b

Main VFA composition at optimal pH

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  67

consequently, the higher production yield of VFAs is hypothesized to be due to the increased rate of hydrolysis and methanogenesis inhibition [39]. In addition to influencing the quantity of volatile fatty acids produced, pH also influences their composition by acting on the individual fractions of VFA [40]. The shift in VFAs mix profile under different pH conditions could be explained by its effect on organic acid dissociation since undissociated organic acids can cause a product shift in bacterial cells [41]. In the case of sewage sludge fermentation, the change in VFA profile mostly involves variations in the acetate, propionate, and butyrate fractions [9]. Cokgor et al. [42] observed that a stepwise increment of the pH value in the range of 5.5 to 7.5, in primary sludge fermentation, led to a constant decrease in acetic acid ratio and an increase in propionic and C4–C5 acids. In the study of Pittmann and Steinmetz [43] a similar change in VFA profile was observed when changing the pH from 6 to 7, but when pH increased to 8 a reverse trend was observed with a maximum acetic acid ratio and a minimum propionic acid ratio. Further increasing the pH to 10 both acetic and propionic acid ratio reduced while butyric acid ratio increased. In another study increasing the pH from 7 to 9, the concentrations of butyric and propionic acids remained at the same level, while acetic acid concentration increased 5 times [34].

4.2.2.3  Influence of retention time The RT is considered as a critical operational parameter for the anaerobic fermentation process. In VFA production from sewage sludge fermentation, hydraulic RT and solid RT often coincide because substrate and microbial culture are contained in the same phase [23]. When using complex substrates such as sludges, higher retention times could lead to improvements in the production of volatile fatty acids (shown in Table 4.3), as the microorganisms would have more time to transform the organic substrate thereby increasing the hydrolysis efficiency of organic matter [20,28]. Jankowska et al. [44] observed that the rate of volatile fatty acid production significantly increased with the increase of retention time from 5 to 15 days, due to the increased solubilization of carbohydrates and proteins. Similar results were obtained by Yuan et al. [45] with Table 4.3  Effect of retention time on volatile fatty acid (VFA) production and composition. SRT range studied (days)

Optimal SRT

VFA yield

Main VFA composition at optimal RT

WASa (VSS = 8.3 g/L) PSb and WAS mixture (VSS = 5 g/L) WAS (VSS = 4.8 g/L)

3–18 5–15

12 15

139 mgCOD/gVSS 620 mg/gVSS

Acetic and iso-valeric acids [28] Acetic and propionic acids [44]

5–10

10

140 mg/g TCOD

WAS (VSS = 8.3 g/L) WAS (VSS = 23.8 g/L)

4–16 1–13

12 5

113 mgCOD/gVSS 330 mgCOD/gVSS

Acetic, propionic, and [45] butyric acids Acetic and iso-valeric acids [46] Iso-valeric, n-butyric, and [69] acetic acids

Type of sludge

a

WAS, waste activated sludge. PS, primary sludge.

b

References

68  Current Developments in Biotechnology and Bioengineering

retention times of 5, 7, and 10 days. At the same time, excessive RT values ​​may not lead to appreciable improvements in VFA production due to the limitation of the substrates. In another study published by Feng et al. [46], an increase in RT from 4 days to 12 days increased the VFA concentration by about 45%, but a further increase to 16 days led to a lower concentration of acids, probably due to the metabolic activity of methanogens. Indeed, relatively short solid retention times can result in methanogenic biomass washout since the growth rate of methanogens is lower than that of acidogens, thus preventing VFA loss due to conversion in methane [9]. The influence of retention time on the composition of volatile fatty acids varies substantially from one study to another, so to date no consistent results have been achieved regarding changes in the distribution of VFAs from fermentation of sewage sludge. Zhang et al. [28] observed that the increase of RT from 3 days to 18 days resulted in the increase in the concentration of acetate from 45% to 59% while the influence of RT on the fraction of other VFAs was not significant. In the study by Feng et al. [46] it was observed that the increase in SRT from 4 days to 16 days, increased the percentage of acetic acid from 32% to 42%, but decreased the percentage of propionic acid from 24% to 14%. On the other hand, Yuan et al. [45] reported that the percentage of acetic acid decreased from 66% to 49% as the SRT increased from 5 days to 10 days, while the percentage of propionic acid remained almost constant (16–18%). Moreover, in another study the influence of RT on VFA composition was different in alkaline environment, where the increase in RT promoted a shift toward acetate production, than in acidic environment where at the same RT a higher butyrate formation was observed [44]. The contradiction in the results of these studies is probably due also to the different operating conditions applied other than RT.

4.2.2.4  Influence of OLR Organic loading rate indicates the amount of organic matter fed daily per unit volume of the reactor. Usually, high OLR allow to a high sludge treatment capacity and high VFA production [9]. The concentration of volatile fatty acids increases with the increase of the OLR, although at high OLRs the overload of substrate and the accumulation of inhibitory compounds could have a negative impact on the degree of acidification [47]. Since the OLR is inversely proportional to the HRT, to increase the former one can decrease the latter, often its influence is not investigated separately from HRT [41]. When higher OLR are obtained by decreasing HTR, is possible to observe a lower production of volatile fatty acids. Banerjee et al. [48] showed that total VFA production from primary sludge fermentation decreased from 0.4 g/L to 0.3 g/L by the time the OLR increased from 4 g TS/(L∙day) to 7 g TS/(L∙day) following the decrease in HRT from 30 hours to 18 hours. OLR can also affect the distribution of VFA although the results reported in literature are not always concordant. Iglesias-Iglesias et al. [47] reported that an increase of the OLR value in sewage sludge fermentation caused a decrease in acetic and propionic acid ratio and an increase in butyric and n-valeric acid ratio in the VFA mix, while Jankowska et al. [49] observed that for sewage sludge fermentation, the OLR increase influenced the total VFA productivity but not the acid mix composition and the microbial community. On the other hand, in the fermentation of food waste, low OLR values were useful in the production of propionic and butyric acid, while the percentage of acetic and valeric acid increases when OLR increases [29].

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  69

4.2.2.5  Influence of other factors Beside the discussed parameters, other factors such as head space pressure and the use of additives could also influence sewage sludges fermentation process and thus the production of volatile fatty acids. The pressure that builds up in the headspace of anaerobic reactors can affect the fermentative microorganisms in different ways depending on the concentration of the different gases. Accumulation of H2 can affect the production of VFA since it alters the flow of electrons in the metabolic pathway of microorganisms [50], therefore maintaining a low H2 partial pressure in the reactor headspace can improve the acid production [20]. Also, a lower headspace pressure is reported to be more conducive to the production of soluble products including VFA and to the transformation of protein compounds [51]. Additives can be used to improve VFA production from sludge fermentation. Surfactants for example have been used since they improve the solubilization of extracellular polymeric substances, breaking the sludge matrix and releasing organic matters to the environment thus providing more substrates for hydrolysis [52]. It was observed that the addition of sodium dodecyl sulfate (SDS), a typical surfactant, with a ratio of 0.05 g SDS/g of dry sludge increased the VFA production by 5 times compared to that without surfactant [53]. Xu et al. [54] utilized sophorolipid (SL), a bio-surfactant, in the acidogenic fermentation o sewage sludge obtaining a VFA production, at 0.1 g SL/g TSS, about 5 times higher than that in the absence of SL. They also reported that SL severely inhibited methanogenesis and increased the abundance of hydrolyticacidogenic bacteria. In another study adding sodium dodecylbenzene sulfonate (SDBS) with a ratio of 0.02 g SDBS/g of dry sludge the maximum concentration of VFA increased by almost 8 times [52]. The use of these surfactants may not be economically advantageous; therefore, an alternative to reduce costs could be to use sludges with a high surfactant content [23]. Another type of additives that can be used to improve VFA production is the chemical inhibitors of methanogenesis. They work by suppressing the activity of methanogens that consume VFAs and have been widely used for the investigation of sewage sludge anaerobic digestion in order to understand the mechanisms of fermentation intermediates production and consumption [55].

4.2.3  Influence of sludge composition and carbon to nitrogen (C/N) ratio on VFA production Substrate composition is fundamental for the synthesis of volatile fatty acids, being able to influence both their quantity and the type of acids produced and the choice of the optimal operating parameters. In general, lipids are less suitable for fermentation than carbohydrates and proteins as they exhibit slower biodegradation kinetics and, when hydrolyzed, they generate long-chain fatty acids which are able to adhere to cell walls, influencing the transport of nutrients and consequently hinder the metabolism of anaerobic bacteria [56,57]. Carbohydrates are easily converted by microbial enzymes into glucose, which is immediately available for glycolysis and fermentation of volatile fatty acids [57]. Instead, proteins are generally characterized by a lower biodegradability, due to their native folded conformation, which makes them less susceptible to protease action [58]. Due to these characteristics, the hydrolysis efficiency of carbohydrates can

70  Current Developments in Biotechnology and Bioengineering

reach 80%, while that of proteins falls within the range of 40–70%. For this reason, the hydrolysis of proteins from organic substrates is considered a rate limiting step during acidogenic fermentation [57,58]. The composition of the substrate in terms of carbohydrates, proteins and lipids determines the main metabolic pathways of the fermentation process, therefore influencing the distribution of the volatile fatty acids produced. Carbohydrate-rich matrices typically support the production of propionic acid and butyric acid while the production of valeric and iso-valeric acids is supported by protein-rich waste streams [59]. The composition of the organic matter contained in sewage sludges is well known and it mainly consists in proteins (50–70%), carbohydrates (10–20%), lipids, lignin, and humic substances [7]. Beside the sludge’s organic components, a parameter that plays an important role in anaerobic acidification is the C/N. A balanced starting C/N ratio is a crucial factor in increasing acidogenic activity. To ensure an adequate nutritional balance to the bacterial population that realizes the process, C/N ratio has to assume an optimal value. Sewage sludges, due to their high protein and low carbohydrate content, have a C/N usually comprised between 5/1 and 7/1 that is lower than the values reported to be optimal for anaerobic digestion that range from 15/1 to 70/1 [7]. In the study by Liu et al. [60], three different sludges with C/N ratios of 5/1, 12/1, and 15/1 were fermented to produce VFAs and the maximum yield was obtained from the sludge with the higher C/N ratio. While this ratio influences the production yield of VFAs, at the same time it acts on the distribution of the individual VFAs. In the same study, using a model substrate, it was observed that as the C/N ratio increased from 5 to 30, the butyrate fraction gradually increased and the acetate fraction decreased. The same authors reported that a too low a C/N ratio was disadvantageous to the growth of acidogenic bacteria. Jia et al. [61] observed an improved production of VFA from sewage sludge when the C/N ratio was increased by adding perennial ryegrass to the sludge. They found that at the optimal C/N ratio of 20/1 the total VFA production yield was 369 gCOD/kgTS while the fermentation of sole sewage sludge (C/N 7/1) and sole perennial ryegrass (C/N 26/1) gave yield of 30 and 88 gCOD/kgTS respectively. At the optimal C/N ratio enzyme activity was enhanced and consequently hydrolyzation and acidification were strengthened. The C/N ratios of the substrate affect also the extent of nitrogen release during the fermentation process. Indeed, a too high C/N ratio could limit the growth of the fermentative microorganisms due to the insufficient amount of nitrogen released. On the other hand, a too low C/N ratio could result in the inhibition of the acidogenic process due to the high ammoniacal nitrogen release that causes the accumulation of free ammonia (FA), one of the most significant inhibitors of the anaerobic digestion process [60,62]. While the former issue is not relevant in sewage sludges fermentation due to their low C/N ratio, the latter is a relevant problem, especially in sludges alkaline fermentation since high pH enhances protein hydrolysis and consequently ammoniacal nitrogen release [63]. Beside inhibiting the microbial activity, FA has also been found to enhance sludge solubilization since it improves the breakdown of the sludge matrix and cell envelope [64]. In order to take advantage of the hydrolysis enhancing properties of free ammonia, Ye et al. [65] proposed a low-alkaline fermentation method for VFA production in which endogenous FA generation is promoted by a mild alkali pH, and ammonia stripping is employed to eliminate its inhibitory effect on acidogenesis. With this method

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  71

it was possible to obtain a VFA yield of 431 mgCOD/gVSS that is higher than that of traditional alkaline fermentation while at the same time reducing the high costs associated with the alkali consumption. In addition to the nutrients balance, trace elements contained in the substrate, such as iron, zinc, nickel, and cobalt, are also important for the metabolism of microorganisms involved in the fermentation process. Indeed, these trace elements are necessary in activating and maintaining enzymatic activities in the anaerobic digestion process [66]. A lot of studies investigated the influence of trace elements on biogas production but few have specially focused on VFA yield [9]. Further investigations are required to understand the influence of trace elements on VFA yield from sewage sludge fermentation.

4.2.4  Control strategies for enhancing VFA production In order to improve the production of VFAs, hydrolysis rate must be increased, the utilization of hydrolysate products in acidogenesis must be enhanced and at the same time methanogenesis must be inhibited to prevent consumption of produced VFAs by methanogenic archaea. Based on the essence of the influencing factors discussed in the previous section, the strategies for controlling the production of VFA from sewage sludge will be discussed under three aspects: (1) improvement of hydrolysis, (2) maximization of acids production, and (3) inhibition of methanogenesis.

4.2.4.1  Hydrolysis improvement Hydrolysis generally constitutes the rate-limiting step of the sludges anaerobic fermentation process [12]. Therefore, it is pivotal to improve the hydrolysis rate due to its importance in the kinetics of the whole process. A great deal of attention has been reserved to this issue and a lot of published studies analyze the influence of process operational parameters, pre-treatments and additives on the acceleration of hydrolysis rate [4]. The most used method to strengthen hydrolysis of sewage sludge is to pre-treat the sludges before the fermentation. Indeed, pre-treatments have shown high efficiency in enhancing the degradation and disintegration of both extracellular and intracellular substances making them immediately available for the production of VFA [7]. Pre-treatment methods will be further discussed in Section 4.3.1. Among operational parameters, pH is important for hydrolysis promotion. An alkaline environment (pH of 8–11) is preferred for hydrolysis. Since this condition promote the release of extracellular enzymes, sludge complex organic matter is more easily solubilized to smaller organic compounds and the breakage of the sludge matrix release soluble substrates [34,37]. Also, alkaline pH can improve the anaerobic degradability of dissolved organic matter [67]. Temperature is another important parameter that has an impact on the hydrolysis rate, higher temperatures are indeed energetically preferred as they increase the hydrolysis rate and overall VFAs production [4]. Zhang et al. [28] reported that the hydrolysis of sewage sludges improved with the gradual increase of temperature from 10°C to 35°C, with concentration of soluble proteins and carbohydrates more than doubled. Mahmoud et al. [68] also found that sewage sludge hydrolysis rate at 35°C was significantly higher than that at 25°C. Similarly, Xiong

72  Current Developments in Biotechnology and Bioengineering

et al. [69] observed that the soluble COD concentration was more than double at 50°C than at 40°, but with a further increase to 60°C hydrolysis performance worsened. Improvements in sewage sludges hydrolysis were found also upon increasing the solid retention time (SRT) [70]. Mahmoud et al. [68] found the most increase in hydrolysis when incrementing the SRT between 10 days and 15 days at 25°C and between 0 day and 10 days at 35°C. In the fermentation of sewage sludge at pH 10, Feng et al. [46] observed an increase of soluble carbohydrate and protein with the increase of SRT from 4 days to 16 days due to improved hydrolysis of extracellular polymeric substances in the sludge. Beside pre-treatments and operational parameters also the use of additives has an influence on the hydrolysis rate of sewage sludges. While some additives, that is, the previous mentioned surfactants, are used with the exact purpose of improving the solubilization of complex organic matter, other substances added to the sludges for different reasons, for example, improve dewatering, are deleterious for the hydrolysis process [4]. Chen et al. [71] studied the effect of poly-aluminum chloride (PAC), a coagulant also used for sludge dewatering in WWTP, on VFA production from sewage sludge and revealed the mechanism of how PAC hindered VFA production. With the increase of PAC addition from 0 mg Al/g TSS to 40 mg Al/g TSS the concentration of both soluble carbohydrate and protein sensibly decreased, and VFA yield decreased from 212.2 mg COD/g VSS to 138.4 mg COD/g VSS. Besides hydrolysis it was found that also the acidogenesis, and methanogenesis processes were inhibited by PAC. Similarly, it was hypothesized that also polyelectrolyte flocculants, used in sludge dewatering, could affect hydrolysis because of the greater size of sludge floc they allow to form [72].

4.2.4.2  Acidification enhancement The critical factors that determine the quantity, distribution and quality of volatile fatty acids that derive from the fermentation process of organic matter have been widely studied. In particular, the effects due to the variation of operational parameters, as reported in Section 4.2.2, have been taken into consideration as function of: pH, temperature, retention time, OLR, substrate and additives [23,73]. Most researchers have looked at one condition at a time, and there are only a few studies evaluating their combined effects; this is due to the complexity of the process even if today it is clear that these parameters have a synergistic effect as they act on cellular metabolism and on the composition of microbial communities involved in the fermentation processes [74]. In sewage sludge fermentation, although alkaline conditions usually enhance VFA production, some studies reported that at the same time they inhibit acidogenesis. Indeed acidogenic bacteria showed poor performance in transforming the hydrolyzed substrate in alkaline environment, while conditions of neutral pH have been found to be more conductive for conversion of soluble substrates in VFA [38,75–77]. Regarding the influence of temperature on acidogenesis, Yuan et al. [27] reported that at 30°C acidogenic bacteria had a greater abundance than at 15°C and the activity of enzymes involved in acidification was significantly higher. Beside process parameters optimization, other methods are reported to promote acid production pathways. For example, acetic acid production can be increased by promoting homoacetogenesis. Nie et al. [78] utilized a 2-phase system, the first phase for syntrophic acetogenesis

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  73

and the second one for homoacetogenesis, in which the gases produced in first phase were recirculated to the homoacetogenesis phase thus greatly increasing acetate yield. Converting the H2 and CO2 produced in the reactor in acetate, has also the advantage of reducing the hydrogen partial pressure hence improving acid production from syntrophic acetogenesis [20]. The use of some additives is reported to be beneficial for the acidification of sewage sludges. Ping et al. [79] reported that adding CaO2 in the fermentation reactor, not only accelerated the sludge solubilization due to the change in alkalinity and its oxidation properties, but also favored the acidification process, significantly promoting acidifying microorganisms, improving the function of lysozyme, the metabolism of monosaccharides and promoting the acid biosynthesis pathway. Liu et al. [80] observed a rise in VFA production after the addition of protease due to an increase utilization of refractory organic matter and a shift in microbial community. Wang et al. [81] reported that Triclocarban, a widely used antibacterial agent that was found to be present in sewage sludge at significant levels, is beneficial for VFA production when sludges are fermented since it enhanced the acidogenesis process, stimulated both syntrophic acidogenesis and homoacetogenesis and inhibited methanogenesis.

4.2.4.3  Methanogenesis inhibition In order to maximize the VFA yield during sludge fermentation, it is essential to prevent their conversion in CH4 and CO2 at the hands of methanogenic microorganisms. The mainly used strategies to pursue this goal are: alkaline fermentation, operational parameters optimization, and inhibitors addition [38,82,83]. Among the operational parameters, pH and SRT have the most influence on methanogenesis, indeed at a high pH range and/or lower SRT methanogens activity is inhibited [34]. Retention time influence the methane production because the growth rate of the other microorganism involved in anaerobic fermentation is higher than that of methanogens [17], therefore shorter retention times are likely to cause methanogens washout, thus preventing VFA loss due to conversion in methane and leading to their accumulation [83]. Pakarinen et al. [84] investigated the coupled effect of retention time and OLR on a methanogenic continuously stirred tank reactor; they reported that by decreasing the RT from 30 days to 6 days and increasing the OLR from 2 kgVS/m3/d to 10 kgVS/m3/d, methane concentration dropped below detection limit and VFAs concentration increased from 6.3 g/L to 15 g/L; hydrogen production was also observed and H2 concentration fluctuated between 10% and 24%. Jankowska et al. [44] reported that when fermenting sewage sludge without pH control, biogas production increased with longer RT, that is, 319 mL/gVS, 409 mL/gVS, and 480 mL/gVS for 5, 10, and 15 days, respectively while at pH 10 no biogas production was observed for RT 5 but it reached 111 mL/gVS and 154 mL/gVS for RT 10 and RT 15, respectively. The effectiveness of high pH levels makes alkaline fermentation a widely used strategy to inhibit methanogenesis in sewage sludge fermentation for VFA production since it has many advantages, such as low cost, ease of operation on large-scale, hydrolysis promotion and reduction of the inhibition to fermentation caused by the produced acids accumulation [7]. However, as discussed in Section 2.2.2, not only methanogenesis but also acidogenesis can be negatively affected by extreme pH condition.

74  Current Developments in Biotechnology and Bioengineering

Another parameter that has an influence on methanogenesis is temperature. Thermophilic conditions are reported to be less favorable for methanogenesis, and more favorable for VFA accumulation than mesophilic conditions [85,86]. Vanwonterghem et al. [83] reported that both shortening the STR and increasing temperature promoted VFA accumulation. Furthermore, they observed that the types of acid produced were mainly driven by temperature rather than retention time. In the last years many studies used chemicals inhibitors in order to suppress methanogenesis [7]. Inhibitors can be specific, that is, they inhibit specific enzymes produced only by methanogenic microorganisms, or non-specific that inhibit both the activity of methanogens and partially that of non-methanogens [55]. Examples of non-specific inhibitors are acetylene, ethylene, and various halogenated aliphatic hydrocarbons such as fluoroacetate, chloroform, and methylfluoride, while a widely used specific inhibitor is 2-bromoethanesulfonate (BES) [23]. Despite the promising results of chemical inhibitors, their employment in large scale processes must be further studied since the use of these chemicals is more expensive than the application of other strategies for methanogenesis inhibition and might be harmful for the environment [87].

4.3  Trends and innovations in VFA production from sewage sludge 4.3.1  Sludge pretreatments The main goal of pretreatments is to improve the hydrolysis rate in order to increase the productivity and yield of the fermentation process. The adoption of this strategy is essential when dealing with substrates rich in proteins and complex polymers, such as sewage sludges, to accelerate the conversion of complex organic matter in VFA [4]. Many pretreatments methods that can be applied on sludges (Table 4.4) and they can be grouped into three main categories: chemical, physical, and biological treatments [9,23]. In addition, studies are being made on hybrid pretreatment, coupling two or more pretreatment methods to take advantage of their synergistic effects on the sludge hydrolysis [7]. The choice of one or more treatments to be applied depends on the quality and quantity of sludge to be treated, the associated capital and operational costs, as well as the extent of the improvement in solubilization of complex organic matter required [88].

4.3.1.1  Chemical pretreatments In chemical pretreatments various reagents, such as acids, alkalis, and oxidizing agents, can be used. Acid and alkaline treatments improve the solubilization of extracellular polymeric substances (EPS) and favor the breakdown of cell walls, with consequent release of intracellular organic matter thus providing more biodegradable substrates for acidogenic microorganisms [89,90]. The advantages of alkaline and acid treatments are linked to ease of operation, good efficiency, and low application costs, while the disadvantages are linked to the corrosion of

Waste activated sludge

Hydrogen peroxide + microwave Thermal + alkaline + Hydrogen peroxide

Waste activated sludge

Waste activated sludge

Waste activated sludge Waste activated sludge Waste activated sludge Waste activated sludge + primary sludge Waste activated sludge

Ozone + ultrasound

Thermal + alkaline

High pressure homogenization Enzymatic Enzymatic Alkaline + microwave Alkaline + ultrasound

Ultrasound

Ultrasound

Hydrogen peroxide Potassium permanganate Thermal Thermal Ultrasound

Ozone Ozone

Acid Ozone

[103]

References

[106] [107] [110] [103]

[104]

[103]

[93]

[97] [98] [101] [94] [94]

[94] [95]

[96]

[97]

VSS reduction of 14.5%; 21.1% [96] particulate COD solubilization 36.7-fold increase in sCOD concentration [93]

VSS reduction of 68% 78.2% increase in SCOD 65.9% particulate COD solubilization 69.4% particulate COD solubilization

23% particulate COD solubilization

5.7-fold increase in sCOD/TCOD ratio

8.7-fold increase in sCOD concentration

22% particulate COD solubilization VSS reduction of 9.5%; 6.4% particulate COD solubilization 3.9-fold increase in sCOD concentration 7.4-fold increase in VFAs yield from WAS 48% particulate COD solubilization 48% particulate COD solubilization COD solubilization of 15%

4-fold increase in sCOD concentration [89] 25.8-fold increase in sCOD concentration [93]

3.7-fold increase in sCOD/TCOD ratio

Effect of pretreatment

Ozone dosage 1 g/h; Ultrasound frequency 21 kHz, energy density 0.26 W/mL H2O2 dosage 1 g/g TS; Microwave 2450 5-fold increase in sCOD concentration MHz, 5 minutes, final temp. 120°C 90°C, 5 h; pH 12; H2O2 dosage 30 mg/g TS VSS reduction of 19.5%; 30.4% particulate COD solubilization

Amylase:protease at 3:1 (w/w), 50°C, 12 h Endogenous amylase, 35°C, 7h pH 11.0; specific energy 38,400 kJ/kg TS pH 13; Ultrasound specific energy input 15,000 kJ/kg TS, frequency 20 kHz 90°C, 5 h; pH 10

Waste activated KOH, pH 13, 1 h sludge + primary sludge Waste activated sludge HCl, pH 1, 24 h Waste activated sludge Ozone dosage 1 g/h, 10 g TS/L, pH 6.8, 27°C, 1 h Waste activated sludge Ozone dosage 0.16 g/g TS Primary sludge + waste Ozone dosage 0.07 g/g TS activated sludge Waste activated sludge H2O2 dosage 1 g/g TS Waste activated sludge 0.1 g KMnO4/g TSS Waste activated sludge 170°C, 30 min Waste activated sludge 190°C, 1 h Waste activated sludge Specific energy 9350 kJ/kgTS, Frequency 20 kHz Waste activated sludge Frequency 21 kHz, ultrasonic energy density 0.26 W/mL, 1 h Waste activated Specific energy 45,000 kJ/kgTS, Frequency sludge + primary sludge 20 kHz Waste activated sludge pressure 12,000 psi; 36.0 mg NaOH/g TS

Alkaline

Pretreatment conditions

Type of sludge

Pretreatment method

Table 4.4  Effect of different sludge pretreatment methods on organic matter solubilization.

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  75

76  Current Developments in Biotechnology and Bioengineering

machinery, the need of neutralization before digestion and the risk of chemical contamination [91]. In acid pretreatment reagents such as HCl, H2SO4, H3PO4, and HNO2 are used to maintain pH between 1.0 and 5.5 [88]. Alkali pretreatment usually employs alkaline reagents such as NaOH, KOH, Mg(OH)2 and Ca(OH)2, with NaOH reported to have the best costeffectiveness [92]. Other pretreatments are based on the oxidizing agents such as ozone, hydrogen peroxide and potassium permanganate. Ozone is a strong oxidant and, along with other ozone-derived oxidation species such as hydroxyl radical, it reacts with the sludge organic matter considerably favoring its solubilization with low reaction times [93]. The disadvantages of ozone pretreatment are linked to the difficulty and high costs related to the generation of ozone that must take place on site. Optimal ozone dose ranges from 0.05 g O3/g and 0.5 g O3/g total solids, with the occurrence of a mineralization phenomenon at higher O3 doses [94], but it was demonstrated that the application of ozone dosages below 0.1 mgO3/gTS are also capable of improving fermentation performance thus making ozonation would more competitive with respect to other pretreatments [95]. Another oxidant that has been successfully used to promote sewage sludge hydrolysis is hydrogen peroxide [96,97]. It has the advantage of being economical, but the disadvantage of having an oxidation potential lower than that of ozone, respectively equal to 1.8 V and 2.1 V, with the consequence of being less reactive [97]. Recently, also potassium permanganate (KMnO4), a widely used strong oxidant, has been applied as pretreatment for sewage sludge fermentation and its effect on the generation of VFA have been investigated. Results showed that the addition of 0.1 g KMnO4/gTSS increased VFA production from 33.9 mgCOD/gVSS to 251.8 mgCOD/gVSS [98]. In addition to enhancing sludge solubilization and the degradation of recalcitrant organics matter in sludge, KMnO4 also inhibited methanogenesis and degraded emerging contaminants thus improving the quality of fermentation liquor and fermented sludges.

4.3.1.2  Physical pretreatments Physical pretreatments of sewage sludge include both mechanical and thermal treatments. Thermal pretreatments are among the most widely used and can be distinguished in: low-temperature pretreatments, with temperatures below 100°C, and high temperature pretreatments, characterized by process temperatures between 100°C and 210°C [88]. At high temperatures the release of soluble organic matter is mainly due to the physical disruption and solubilization of sludge matrix and cellular envelopes organic matter, while in low temperature hydrolysis other mechanisms such as stimulation of thermophilic bacteria and solubilization by hydrolytic enzymes contribute to sludge solubilization [99]. The use of high temperatures is reported to give better performance in terms of sludge biodegradability and methane yield [94] but, in addition to requiring more energy, temperatures higher than 180°C can lead to the formation of recalcitrant organic compounds like melanoidin that are difficult to degrade and can inhibit the fermentation process [31,100]. Thermal hydrolysis has been recently reported to be a suitable pretreatment for VFA production from sewage sludges. Zhang et al. [31] obtained a 44.6% increase in VFA production when a thermal hydrolysis pretreatment (155–175°C, 6 bar

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  77

for 30 min) was used. Kakar et al. [101] reported an enhancement in VFA yield from 0.12 gCOD/ gCOD to 0.42 gCOD/gCOD when thermal pre-treating sewage sludge at 170°C for 30 min. Microwave pretreatment is another option to improve the hydrolysis of sludges that act through thermal and athermal effects. The thermal effect is related to the generation of heat through the rotation of dipolar molecules, such as water, in the oscillating electromagnetic field. The athermal effect is caused by the alignment of the polarized side chains of the macromolecules with the poles of the electromagnetic field, thus breaking the hydrogen bonds and destabilizing the structure of the molecules [102]. However, microwave pretreatment is energy intensive and results in a higher energy consumption than conventional thermal pretreatment [88]. The mechanical pretreatments, such as sonication, and high-pressure homogenization, essentially act by disintegrating the sludge flocs and destroying the cell membranes of microorganisms through shear stresses generated by an external action [4]. This led to the release of soluble organic matter and a reduction of particles dimension with an increase in the specific biomass surface available for hydrolytic bacteria. All this can be translated into an improvement in the biodegradability of the substrate and an acceleration of the fermentation process [88]. In sonication pretreatment ultrasound waves trigger the formation of cavitation bubbles, which at the moment of collapse generate hydromechanical shear forces strong enough to break up the macromolecules present in the sludge and destroy the cell walls of microorganisms. Following the collapse, an increase in temperature and pressure is also observed, which lead to the formation of reactive hydroxyl radicals [94]. Sonication is considered an effective method for sewage sludge pretreatment and it has also been used in full scale application, however its high energy requirement is a major drawback [103]. Another mechanical pretreatment, reported to be less energy intensive, that can enhance sewage sludge fermentation is high-pressure homogenization [104]. It consists in pressurizing the sludge of up to 600 bar and make it pass through a valve where the abrupt pressure drop causes high shear forces within the sludge and cavitation that lead to cell membrane ruptures [99].

4.3.1.3  Biological pretreatments Biological pretreatments are a more recent approach compared to other types of pretreatments used to improve sewage sludge hydrolysis and thus VFA production. In general, these kinds of pretreatments make use of biologically active enzymes to improve hydrolysis and thus the biodegradability of the sludge. Enzymatic hydrolysis involves the addition of hydrolytic enzymes before, or during, the anaerobic fermentation process which allows to increase the solubilization of the particulate organic substance, destroy the structure of sludge flocs and the bacterial cells wall [4]. In several studies it was observed that dosing a mixture of enzymes had a greater impact on hydrolysis than using a single enzyme, as different enzymes are able to hydrolyze different components in the sludge [105,106]. In the case of sewage sludge, the enzymes most used, individually or in combination, are protease and amylase, which act on the main components of the sludge, namely proteins and carbohydrates. Amylase has shown greater efficiency in sludge solubilization than protease [106,107]. Xin et al. [108] reported that by pre-treating

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sewage sludge with a hydrolytic enzymes blend containing lysozyme, protease, α-amylases and cellulase, acidogenesis was strongly enhanced leading to a VFA concentration 2.5 time higher than the one obtained from untreated sludge. The advantages of enzymatic pretreatment are linked to the fact that it has a high solubilization efficiency and it does not involve the production of refractory products while being energy-efficient, environment-friendly. The major disadvantage is the high cost that, together with the low purity of commercial enzymes, constitute a fundamental limit for its widespread application [9]. However, in the last years some low-cost enzyme sources such as aged refuse, organic garbage and fungal mash have been used, suggesting that these could be a viable solution for a larger application of enzymatic pretreatment in the future [108].

4.3.1.4  Hybrid pretreatments In recent years, studies have been carried out by coupling two or more pretreatment methods to exploit the synergistic effect resulting from the combination of different technologies [99]. In this way, it is also possible to overcome the inherent disadvantages and limitations of some individual pretreatment methods. The combined alkaline and ultrasonic pretreatment is an example of a pretreatment in which the synergistic effects on hydrolysis improvement are exploited. In particular, the microbial cell walls weakened by the alkaline pretreatment results more vulnerable to the shear stress generated by the sonication pretreatment. In this way, not only a better disintegration and solubilization of the sewage sludge can be obtained, but also energy could be saved since a tendency toward a higher synergistic effect with a lower specific energy input in ultrasonic pretreatment has been reported [103]. In the ozone-ultrasound hybrid pretreatment it has been observed that the latter improve the decomposition of ozone into free radicals, while the micro bubbles of ozone act as cavitation nuclei thus helping to generate higher acoustic cavitation [93]. In order to improve sewage sludge acidogenic fermentation, Liu et al. [109] performed a study in which the effects of combined thermo-alkaline, thermo-acid, ultrasonic-acid and ultrasonic-alkaline pretreatments were assessed. It was observed that the ultrasonic-alkaline and thermo-alkaline pretreatment methods had a higher efficiency in the production of volatile fatty acids due to the significant improvement in sludge solubilization, leading to an increase in VFA produced of more than 68% and 59% respectively. The acid hybrid pretreatments on the other hand were not effective resulting in a reduction of VFA production. The combination of microwave and alkaline pretreatments through the initial addition of an alkali solution and a subsequent microwave treatment has been also investigated [110]. The preliminary alkaline breakdown of the sludge flocs, that also weakened the cell walls of the bacteria, made them more susceptible to lysis by microwave irradiation and resulted in an approximately 80% increase in VFA yield while reducing the fermentation time. Another hybrid pretreatment was investigated by Luo et al. [111] that exploited the synergic effects of surfactants and biological enzymes. They reported that surfactants cause a significant increase in the water solubility of hydrolytic enzymes, thus enhancing subsequent hydrolysis and acidification.

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  79

In general, hybrid pretreatments can produce a synergistic effect that favors the enhancing of hydrolysis and acidification, with a consequent improvement of the overall VFA production. However, not all possible combinations have been studied to date and not all the mechanisms linked to synergistic effects are yet known in detail. Therefore, there is a need for further efforts, including technical-economic feasibility studies, to identify the optimal combinations of pretreatment methods that allow to obtain better performances with a lower energy and chemicals consumption than that of single pretreatments [9,88].

4.3.2  Fermentation reactor configurations Sewage sludge acidogenic fermentation for VFA production has been usually applied, at pilot or full scale, in batch, semicontinuous and continuous stirred-tank reactors [82,112,113]. This is due to the high solids concentration of sewage sludge that does not allow the use other highrate anaerobic reactors configuration, such as packed bed reactor or up flow anaerobic sludge blanket reactor (UASB), widely used in acidogenic fermentation of much diluted wastewaters [23,73]. Although stirred-tank reactors are established technologies that ensure stable operation and good production of volatile fatty acids, they have some disadvantages. In particular, they are not suitable for high concentrations of substrate, as a part of it remains in the effluent, and the accumulation of VFA and ammonia produced in fermentation can inhibit the bacterial activity lowering the VFA yield [20]. In addition, since the fermentation process requires a long retention time, a large reactor volume is usually required, which results in high cost and energy consumption [114]. Membrane technology has been tested as a solution to these problems, indeed it allows to decouple solid and hydraulic retention times, resulting in increased production of volatile fatty acids and smaller reactor size [115,116]. Since membranes hold and concentrate functional bacteria in the reactor, the degradation of refractory organic matter is accelerated and higher organic loads can be applied. Furthermore, in the membrane bioreactor (MBR) for sludge anaerobic fermentation the product inhibition problem is alleviated since fermentation products are continuously separated from the biomass [116]. Despite all the advantages, the main bottleneck that penalizes the application of membrane technology for VFA recovery from sludge is the membrane fouling. Indeed, fermented sludge has smaller particle size, higher concentration of suspended solids, higher viscosity, and a greater filtration resistance than activated sludge and this cause a more severe membrane fouling hindering the stable operation of MBR in anaerobic fermentation of sludge [117]. Self-forming dynamic membranes (SFDM) have recently been used to replace conventional membrane technology since they have a higher membrane fouling resistance and the membrane flux can be fully restored [114,118]. The separation layer of SFDMs is formed by the precipitation of microbial flocs and suspended solids on an inexpensive porous carrier material, such as a silk screen or steel mesh, with a large pore size. Compared with conventional membrane filtration, SFDM has a flux 1 to 4 folds higher, lower filtration pressure (the effluent can flow out by gravity), lower investment costs and the membrane fouling can be completely cleaned [118]. Liu et al. [114] used a SFDM in order to improve the production of volatile fatty

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acids from the fermentation of sewage sludge and reported a VFA yield increase of 228%. An increase in the activity of relevant enzymes that promoted substrate degradation was observed, and SFDM allowed the accumulation of acid-forming bacteria with a low growth rate capable of degrading organic refractory matter. Another approach for VFA production from sewage sludge, that allows the use of high loading rate fermentation systems, such as UASB reactors, is liquid fermentation [119]. Raw sludge is pre-treated in order to maximize the solubilization of organic matter, subsequently the sludge liquid is separated and fermented in a continuous operated UASB reactor. This process allows to use organic loads up to 5 times higher than the traditional slurry fermentation resulting in a high-rate VFA production and above all in much lower energy consumption. Zhang et al. [119] operated a continuous liquid fermentation of pre-treated sewage sludge with an energy consumption of 9.9 kWh/kg VFAs that was almost 10 times lower than that of batch slurry fermentation. Liquid fermentation of sewage sludge was performed also by Liu et al. [120] in a SFDM reactor obtaining an energy consumption of 18 kwh/kg VFA. Removing solids before the fermentation process indeed decrease the viscosity of the mixture that has to be stirred, thus substantially decreasing the energy required for mixing that usually account for more than 80% of the total energy consumed in the fermentation process [119,120].

4.3.3  Enhanced strategies for VFA production from sludge fermentation In the last two decades, research on VFA production from sewage sludge has increased dramatically [4,7] and consequently many methods to improve the process have been reported. One example is the stepwise alkaline fermentation. Indeed, while it is widely demonstrated that alkaline pH promotes hydrolysis of sludges organic matter, resulting in a greater availability of soluble substrate for VFA production [27,34,36,41,65,73,80,110], the mayor drawbacks of alkaline fermentation are that at high pH: (1) acidogens bacteria might be inhibited, (2) the biodegradability of soluble organic matter can decrease, and (3) the cost of chemicals to maintain alkaline condition impact the economic sustainability of the process [37,63]. To overcome these issues and further improve alkaline fermentation performances, stepwise alkaline fermentation has been proposed [34,121]. It consists in dividing the sludge fermentation process into two or more stages in which the pH is controlled at different values. The aim is to maximize chemical hydrolysis in the first stage, with high pH levels, while promoting biological hydrolysis and acidogenesis in the subsequent stages by applying milder alkaline conditions, thus obtaining a better VFA yield [35]. A conceptual model showing the advantages of stepwise pH control is reported in Fig. 4.2. Wang et al. [35] found that the optimal condition for acidogenic fermentation of sewage sludge was a stepwise decrease in pH from 11 to 9 after the firsts 3 days. The early sludge fermentation at pH 11 enhanced the hydrolysis of sludge and reduced the total amount of methanogens thus providing a niche for acidogenic microorganisms to convert soluble proteins into VFA. The subsequent pH decrease promoted the proliferation of a variety of acetogens, and facilitates the acidification of the hydrolyzed substrate thus promoting the

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  81

Fermentation pH 11

Strong influence

Hydrolyzers

SCOD solubilization

Fermentation pH 9

Normal influence

Diverse acidogens

SCFA production

Microbial community

Microbial community

Stepwise pH fermentation Process synergies pH 11

Environmental variable

pH 9

Performance dynamics

Environmental variable

Performance dynamics

FIG. 4.2  Conceptual model showing the potential of stepwise pH control in enhancing volatile fatty acid (VFA) production. From Wang et al., 2019 [35].

efficient production of VFA from sludge. Stepwise alkaline fermentation is therefore regarded as a significant improvement of the traditional alkaline fermentation since it enhances the economic efficiency of the whole process by allowing an increased revenue from VFA production and a reduction in the cost of chemicals [35]. Another strategy recently proposed to enhance VFA production from sewage sludge is bioaugmentation [74]. It consists in the addition of a pure culture of efficient acid-producing bacteria in order to enrich the fermentative microorganism population and improve the conversion of organic matter in VFA. The most promising feature of bio-augmentation is that, by choosing the appropriate bacterial strain to enrich the mixed culture in the fermentation reactor, it is possible to enhance the production of a specific acid type [122]. Co-fermentation of sewage sludges with other carbon-rich substrates is another technique that has been applied to improve fermentation efficiency and maximize the VFA yield [6]. Sewage sludge indeed generally has a relatively low C/N ratio, low carbon biodegradability and during its fermentation ammonia, which at high concentrations is toxic to microorganisms and might even completely stop fermentation, is released [60,63]. Co-fermentation with other kinds of biodegradable organic wastes, such as agricultural residues and municipal solid waste, is advantageous because it can balance the C/N ratio, increase the organic content, and dilute the inhibitory and/or toxic compounds such as ammonia or other pollutant that can be contained in sludges [123–125]. On the other way round, the sludge addition can improve the buffer capacity in the fermentation of organic wastes such as food residues and improve hydrolysis of lignocellulosic substrates [124].

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4.4  Final applications of sludge-derived VFA and economic evaluation 4.4.1  Sludge-derived VFA applications Volatile fatty acids are widely used for various applications in industries, such as pharmaceutical, food processing, chemical, textile, plastic industries, and many others. The VFA mix obtained from sewage sludge fermentation is, in most cases, not suitable for direct industrial applications as high value-added chemicals and, unless it is separated into the pure form of each component, has little value itself [8]. The separation and concentration of VFAs contained in fermentation liquid is however an energy intensive process, and their sustainable recovery still poses a challenge to the scientific community [10]. On the other hands, VFA rich fermentation broth can be directly used in various applications such as the widely applied direct conversion in methane or hydrogen by anaerobic fermentation [73], biological nutrients removal in WWTPs in place of expensive external carbon sources [112] and microbial conversion in high value-added products such as polyhydroxyalkanoates (PHA) [126] and microbial lipids for biofuels production [127].

4.4.1.1  Carbon source for nutrient removal in WWTP A viable application of VFA produced from sewage sludge is their use as external carbon source for the biological nutrient removal (BNR) in wastewater treatment plants. The carbon requirement for nitrogen removal in nitrification/denitrification process is in the range of 5–10 mgCOD/mgN, while 7.5–10.7 mgCOD/mgP are required for biological phosphorus removal [128]. Since the organic carbon present in urban wastewater is often limited, an external carbon source can be required to meet the wastewater discharge standards concerning nutrients. Usually, when external carbon sources are involved into the process, commercial chemicals such as methanol, acetate and glucose are used [74]. Since commercial chemicals can be expensive, VFAs produced from sewage sludge fermentation have been identified as suitable and more economical alternative. Fermentation liquid however could contain high concentrations of nutrients depending on substrate composition, fermentation operational conditions and pre-treatments used [32]. Therefore, it might be necessary to recover the P and N released in the fermentation liquid in order to reduce their load on the BNR process. This could be achieved through conventional physicochemical methods or more sustainable strategies such as struvite precipitation, with the additional advantage of recovering a valuable resource [129]. However, it is still possible to achieve high nutrients removal efficiency in BNR without removing nitrogen and phosphorus from fermentation effluent [112]. A number of studies have shown that the use of VFA derived from sludge fermentation can obtain nutrients removal efficiency similar or even better than that of commercial chemicals such as synthetic acetic acid or methanol [112,125,130,131]. A study by Su et al. [132] showed that sulfur-containing cysteine amino ​​ acid, present in sludge fermentation liquid, help in reducing the toxicity caused by reactive nitrogen species that generate during biological nutrient removal and inhibit the BNR microbes. Zhu and Chen

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  83

[133] reported another advantage of using VFA-rich liquid from fermented sludge. In their study, by using sewage sludge alkaline fermentation liquid instead of synthetic acetic acid as external carbon source in an anaerobic−aerobic (low dissolved oxygen) processes, the production of NO and N2O reduced by 50.0% and 68.7%, respectively, while the removal efficiency of total nitrogen and phosphorus improved. Furthermore, Liu et al. [112] proved the economic feasibility of a full-scale process in which alkaline fermentation of sewage sludge was used for VFAs production then used for biological nutrients removal obtaining phosphorus and nitrogen removal efficiencies of about 89.7% and 73.4%, respectively. The use of sewage sludge fermentation liquid for BNR processes have a lot of advantages but scarce, if any, disadvantages.

4.4.1.2  Polyhydroxyalkanoates production Polyhydroxyalkanoates are a class of biodegradable, thermoplastic biopolyesters synthesized by many groups of bacteria as intracellular carbon and energy stocks under unbalanced growth conditions [126]. PHAs are industrially produced using pure bacterial culture under strict sterility conditions and refined carbon substrates, such as sucrose, glucose, or glycerol [134]. In order to lower production cost in last decades a lot of efforts have been directed to develop efficient PHAs production processes using open mixed microbial cultures (MMC) and waste derived substrates such as VFA. These processes typically involve a first acidogenic fermentation step for volatile fatty acids production followed by a culture selection step and the subsequent accumulation phase in which VFAs are used for PHA accumulating biomass production and polymer accumulation respectively [126]. A diagram of the typical MMC PHA production process is shown in Fig. 4.2. In many studies sewage sludge derived VFA have been successfully used for PHA production, including the first demonstration project integrated in a full scale WWTP in which biomass was able to accumulate up to almost 0.5 gPHA/gVSS [135]. Before using VFA-rich fermentation liquid for PHA accumulation, it might be necessary to adjust its nutrients content, since phosphorus and ammonium excess could promote microorganisms’ growth and reduce PHA yield [136,137]. Another factor that could hamper PHA production from sludge-derived VFA, is the poor culture selection due to the presence of non-VFAs organic compounds, such as soluble carbohydrate and protein, in the fermentation liquid [31]. Indeed, the slow biodegradable nonVFA fraction keeps the microorganisms always exposed to substrate availability, thus hindering the establishment of a strong selection pressure on the mixed culture [137]. In order to improve PHA production when high concentration of non-VFAs organics is involved. Tu et  al. [138] introduced a sedimentation and discharge step after VFA consumption in the culture selection phase, this resulted in a reduction of non-VFAs concentrations during the famine phase and increased the culture maximum PHA storage capacity from 0.43 gPHA/gVSS to 0.56 gPHA/gVSS (Fig. 4.3). When the final use of sewage sludge derived VFA is PHA production strong sludge pretreatments yielding high levels of readily available nutrients and non-VFA organics, such as high-pressure thermal hydrolysis, should be avoided. Fermentation strategies such as bioaugmentation could be instead very advantageous since enhancing the production of specific acids type allows the tailored synthesis of a biopolymer with desired properties and functionality.

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Sludge fermentation

Selection of PHA storing biomass

VFAs

Selected biomass

VFAs

PHA accumulation

PHA containing biomass

PHA extraction from biomass

FIG. 4.3  Diagram of the typical mixed microbial cultures (MMC) polyhydroxyalkanoates (PHA) production process.

Indeed, the composition of PHA produced by MMC is determined by the VFA composition with propionic and valeric acids favoring the synthesis of 3-hydroxyvalerate and acetic and butyric acids promoting the production of 3-hydroxybutyrate [139].

4.4.2  Economical evaluation of VFA production from sewage sludge Fermentation of sewage sludge to generate volatile fatty acids is reported to be more profitable than biogas production due to the higher economic value of VFA compared to methane and the lower retention time needed for acidogenic fermentation which results in lower reactors volume [125]. Nevertheless, techno-economic assessments of the whole VFA production process interrelated to the wastewater treatment process, are rarely reported in literature [140] and not easily comparable between them due to the difference in the parameters (such as capital costs, impacts on the other wastewater treatment process units and sludge handling and disposal costs) and final uses of the recovered products considered in different studies. Furthermore, due to the variable quality and purity of the recovered materials, their market price is often very volatile or even unknown [141]. Yuan et al. [27] conducted an economic analysis based on a 5 L semicontinuous sewage sludge fermentation reactor for VFA production, to compare the economic performances of microtherm (15°C) and mesotherm (30°C) fermentation. Taking in consideration the costs for energy, sludge management and external carbon source for nutrient removal, a total cost saving of 56 USD/year and 63 USD/year was obtained for microtherm and mesotherm fermentation respectively, thus demonstrating the economic feasibility of mesotherm sludge fermentation for VFA production. In a recent study Liu et al. [112] operated a full-scale sludge fermentation system to produce VFA for nutrient removal application, and provided an economical evaluation of the process in comparison to anaerobic digestion for methane production. Not taking in consideration the costs related to fermented sludge dewatering and disposal, VFA and biogas recovered produced a net profit of 9.12 USD/m3 and 3.71 USD/m3 sludge respectively. Bahreini et al. [140] assessed the impact of VFA recovery from primary and rotating belt filtration sludges, on the sludge management and disposal cost by integrating a fermentation reactor for VFA production with the subsequent anaerobic digestion for biogas production from the residual solids. Considering the revenues of methane and VFA recovery, and the reduction in sludge volume, the integration of acidogenic fermentation reactor for VFA production, could recover about 12.8% and 17.3% of the sludge management costs for primary and rotating belt filtration sludge respectively.

Chapter 4 • Production of volatile fatty acids from sewage sludge fermentation  85

Furthermore, they estimated a payback period of about 3.6 years and 2.7 years for a WWTP with a treatment capacity of 100,000 m3/d, equipped with primary clarification or rotating belt filtration respectively. Da Ros et al. [113] performed a cost-benefit analysis of cellulosic primary sludges (CPS) fermentation comparing three different scenarios: scenario 1 (S1, used as reference scenario) where CPS is used only for biogas production through anaerobic digestion and the carbon source for nutrients removal is purchased from the market; scenario 2 (S2) where an acidogenic fermentation unit is introduced prior to digestion for biogas production, and the produced VFA are used as external carbon source; scenario 3 (S3) where the produced VFA are used as substrate for PHA production. Taking into consideration costs, revenues, and potential savings derived from the internal use of the recovered resources, net incomes for each scenario were found to be 53, 67, and 95 €/ton TS for S1, S2, and S3 respectively. In a recent study [142], a more comprehensive economic evaluation of sewage sludge fermentation for VFAs and biogas production was performed. Three scenarios were evaluated: classical activated sludge process, in which excess sludge is digested to produce biogas (CAS1) or fermented to produce VFA (CAS2) and a modified process (MAS) in which organics from wastewater are recovered directly by adsorption on the excess activated sludge and then fermented for VFA production. This strategy allows to prevent organic matter loss through CO2 formation and also avoid its synthesis into bacterial organism thus enhancing its bioavailability for VFA production, furthermore it allows to reduce the size of the activated sludge process and the energy needed for aeration. Each scenario was evaluated considering all the running costs for the global operation of a WWTP treating 100,000 m3/d of urban sewage, but the depreciation cost was not considered. Taking in consideration the revenues from VFA and methane produced, the net incomes from the plant operation were estimated to be –4575, –3620, and 89.6 USD/day for of CAS1, CAS2 and MAS respectively, demonstrating that the new strategy of direct organics recovery for VFA production could allow WWTP to make net profits.

4.5  Conclusions and perspectives In recent years, research on anaerobic digestion of sewage sludge for VFA production has increased in prominence and many advances have been made. The economic feasibility of the process for biological nutrient removal has been demonstrated at full scale [112], and an increasing number of studies are focusing on the optimization and combination of different treatment technologies to improve their efficacy. However, there are still some issues that should be further studied to enhance industrial scale VFA production from sewage sludge. In the future, VFA recovery from sewage sludge should be addressed from a wider perspective. The whole process has to be taken in consideration, starting from the sludge production with the aim to improve its organic carbon bioavailability at the source and not only whit expensive pre-treatments. In order to achieve this result, more focus should be given on changing the conventional way in which wastewater organic matter is concentrated in the sludges before their fermentation. Improving the bioavailability of sludges organic matter, for example, by direct adsorption on excess activated sludge, would improve the VFA yield while the process energy consumption and direct CO2 emissions would decrease.

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Regarding the sludge acidogenic fermentation, although the influences of the operational conditions on the process have been widely examined, the effects of their synergistic or antagonistic interaction on VFA production and composition, still need to be systematically studied. In order to enhance the sustainability of the process, more efforts should be directed on the integration of the fermentation process for VFA production with the recovery of other resources such as nutrients, and suitable reactor design should be studied. Also, attention should be paid on the recovery of residual organics in the fermented sludge. Novel strategies for VFA production such as bioaugmentation, self-forming dynamic membranes and liquid fermentation, seems promising to improve the sustainability of the process while providing better quality high value-added products that could allow for more profitable applications, but they should be further developed and upscaled. Another challenge that the scientific community has not solved yet, is to achieve a sustainable separation and concentration of VFAs contained in fermentation liquid. This would allow their use, as high value-added chemicals, for industrial applications in place of petroleumderived VFA and it would greatly foster the development of industrial scale sludge-derived VFA production plants. While the technology and the understanding of the process is improving, it is increasingly evident that sewage sludge acidogenic fermentation have a great potential to become a significant source of volatile fatty acids for many applications, especially for biotechnological processes that does not require pure VFA such as bioplastic and biofuels production, that will sustain the transition toward the circular economy model.

Acknowledgments This work was funded by the Project “Achieving wider uptake of water-smart solutions—WIDER UPTAKE” (Grant Agreement number: 869283) financed by the European Union’s Horizon 2020 Research and Innovation Programme, in which the first author, Prof. Giorgio Mannina, is Principal Investigator for the University of Palermo; local project website: https://wideruptake.unipa.it/.

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5 Zeolites for the nutrient recovery from wastewater Sofia Maria Muscarellaa, Luigi Badaluccoa, Vito Armando Laudicinaa, Giorgio Manninab a DEPARTM ENT O F AGRI CULT U R A L , F O O D A N D F O R E S T S C I E N C E S , U N I V E R S I T Y OF  PA L ERM O , PAL ERM O , I TALY b E N G I N E E R I N G D E PA RT ME N T, PA L E R MO U N I V E R S I T Y, PA L E R MO , I TA LY

5.1 Introduction To meet the growing food demand of the world population, excessive use of chemical fertilizers is occurring to improve soil fertility and crop production. The excessive use of chemical fertilizers is not economically and environmentally sustainable. Indeed, from one hand, due to the increasing demand of fertilizers is rising their costs whereas, on the other hand, the accumulation of fertilizers in wastewaters is altering the homeostasis of the ecosystems thus causing serious damages to human health [1,2]. The recovery of nutrients, such as nitrogen (N) and phosphorus (P), from wastewaters is a good option to counteract both economic and environmental issues raised by the excessive use of fertilizers [3]. Adsorption is among the most widely used methods for nutrient recovery from wastewaters due to its efficiency and simplicity. The choice of appropriate adsorbent materials is a key issue for ensuring high performance and low costs of the process [4]. Over the years, several materials have been studied to absorb nutrients from wastewaters. Zeolites, both natural and modified, have attracted great attention due to their relevant specific capacity, selectivity, safety, and stability [5]. However, considering that in municipal effluents the inorganic P exists as the anionic forms of dihydrogen or monohydrogen phosphates (H2PO4− and HPO42−, respectively) and N in both cationic (ammonium, NH4+) and anionic (nitrate, NO3−) form [6], natural zeolites can be only used for the direct recovery of NH4+. This derives from the chemical properties of zeolites, which consist of an aluminosilicate structure comprising a tridimensional tetrahedral arrangement of silicon cations (Si4+) or aluminum cations (Al3+) each surrounded by four oxygen anions (O2−). Some Si4+ ions are substituted by Al3+ ions, resulting in the imbalance of net negative charges in the structure of the tectosilicate [7]. Thus, thanks to their negative charges, zeolites can adsorb cations from the surrounding environment. In recent years, also P recovery by zeolites has been studied, focusing on the possibility of simultaneous N and P recovery. In fact, being the NH4+ adsorption from wastewaters by zeolites mainly based on exchange reactions, other cations are released into the solution, thus potentially leading to synchronous phosphate (PO43−) precipitation [8]. The Current Developments in Biotechnology and Bioengineering. DOI: https://doi.org/10.1016/B978-0-323-99920-5.00012-3 Copyright © 2023 Elsevier Inc. All rights reserved.

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simultaneous recovery of N and P by zeolites can be also obtained by improving their affinity for PO43− through their activation by aluminum salt solution [9]. Once recovered, the nutrients can be reused as a soil fertilizer within a circular economy vision [10]. In fact, it is possible to reuse both the nutrients desorbed from the zeolites after regeneration through the production of fertilizers or by applying the enriched zeolites directly to the soil. This second option of reusing the recovered nutrients is possible because zeolites are considered one of the natural inorganic soil conditioners widely used to improve the physical and chemical properties of the soil; moreover, when enriched, they are able to retain the most important nutrients for plants, making them available when they are required by the plants themselves thus acting as a slowrelease fertilizer [11]. In the present chapter, firstly, the chemical composition, structure and properties of zeolites are reported and, thereafter the differences between natural and synthetic zeolites and their main fields of use are discussed. Finally, the possibility of using zeolites for the recovery of nutrients from wastewaters, the mechanisms that allow the nutrient adsorption, the methods of regeneration and reuse of recovered nutrients and enriched zeolites in agriculture are discussed.

5.2  Structure and chemical composition of zeolites 5.2.1  Chemical composition and structure of zeolites The name “zeolite” was proposed by the Swedish mineralogist Crönstedt in 1756. He observed that these minerals (especially stilbite) emit bubbles due to the release of interstitial water when heated, without changing the aluminosilicate structure. This characteristic gives rise to the name zeolite from ζέω (boil) e λίθος (stone) in Greek [12]. Zeolite is a crystalline hydrated aluminosilicate. Its basic structure consists of an interconnected tetrahedral arrangement in which silicon (Si4+) and aluminum (Al3+) cations placed in the center of each tetrahedron coordinate four oxygen anions (O2−) placed at the vertexes. Each oxygen atom is shared by two tetrahedra thus connecting them and resulting in inorganic macromolecules with a unique three-dimensional framework. The resulting Si (or Al) to O ratio is 1:2. The isomorphic substitutions of some Si4+ by Al3+ ions determine a net negative charge in the structure of the zeolites, arising from the difference in formal valence between the tetrahedra (AlO4)5− and (SiO4)4− and normally found on one of the oxygen anions connected to an aluminum cation. Such negative charges are balanced by alkaline earth metals, such as sodium (Na+), potassium (K+) or calcium (Ca2+); also, lithium (Li+), magnesium (Mg2+), strontium (Sr2+), and barium (Ba2+) can be found in some zeolites [13], beyond water. These ions are found on the external surface of zeolite, bound with the aluminosilicate structure by weaker electrostatic bonds [14,15].  The zeolite framework contains open cavities in the form of pores, channels and cages that host the water molecules and cations, but only those of appropriate molecular size to fit into the pores are admitted. The framework can be interrupted by (OH, F) groups; these occupy the vertexes of a tetrahedron and are not shared with adjacent tetrahedra [16].

Chapter 5 • Zeolites for the nutrient recovery from wastewater  97

O O

Si

Oxygen O

=

O

Silicon or Aluminum

FIG. 5.1  Primary building unit of zeolites.

The chemical composition of a zeolite can hence be represented by a formula of the type:

M x/n , (H2 O)z [ AlO2 )x (SiO2 )y ] (5.1)

where M is an extra-framework cation with valence n, and x, y are pertinent values of molar concentrations of Al and Si in the zeolite framework, and z is molar concentration of H2O [17].

5.2.1.1  Primary and secondary building units of zeolites The crystal structure of zeolite is usually divided into primary building unit (PBU) and secondary building unit (SBU). The PBU framework is a TO4 tetrahedron, where the central T-atom is usually Si or Al, and the outer atoms are O (Fig. 5.1). The simple geometric shape is formed by the interconnection between two or more tetrahedra. Therefore, the formed link is called the SBU. The complexity of SBU starts with simple rings, double rings, polyhedra, and even more complex units linked together. The unit cell of zeolite always contains an integer number of SBUs [18] (Fig. 5.2). In addition to the above-mentioned PBU and SBU, the zeolite can also contain other components, such as prisms and cages, which are called composite building units (CBU). The cage is defined as a polyhedron whose largest ring is not enough to allow molecules larger than water to pass through.  They appear in several different framework structures and can be useful in identifying relationships between framework types [7]. CBUs are therefore a grouping of finite and infinite building units and involve various forms, such as SBUs, polyhedra such as the sodalite unit (β cavity) and chains of tetrahedra [18,19]. To date, 253 different zeolite framework types

Oxygen Cations Silicon or aluminum H2O

FIG. 5.2  Secondary building unit of zeolites.

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Table 5.1  Division of zeolites according to pore size, which influences their properties and fields of use. Category

Number of rings

Diameter

Small-pore Medium-pore Large-pore Extra-large-pore zeolites

≤8 rings 10-ring 12-ring >12-ring

Having free diameters of 0.30–0.45 nm 0.45–0.60 nm in free diameter 0.6–0.8 nm more than 8 nm

have been identified in natural or synthetic zeolites, and the International Zeolite Association (IZA) has assigned a three-letter code for each; this short note is used to describe the pore system [20]. Each porous network is characterized by the direction of the channel, the number of atoms in the pore, and the crystallographic free diameter of the pore (in Å), the number of asterisks indicating whether the system is one-, two- or three-dimensional [21].

5.2.1.2  Pores, cages, and channels The structure of zeolites is characterized by tetrahedra bond in different ways thus producing different types of tetrahedral structures. Each tetrahedron is composed of TO4 (T = Si, Al, and P, among others). This structure contains open cavities in the form of pores, channels, and cages. These are usually occupied by H2O molecules, and by extra-framework alkali metal ions or organic amines which can be removed during the ion exchange process. The pores in zeolites are classified according to their diameter into micropores, mesopores or macropores (Table 5.1). A cage is a polyhedral pore whose faces are so narrow that cannot host species larger than a water molecule, while a cavity is the pore with at least one face large enough to host a water molecule. In practice, a cage has no face larger than 6 rings (a ring is usually referred to as a cycle of tetrahedrally coordinated silicon or aluminum and oxygen atoms), whereas a cavity has at least one. Channels, that are infinitely large in one dimension, are large enough to allow the spread of host species (i.e., larger than six rings). The internal cages and channels of zeolites depend on the specific arrangement of the SBUs and are organized according to a hierarchical structure, which gives zeolites adsorption, catalysis, and ion exchange properties [22–26].

5.2.1.3  Cation exchange capacity of zeolites Natural zeolites have different properties, one of these is the cation exchange capacity (CEC), that represents the theoretical quantity of cations that can be exchanged per mass of zeolite at a given pH. The CEC is reported to range between 60 and 230 cmol(+) kg−1 and depends on the amount of negative charge resulting from the isomorphic substitution of Si by Al [27].  Generally, the process takes place with the uptake of one or more cationic species, that in the natural zeolites are usually represented by: Na+, K+, Mg2+, and Ca2+ [7], and the simultaneous displacement of an equivalent amount of one or more other cationic species [28]. The cation exchange process may be presented by the following equation:

Z A B(Zz+) B + Z B A(Zs +) A ↔ Z A B(Zs +) B + Z B A(Zz+) A (5.2)

Chapter 5 • Zeolites for the nutrient recovery from wastewater  99

where zA and zB represent the A and B exchangeable ion charges and the coefficients (z) and (s) refer to zeolite and to aqueous solution, respectively. The reaction proceeds until equilibrium is attained [29]. Due to such property, zeolites have been found to be a good adsorbent of gases and liquids materials thus being useful to separate them from the surrounding environment. Furthermore, the cation exchange has been identified as the mechanism for NH4+ adsorption.

5.2.1.4  Selectivity of zeolites Selectivity is defined as the affinity or preference of certain types of materials for certain ions [30]. It depends on the interaction of materials and ions in relation to ionic radius (molecular sieve), ion hydration energy and electrostatic bond energy [31]. Shape selective adsorption is a process that separates molecules based on inclusion or exclusion from specific zeolite pores. This phenomenon is possible because the large structural cavities and the entry channels leading into them contain water molecules, which form hydration spheres around exchangeable cations. On removal of water by heating at 350–400°C, small molecules can pass through entry channels, but larger molecules are excluded—the so called “molecular sieve” property of crystalline zeolites [32]. The ion exchange selectivity usually depends on the hydrated ionic radius of cations, that is, on the number of water molecules that surrounds the cations. An example of cation selective sequence in natural zeolite is, as reported by Sarioglu [33]:

K + > NH 4 + > Na + > Ca 2+ > Fe3+ > Al 3+ > Mg 2+ (5.3)

5.3  Natural zeolites and synthetic zeolites Many zeolites are found naturally as minerals and are mined extensively in many parts of the world. Others are synthetic and manufactured commercially for specific uses or produced by scientists trying to improve their chemistry.

5.3.1  Natural zeolites Natural zeolite is found all over the world. Although there are no accurate data on the total amount of zeolites, it is reported that they occur on all continents with different mineral concentrations and types [34]. Natural zeolites are produced by the interaction of volcanic rock and volcanic ash with alkaline groundwater, usually related to mountainous areas such as the Caucasus and Balkans. Deposits have also been found in the Himalayas, near the Gulf of Mexico, Switzerland, and the United States [29,35]. They can occur in the crystalline form found in igneous and metamorphic rocks, as well as in smaller-diameter crystal grains accumulated in sedimentary rocks [30,36]. Indeed, the performance of a natural zeolite depends on several factors including structure, size and shape, the charge density of the anionic framework, and the ionic charge and concentration of the external electrolyte solution [31]. Among all the natural zeolites discovered (over 60), only six are abundant in natural deposits worldwide: analcime (ANA), chabazite (CHA), clinoptilolite (HEU), erionite (ERI), mordenite

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(MOR), and phillipsite (PHI). The most common natural zeolite can be purchased at very low cost, making it an economical alternative to synthetic zeolite [32]. Due to its characteristics, natural zeolite is used as soil amendment, animal feed additives, ion exchangers in industrial, agricultural, and municipal wastewater treatment, Sr and Cs radioisotope absorbents in the nuclear industry, and nuclear accident clean-up (Chernobyl), soil replacement in horticulture, veterinary and medical applications [33]. Natural zeolites are used for wide applications due to its low cost and availability, on the other hand, the industrial application of natural zeolite is limited because, as mentioned above, its properties strictly depend on its crystal structure. The main disadvantage is that the channel diameter is too small to absorb larger gas molecules and organic compounds [34].  Furthermore, zeolites, due to the very different conditions under which they are formed in nature, are rarely pure but contaminated to varying degrees by other minerals, metals, quartz, or other zeolites [7]. Therefore, natural zeolite is excluded from many important commercial applications where uniformity and purity are essential [35]. It is important to clarify that zeolite deposits are non-renewable resources.

5.3.1.1  Most important natural zeolites Clinoptilolite-heulandite Clinoptilolite is one of the most common natural zeolites. It is easily obtained from mines and is suitable for use as an adsorbent due to its characteristics. Although many researchers classify it as heulandite, its Si/Al ratio and thermal stability make it different from heulandite (4 ≤ Si/ Al < 5.2 and Si/Al < 4 at 750–800°C and 450–550°C, respectively [35,37]. The heulandite group minerals are classified according to the main exchangeable cations as follows: K-, Na-, Ca- and Sr-heulandite and K-, Na- and Ca-clinoptilolite [18,38]. The IZA identification code is HEU and the chemical formula is (Na,K,Ca)2-3Al3(Al,Si)2Si13O36•12(H2O). Chabazite Chabazite is a zeolite with the chemical formula CaAl2Si4O12•6H2O, which consists of a hydrated calcium and aluminum silicate contained in white to yellow or red creamy glass crystals. It has small pores with an effective pore size of about 4 Å and a good cation exchange capacity. Nowadays, due to their unique properties, chabazites are widely used to remove trace gases and as a water softener [39,40] Phillipsite The name phillipsite is derived from the name of the English mineralogist William Phillips. It is a mineral belonging to the group of zeolites, potassium and calcium silicate, with the formula (K,Na,Ca)1-2(Si,Al)8O16•6(H2O). It crystallizes in the monoclinic system and forms pseudorhombic, colorless, brittle, and twin crystals. It is often found in leucitic and tefritic rocks, and less frequently in basalts. Major deposits of phillipsite occur at Aci Castello in Sicily and at Capo di Bove near Rome in Italy [41].

5.3.2  Synthetic zeolites Synthetic zeolites, like natural zeolites, have different structures and properties. However, in contrast to the natural zeolites, the synthetic ones being obtained under controlled conditions,

Chapter 5 • Zeolites for the nutrient recovery from wastewater  101

have standardized and pure form (e.g., pores diameters are known), and better adsorption performance and cation exchange capacity thus being more useful for industrial applications [42]. Synthetic zeolite properties can be improved by physical and chemical treatments (hydrothermal synthesis, molten salt method, fusion method, alkali activation, microwave-assisted synthesis, synthesis by dialysis [42] thus increasing the hydrophilicity/hydrophobicity towards many ions or organic adsorbents [36,43]. Considering the industrial importance of zeolite, and due to its structural complexity and inherent scientific interest in its chemistry, a lot of effort has been made for zeolite synthesis [44]. The history of artificial zeolite can be traced back to the levynite produced in the laboratory claimed by St. Claire Deville in 1862. However, the zeolite synthesis as known today originated from Richard Barrer and Robert Milton. They showed that the final product can be obtained by heating the aluminosilicate raw material in the presence of an alkaline solution within a few hours or days, depending on the type of raw material and the process conditions (temperature, pressure) [45]. Among all the methods, hydrothermal method is effective, cheap, easiest, and commonly adopted by IZA (International Zeolite Association). It usually takes place in an alkaline (high pH) medium, in which an amorphous reagent containing silica and aluminum is mixed with a cation source. This aqueous mixture is then heated in an autoclave at temperatures above 100°C. Initially, after raising the synthesis temperature (induction period), the reactants maintain an amorphous composition and following this it is possible to detect the products, that is, the crystalline zeolites that have formed. Subsequently, all the amorphous materials are replaced by a single and equal mass of zeolite crystals, which are recovered by a process of filtration, washing, and drying [46,47].

5.3.2.1  Most important synthetic zeolites More than 150 zeolites have been synthesized, the most common are zeolites A (commonly used as a laundry detergent), X and Y (two different types of faujasites, used for catalytic cracking), ZMS-5 (a branded name for pentasil-zeolite). Zeolite A Zeolite A has a framework structure called Linde A (LTA), which consists of sodalite (SOD) cages connected by four rings and has a Si/Al ratio of 1:1. The SOD units are combined to produce an alpha cage with a diameter of 11.4 Å (a large cavity in the center of the structure), and two channel systems are connected to allow movement of Na+ ions and water molecules [45]. It is characterized by the formula |(Na12+(H2O)27|8[Al12Si12O48]8. Zeolite X and zeolite Y Zeolite X and zeolite Y belong to the family of aluminosilicate molecular sieves with faujasite type structure (FAU). The chemical formula is |(Ca,Mg,Na2)29 (H2O)240|[Al58Si134O384]–FAU (IZA). Faujasite is a rare zeolite, although its synthetic counterparts Linde X and Linde Y are widely used as adsorbents and catalysts. The difference between zeolite X and zeolite Y is determined from their Si/Al atomic ratio (usually between 1 and 1.5 for X and higher for Y zeolite) [48].

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Table 5.2  Chemical and physical characteristics of natural and synthetic zeolites. Si/Al ratio

CEC Structure (meq g−1) type Application

Clinoptilolite- (Na,K,Ca)2heulendite 3Al3(Al,Si)2Si13O36•12(H2O)

5

2.2

HEU

Chabazite

CaAl2Si4O12•6(H2O)

2–5

3.9

CHA

Phillipsite Zeolite A Zeolite X and Y Zeolite ZMS-5

(K,Na,Ca)1-2(Si,Al)8O16•6(H2O) Na12[(SiO2)12(AlO2)12].27 H2O |(Ca,Mg, Na2)29 (H2O)240|[Al58 Si134O384]–FAU NanAlnSi96–nO192•16H2O (n is between 0 and 27)

2 4.5 1 5.3 1.2–2.5 3.2–4.5

PHI LTA FAU

20

MFI

Name

Chemical formula

2

Different application fields, in particularly is used as an adsorbent due to its natural characteristics Widely used to remove trace gases and as a water softener Widely used as a water softener Laundry detergent Catalytic cracking Used to convert alcohols directly into gasoline

Zeolite ZMS-5 ZMS-5 is the prototype high silica zeolite ZSM-5, which was discovered in 1972. ZSM-5 belongs to the molecular pentasil (MFI) family, and its chemical formula is NanAlnSi96–nO192•16H2O (n is between 0 and 27). The catalyst is called ZSM-5 because its pore size is 5Å (Angstroms) and it has a Si/Al ratio of more than five [49]. Table 5.2.

5.4  Applications of zeolites Due to its characteristics, zeolite is one of the most important inorganic cation exchangers used in catalysis, agriculture, water, and wastewater treatment, as well as in industrial applications, nuclear waste, animal feed additives and biochemical applications [50].

5.4.1 Catalysis In terms of value, the consumption of zeolite catalysts accounts for about 27% of the global zeolite market, but this percentage may increase [51]. In fact, zeolites are used as catalysts in the oil refining and petrochemical industry because of their superiority performance, stability and selectivity in large conversion and upgrade processes compared to their amorphous equivalents. The action of catalysts takes place within the pores of zeolites selected on the basis of their unique characteristics of structure, morphology and composition [52,53]. The compatibility of the zeolite as a catalyst with the specific reaction can be considered the key to the economic success of this method [54]. Two important processes involving the use of zeolite as a refinery catalyst are fluid catalytic cracking (mainly gasoline production process) and hydrocracking (mainly middle distillate production process) [55].

5.4.2 Agriculture Zeolite is commonly used in agriculture to improve physical and chemical properties of the soil, such as water and nutrient retention capacity, electrical conductivity, soil aggregation,

Chapter 5 • Zeolites for the nutrient recovery from wastewater  103

Precipitation

Run-off N K P

S Zeolite

N

S

K

Infiltration

K

H2O

H2O

H2O K

K

P

P

N

K

S

H2O

H2O P

P

N S

S

N

FIG. 5.3  Mechanism by which zeolites retain nutrients in the soil.

permeability, infiltration rate and saturated hydraulic conductivity [56]. This is because of the unique characteristics of zeolites such as high pore volume, low bulk density and cation exchange capacity [57]. Ravali et al. [58] found the application of 7.5 t ha−1 of zeolite, sieved at 2 mm, increases CEC by 37% compared to the untreated soil. CEC increase is of paramount importance because in turn affects nutrient retention capacity. MacKown and Tucker [59] found that the application of zeolites reduces nitrate leaching in sandy soils probably due to a decrease of the nitrification process. Indeed, clinoptilolite is able to slow down the NH4+ oxidation by nitrifying bacteria [56]. With regard to the effect of zeolite on water infiltration rate in soils, Xiubin and Zhanbin [60], in a study conducted under laboratory and field conditions, observed that natural mordenite with a particle size lesser than 0.25 mm, increased infiltration by 7–30% on gentle slope soils and more than 50% on steep slope soils. In addition, they found that soil moisture could increase by 0.4–1.8% in extreme drought conditions, and by 5–15% in general situation. It is undoubtedly that zeolite ameliorate physical and chemical properties of soils as demonstrated by the larger number of studies. However, more studies are needed to elucidate the effect of the soil quality improvement on crop grown and productivity (Fig. 5.3).

5.4.3  Industrial wastewater treatment Heavy metals are classified as metal elements with a relatively high density (> 5 g cm−3) that are hazardous and harmful to human health, including impaired growth and development, cancer, organ damage, nervous system damage, and even, in extreme cases, death. Industrial wastewaters containing heavy metals comes from different industries [36]. A large number of elements fall into this category, but the most environmentally relevant elements are Cd, Cr, Cu, Ni, Zn, Pb

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and Hg [61]. Natural zeolites as adsorbents for heavy metals have attracted great interest due to their valuable ion exchange capacity. In fact, they can exchange or adsorb various cations such as cesium (Cs) and strontium (Sr), and heavy metals such as cadmium (Cd), lead (Pb), nickel (Ni), manganese (Mn), zinc (Zn), chromium (Cr), iron (Fe), and copper (Cu) [62].

5.5  Use of zeolite for nutrients recovery One of the main fields of use for zeolites is the recovery of nutrients from wastewaters. This capacity depends on the negative charges of the zeolites, which allows them to adsorb the cations from the surrounding environment. They have been used mainly to recover NH4+ from wastewater as this is one of the cations for which zeolites have the greatest selectivity. In recent years, the use of zeolites is also being extended to the recovery of PO43− from wastewaters, using modified zeolites.

5.5.1  Nutrients recovery mechanism Over the years, several studies have been conducted to evaluate the application of zeolite in NH4+ adsorption, with a particular focus on operational capacity. The ability of zeolite to adsorb NH4+ depends on several factors such as initial NH4+ concentration, medium size, contact time and temperature [27,63]. Currently, the most used ion exchange media for NH4+ removal is the natural zeolite clinoptilolite (in the activated Na-form) which has been shown to have NH4+ an exchange capacity of 29 mg N-NH4+ g−1 in single NH4+ solutions and 23 mg N-NH4+ g−1 in treated wastewater simulating up-concentration effluent at pH 8 [64]. However, most of the latest research on NH4+ removal in the ion exchange process is limited to laboratory-scale analyses, and to the best of our knowledge, only one full-scale application has been performed worldwide in the last half century. This is the Battelle Northwest/South Tahoe Public Utility District treatment unit, created to protect the quality of the water of the Truckee River in California. This unit includes a 22,500 m3 d−1 facility in California for the removal of NH4+ from domestic wastewater and is the only industrial facility using natural zeolite in the world [65]. However, unfortunately no data are available in literature regarding the performance of such full-scale treatment plant. The transition from batch testing to full testing is of great importance because allows to assess other parameters in the NH4+ adsorption by zeolite. For example, the mechanical strength of the medium that is considered to be the key characteristics to prevent the breakdown and disintegration of the medium during the ion exchange process. The mechanical strength of the zeolites refers to the resistance of the media to friction and compression. Friction resistance is correlated with turbidity measurements, following the principle that increasing turbidity was related to media disintegration over time. This parameter was recently studied by Guida et al. [66]. In their batch study, six synthetics (named progressively from Zeolite1 to Zeolite6), one natural (clinoptiolite), and one engineered (Zeolite-N) zeolites were analyzed. They observed that after the first 2 hours of stirring, the disintegration values ranged from 17.5 NTU h−1 to 2.7 NTU h−1. Clinoptilolite showed the best friction resistance together with other two natural zeolites. To evaluate compression, the different zeolites were subjected to increasing load pressure and recording the force at the breaking point. Zeolite-N resisted to a pressure up to 7.9 N before

Chapter 5 • Zeolites for the nutrient recovery from wastewater  105

fracture, synthetic zeolites 4 and 5 resisted a load up to 11.3 N and 16.5 N, respectively, for the other synthetic zeolites the average load before fracture was between 3.9 N and 5.7 N. Among all, clinoptilolite showed the highest compressive strength with a loading pressure of 38.6 N before fracture. This variability in the mechanical strength of synthetic zeolites was attributed to the difference in the manufacturing process (temperature, time, and furnace size during calcination), which could explain the higher strength of natural zeolite. Recently, zeolites have been investigated also to recover PO43− from wastewater. Natural zeolites usually have little or no affinity for anions such as PO43- because of its net negative charge [67]. Therefore, natural zeolite may demonstrate only low adsorption capacity for PO43−[68]. To overcome this limit, several studies have analyzed the possibility to modify zeolites to improve their ability in recovering PO43− from water. Onyango et al. [9] studied the removal of inorganic PO43− from aqueous solutions by improving the affinity of zeolite for PO43− through its activation by aluminum salt solution. In their study, they used the synthetic zeolite HSZ 330 HUD (Si/ Al ratio: 2.75–3.25) and its Al3+-activated form (Al HUD). Equilibrium and kinetic experiments were performed to investigate the effects of operating conditions such as adsorbent mass, solution pH, coexisting ions, and initial PO43− concentration on the ability or rate of PO43− adsorption by zeolites. Activated zeolite was prepared by adding 50 g of HUD to 1 L of 0.075 M aluminum sulphate solution, the mixture was stirred for 2 days and then washed several times with demineralized water to lower the electrical conductivity. Finally, the Al3+ -activated HUD zeolite was air-dried at room temperature for 2 days. The ability of the HUD and Al-HUD zeolites to remove PO43− was tested by varying the masses of the zeolite at a fixed PO43− concentration of 100 mg L−1, at room temperature of 25°C and, pH of 5.7. Zeolite activated with Al3+ (Al-HUD) showed higher removal efficiency; indeed, 150 g of Al-HUD lowered the PO43− concentration below 10 mg L−1, whereas HUD zeolite was not able to low the PO43− concentration below 20 mg L−1 under the same experimental conditions. Furthermore, it was observed that the efficiency of PO43− removal by zeolite increased by increasing the mass of the adsorbent and decreasing the pH of the solution. Other studies have focused on the simultaneous recovery of N and P from wastewater by zeolite. In a study conducted by Karapinar [69], natural zeolite was studied in lab scale using a batch system to remove NH4+ and PO43− in two steps. Specifically, zeolite with an average particle size of 13 μm was used as adsorbent for NH4+ and then as a seed material for the precipitation of calcium phosphate [Ca3(PO4)2]. The Author found that zeolite and NH4+ did not affect the amount of calcium phosphate precipitated. On the other hand, phosphate precipitation increased with rising pH and was achieved at low super saturations (pH 6.9–7.5) through secondary nucleation and crystal growth. Thus, this study pointed out that zeolite can be appropriately used as an adsorbent/seed material for the removal of both NH4+ and PO43− nutrients from wastewater. Lin et al. (2014) in their study for the first time used natural zeolite in batch and continuous tests to simultaneously remove P and N from orthophosphate and ammonium-nitrogen laden wastewaters with a pH range from 3–11. They used an unmodified Chinese natural zeolite formed from clinoptilolite-Na, heulandite, and quartz, with a specific surface area of 14.33 m2 g−1, sieved to 0.8–1.43 mm, with a CEC of 0.092 cmol(+) g−1. They found that P removal was efficient only at pH > 9, decreasing from 102 to 14.5 mg L−1. Moreover, when ammonium was also present, P removal was enhanced by about 60%. They suggested that

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ammonium was first adsorbed onto the zeolite via ion exchange to release Ca2+ ions from the zeolite lattice [70]. This ion exchange reaction was independent on the presence of P. Then the released Ca2+ ions precipitated the dissolved phosphate, which was greatly dependent on the concentration of ammonium. More recently You et al. [71] investigated the performance of three synthetic Na-zeolites, modified in Ca- and Mg-zeolites, for the simultaneous removal and recovery of NH4+ and PO43− from synthetic solutions, simulating wastewater treatment effluent. The simultaneous removal of NH4+ and PO43− can occur through two mechanisms: 1. ion exchange and formation of calcium phosphate minerals in the case of Ca-zeolite, and 2. precipitation of struvite (MgNH4PO4) in the case of Mg-zeolite. Results suggested that the activation of Na-zeolite in calcium and magnesium forms improved the PO43− adsorption capacity of zeolite. The maximum adsorption capacities obtained were 123 mg NH4 g−1 and 119 mg PO4 g−1 in an ammonium/phosphate binary system for Ca-zeolite, whereas Mg-zeolite adsorbed while 55 mg NH4+ g−1 and 60 mg PO43− g−1. However, the sorption process was faster for Mg-zeolites and the sorption mechanism that controls the overall process, for both zeolites, was diffusion into the particle.

5.5.2  Regeneration of zeolites Once saturated, the zeolite can be regenerated or used as slow-release fertilizer. Regeneration is the biggest obstacle to the widespread use of zeolite-based wastewater treatment on a large scale; indeed, it is estimated to account for up to 80% of the total operating cost of zeolite use, which is why it is necessary to develop a method that is economically feasible [65]. For this reason, it is crucial to assess the potential applicability of a zeolite for industrial or large-scale applications and to study the possibility of its regeneration in order to consider its use in multiple cycles. Regeneration allows the zeolite to be reused for several cycles until when the adsorption capacity decrease towards levels no more sustainable from an economic point of view. Regeneration can be carried out by various techniques such as chemical, acid, heating, and biological regeneration [72]. However, the most used regeneration methods are chemical and biological ones. Chemical regeneration is commonly achieved using high strength ionic brines, for example, aqueous sodium chloride (NaCl), where Na+ ion replaces the NH4+ adsorbed liberating it into the liquid phase [65,73]. The desorption and regeneration are based on the following cation exchange equation [74]:

NH 4 - zeolite + Na + → Na - zeolite + NH 4 + (5.4)

The concentrated regenerating solution, after increasing its pH above 9.3 (pKa of the NH4+-NH3 system), can be stripped from the air of its ammonia content and the brine can be recovered [75]. The NH3 gas can be further absorbed into a solution of H2SO4 or HNO3. In addition, the H2SO4 solution absorbed by NH3 could be used as a fertilizer. Widiastuti et al. [15] studied the desorption and regeneration of zeolite in a batch experiment using a 1% NaCl solution. Their results showed that the NH4+ desorption was up to 97% depending on contact time, zeolite loading, initial NH4+ concentration and pH of the solution and that the adsorption and desorption efficiency decreased by increasing the cycle numbers. In fact, the adsorption of NH4+ on the zeolites decreased from 4.49 mg g−1 in the first cycle

Chapter 5 • Zeolites for the nutrient recovery from wastewater  107

to 3.71  mg g−1 in the third cycle, whereas the desorption decreased from 4.27 mg g−1 in the first cycle to and 3.57 mg g−1 in the third. However, although NH4+ is selectively adsorbed by zeolite, its desorption was sufficiently high following the treatment with NaCl that is the most widely used method mainly for the low cost of the NaCl. Also, Li et al. [76] carried out column experiments and batch tests to test the efficiency of silicate-carbon modified zeolite (SCMZ) in removing NH4+ from drinking water with a concentration ranging from 2.5 mg NH4+ L−1 to 10 mg NH4+ L−1 and how this efficiency varies when regenerated by NaCl solution. They observed that the NaCl regeneration method had good regeneration capacities for NH4+ adsorption of SCMZ. Indeed, NH4+ adsorption capacity was 0.1117 mg L−1, at the first regeneration cycle and 0.11047 mg L−1 at the third regeneration cycle, both values close to those of fresh SCMZs (0.1155 mg L−1). Finally, Deng et al. [77] conducted both batch and column test to recover NH4+ from zeolite with NaCl solution at different pH (9–12) and salt concentration (20–160 g L−1). In the batch study, they achieved 95% of regeneration efficiency after only 1 h at NaCl concentration of 80 g L−1. Interestingly, the regeneration efficiency was 96% at the low NaCl at pH 12 in 2 h. Results from the column study, on the other hand, suggested that an optimum NaCl concentration for regeneration of exhausted zeolite determined in batch tests cannot represent that in continuous zeolite columns, due to different volumes. Indeed, at pH 9, the maximum RE was 76% at NaCl concentration of 160 g L−1, while it was only 38% and 13%, respectively, at NaCl concentration of 80 and 20 g L−1. The Authors [77] attributed the low regeneration efficiency of NH4+ in the continuous column compared with in batch studies to the small mass ratio of Na+ to Zeolite-NH4+−N in the column during regeneration. Thus, based on results of Deng et al. [77] the maximum regeneration efficiency (85%) of zeolite in column test can be achieved within 2 h of reaction time at pH 12 using low NaCl dosages from 10 g to 40 g of NaCl L−1. Besides to NaCl, also other salts can be used to recover NH4+ from zeolite. Czárán et al. [78], for example, suggested KCl solution at pH ranging from 9 to 11 to regenerate zeolite. Indeed, KCl should be preferred to NaCl due to the much higher selectivity of clinoptilolite toward K+ than Na+; on the other hand, it is less used due to its cost. On the other hand, Metcalf et al. [79] recommended calcium hydroxide for the regeneration of zeolite, although this method waste time, energy, and chemical reagents, causing secondary pollution that requires further treatment. The direct result of zeolite regeneration is the desorption of NH4+, which can be recycled and used for other purposes, such as fertilizer. In some cases, however, if the aim is not to reuse the zeolite, it may be possible to desorb the nutrients without regenerating the zeolite itself. Cyrus and Reddy [80] proved that adsorbed NH4+ can be desorbed by washing the zeolite with water for 250 hours; however, a large amount of N (>20%) is retained in the zeolite column. This experiment suggested that NH4+ enriched zeolite can be a suitable medium to slowly release N in the soil. An alternative to the chemical regeneration is the biological regeneration, a two-mode process consisting of aerobic nitrification and further processing of NO3− products [81]. Biological regeneration occurs putting in contact the NH4-zeolite with nitrifying bacteria. In this system, a solution containing cations is recirculated through the bed of zeolite to desorb NH4+ into the solution:

Z - NH 4 + + Na + ↔ Z - Na + + NH 4 + (5.5)

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The effluent containing the displaced NH4+ is then oxidized to NO3− by nitrifying bacteria in presence of oxygen.

NH 4 + + 2O2 ↔ NO3 − + 2H + + H2 O (5.6)

The NO3− brine, after a clarification process, is easy to dispose of or can be mixed with raw wastewater and denitrified into harmless N gas [82].

5.5.3  Reuse of enriched zeolites Zeolite cannot be regenerated indefinitely. After a certain number of regeneration cycles, they are exhausted. Several studies have been performed with the purpose of evaluating after how many cycles of use the ammonium exchange capacity (AEC) of regenerated zeolites is significantly reduced, compared to virgin zeolite. Deng et al. [77] evaluated the AEC of four natural zeolites, breakthrough time, in a continuous zeolitic column fed with anaerobic membrane bioreactors permeates and NH4Cl solution to optimize the operating conditions and evaluate the effect of competitive cations on the AEC. Regeneration was performed using 1% NaCl solution at pH 12. They observed that the AEC decreased by 13% after the first cycle (from 3.1 mg N g−1 to 2.7 mg N g−1 zeolite). After the third and fourth cycles the AEC was reduced by 12.5% compared with virgin zeolite. After the fifth cycle, the AEC started to decrease until it reaches a plateau of about 1.8 mg N g−1 zeolite after the eighth cycle. Thus, the regeneration efficiency of the first five cycles were almost constant at 84% and then slowly decreased to about 70%. In another study, Guida et al. [66] evaluated the regeneration capacity of zeolites over multiple cycles using wastewater. Of the eight zeolites analyzed, including six synthetics (Zeolite1, Zeolite2, Zeolite3, Zeolite4, Zeolite5, Zeolite6), one natural (clinoptilolite), and one engineered (Zeolite-N), the latter showed the highest AEC, with a 6% decrease between the first and tenth cycles. Zeolite-N also presented the highest regeneration capacity which increased by 7% between the first and tenth cycles, achieving a regeneration efficiency of 90–100%. For clinoptilolite and Zeolite2, after 10 cycles, the AEC decreased by 15% and 3%, respectively, and the regeneration capacity also decreased by 11% and 3%, respectively, while no clear pattern in the variation of AEC and Qreg was identified for the other zeolites. Non-regenerated enriched zeolite, however, can be reused for other purposes, such as slowrelease fertilizer in agriculture, in which case, the enriched zeolite is applied directly to the soil. Based on the high affinity of Ca/Mg zeolites for NH4+ and PO43− and the slow release of NH4+ and PO43−, especially when removed through the formation of Ca and Mg-phosphate minerals. You et al. [71] suggested the use of these materials as potential nutrient carriers for soil quality improvement. Several studies have been conducted to investigate this possible alternative use of zeolite. Kocatürk-Schumacher [83] conducted a pot experiment with ryegrass to test the ability of zeolite clinoptilolite enriched with digestate nutrients to supply nitrogen when used as fertilizer. Clinoptilolite had particle size of 1–3 mm, pH of 8.5, with 150–210 cmol(+) kg−1 cation exchange capacity, and 35.5 m2 g−1 specific surface area. Nutrient‐enriched clinoptilolite, and quartz sand, was used to fill in pots to have 15 or 45 mg N per pot. Pots with the lowest N resulted in more than a twofold increase in yield and more than a threefold increase in N uptake when compared to the control (untreated) clinoptilolite treatment. The increases were even

Chapter 5 • Zeolites for the nutrient recovery from wastewater  109

Table 5.3  Plant growth indicators take in consideration in the different treatments by Guaya et al. [6]. NEZ, nutrient enriched zeolite; CS, clay soil; SS, sandy soil. Plant moisture Evapotranspiration content The incorporation of NEZ systems did not modify the NEZ-Mn evapotranspiration on both clayey and NEZ-Fe sandy amended soils. NEZ-Al

The addition of the NEZ system changed the plant water contents of sunflowers grown on clay soils but for sandy soils the water content was the same in all treatments.

Plant biomass CS SS +269% +77%

Nitrogen and Root/ Phosphorus branch ratio uptake

CS SS CS –47% +34% +19% N +110%P +401% +131% –13% +17% +44% N +77% P +95% +50% –21% –12% +43% N +54% P

SS +17% N +210% P +45% N +174% P +5% N +140% P

C/N ratio CS –1%

SS –10%

–21% –38% –24% –10%

greater when nutrient‐enriched clinoptilolite was added to reach 45 mg N per pot, with almost a fourfold and sixfold increase in yield and N uptake. Guaya et al. [6] activated a natural clinoptilolite in its potassium form with hydrated metal oxides (Fe, Mn, Al) and used it to recover nutrients from wastewater. Then, the nutrientenriched zeolites were used as a soil conditioner in sandy and clay soils for the growth of sunflower (Helianthus annuus) plants. They observed that enriched zeolites as soil conditioner improved sunflower production by more than 50% (Table 5.3), thus confirming the reasonable use of enriched zeolites as a slow nutrient release avoiding, at the same time, nutrient leaching, and the consequent environmental issue (Table 5.3).

5.6  Conclusions and perspectives Due to its physical and chemical properties, natural abundance and low cost, zeolite is widely used for its catalytic abilities, cation exchange capacity, and application potential in various industrial processes. Due to the high CEC, zeolite can be used for the recovery of nutrients, for example, NH4+, from treated wastewater. In addition, zeolite can also be modified to recover negatively charged nutrients such as PO43−. Furthermore, zeolite can also be used as a soil conditioner to improve the physic-chemical characteristics of soils, such as nutrient retention capacity. Finally, the application of nutrient-enriched zeolite can be a suitable option as a source of nutrients for plants in a circular economy perspective. The use of zeolites is set to increase significantly due to the various fields in which they can be used. It should be mentioned that there are still immediate challenges for further successful environmental applications of this interesting class of natural materials. Future research should find sustainable methods to increase the operating capacity and mechanical strength of natural zeolites to enable their use in large-scale plants. Therefore, it is essential to find zeolites with a good adsorption capacity and capable of maintaining this capacity after several regeneration cycles, thus ensuring the sustainability of the process. Furthermore, after their depletion, enriched

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zeolites can be applied to the soil as slow-release fertilizers. The possible use as large-scale fertilizers may increase their possibilities for environmental application but has to be deeply investigated.

Acknowledgments This work was funded by the Project “Achieving wider uptake of water-smart solutions—WIDER UPTAKE” (Grant Agreement number: 869283) financed by the European Union’s Horizon 2020 Research and Innovation Programme, in which the first author, Prof. Giorgio Mannina, is Principal Investigator for the University of Palermo; local project website: https://wideruptake.unipa.it/.

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[65] Canellas, J (2018) Tertiary ammonium removal with zeolites. School of Water, Environment and Energy Research Degree, Cranfield University. [66] Guida S, Potter C, Jefferson B, Soares A. Preparation and evaluation of zeolites for ammonium removal from municipal wastewater through ion exchange process. Sci Rep 2020;10(1):1–11. [67] Huang H, Xiao D, Pang R, Congcong H, Ding L. Simultaneous removal of nutrients from simulated swine wastewater by adsorption of modified zeolite combined with struvite crystallization. Chem Eng J 2014;256:431–8. [68] Markou G, Depraetere O, Vandamme D, Muylaert K. Cultivation of Chlorella vulgaris and Arthrospira platensis with recovered phosphorus from wastewater by means of zeolite sorption. Int J Mol Sci 2015;16(2):4250–64. [69] Karapinar N. Application of natural zeolite for phosphorus and ammonium removal from aqueous solutions. J Hazard Mater 2009;170(2–3):1186–91. [70] Lin L, Wan C, Lee DJ, Lei Z, Liu X. Ammonium assists orthophosphate removal from high-strength wastewaters by natural zeolite. Sep Purif Technol 2014;133:351–6. [71] You X, Valderrama C, Cortina JL. Simultaneous recovery of ammonium and phosphate from simulated treated wastewater effluents by activated calcium and magnesium zeolites. J Chem Technol Biotechnol 2017;92(9):2400–9. [72] Sengupta S, Nawaz T, Beaudry J. Nitrogen and phosphorus recovery from wastewater. Curr Pollut Rep 2015;1(3):155–66. [73] Koon JH, Kaufman WJ. Ammonia removal from municipal wastewaters by ion exchange. J (Water Pollut Control Federation) 1975;47(3):448–65. [74] Du Q, Liu S, Cao Z, Wang Y. Ammonia removal from aqueous solution using natural Chinese clinoptilolite. Sep Purif Technol 2005;44:229–34. [75] Liberti L, Petruzzelli D, Florio LDe. REM NUT ion exchange plus struvite precipitation process. Environ Technol 2001;22(11):1313–24. [76] Li M, Zhu X, Zhu F, Ren G, Cao G, Song L. Application of modified zeolite for ammonium removal from drinking water. Desalination 2011;271(1–3):295–300. [77] Deng Q, Dhar BR, Elbeshbishy E, Lee HS. Ammonium nitrogen removal from the permeates of anaerobic membrane bioreactors: economic regeneration of exhausted zeolite. Environ Technol 2014;35(16):2008–17. [78] Czárán E, Mészáros-Kis Á, Domokos E, Papp J. Separation of ammonia from wastewater using clinoptilolite as ion exchanger. Nucl Chem Waste Manag 1988;8(2):107–13. [79] Metcalf L, Eddy HP, Tchobanoglous G. Wastewater engineering: Treatment, disposal, and reuse. New York McGraw-Hill; 1991. [80] Cyrus JS, Reddy GB. Sorption and desorption of ammonium by zeolite: batch and column studies. J Environ Sci Health - Part A Toxic/Hazard Substances Environ Eng 2011;46(4):408–14. [81] Wu Z, An Y, Wang Z, Yang S, Chen H, Zhou Z, et al. Study on zeolite enhanced contact–adsorption regeneration–stabilization process for nitrogen removal. J Hazard Mater 2008;156(1):317–26. [82] Rahmani AR, Mahvi AH, A.H.M. Use of ion exchange for removal of ammonium: a biological regeneration of zeolite. Pak J Biol Sci 2004;8(1):30–3. [83] Kocatürk-Schumacher NP, Zwart K, Bruun S, Stoumann Jensen L, Sørensen H, Brussaard L. Recovery of nutrients from the liquid fraction of digestate: use of enriched zeolite and biochar as nitrogen fertilizers. J Plant Nutr Soil Sci 2019;182(2):187–95.

6 Wastewater treatment sludge composting Sofia Maria Muscarellaa, Luigi Badaluccoa, Vito Armando Laudicinaa, Zhiwei Wangb, Giorgio Manninac a

DEPARTMENT OF AGRICULTURAL, FOOD AND FOREST SCIENCES, UNIVERSITY OF PALERMO, PALERMO, ITALY b TONGJI UNIVERSITY, SHANGHAI INSTITUTE OF POLLUTION CONTROL AND ECOLOGICAL SECURITY, STATE KEY LABORATORY OF POLLUTION CONTROL AND RESOURCE REUSE, SCHOOL OF ENVIRONMENTAL SCIENCE AND ENGINEERING, SHANGHAI, CHINA c ENGINEERING DEPARTMENT, PALERMO UNIVERSITY, PALERMO, ITALY

6.1 Introduction In recent years, the amount of sewage sludge generated by wastewater treatment plants (WWTPs) has increased due to worldwide population growth and to efficiency of biological treatment processes [1,2]. Sludge is an important source of secondary pollution to aquatic environments and a potential risk to human health; moreover, it represents one of the most important cost items in the functioning of water treatment plants [3–5]. About 60% of the operating costs of secondary wastewater treatment plants in Europe can be associated with the treatment and disposal of products [6]. For this reason, proper sludge management becomes increasingly important, at both national and international level, and it becomes necessary to find effective measures to limit the environmental impacts and to reuse sludge as a resource, within a circular economy vision [2,7]. Current methods of utilization of sewage sludge include agricultural application, landfilling, incineration, drying, and composting and/or vermicomposting. Composting is a widely used cost-effective and socially acceptable method for treating solid or semisolid biodegradable waste [8]. In agriculture sewage sludge is used for rehabilitation of degraded soils, reclamation, or adaptation of land to specific needs [9]. The above consideration comes from several studies showing that the application of sludges to agricultural land can improve soil fertility and, therefore, crop productivity [10–12]. This field of use is also possible due to its composition; in fact, it is rich in organic matter, nitrogen, phosphorus, calcium, magnesium, sulfur, and other microelements needed by plants and living native organisms in the soil. However, sewage sludge may contain a wide range of harmful toxic substances such as heavy metals, polycyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzo-p-dioxins and dibenzo-p-furans, polychlorinated biphenyls, di(2-ethylhexyl) phthalate, polybrominated diphenyl ethers, detergent and drug residues, pharmaceutical and personal care products (PPCPs), endogenous hormones, synthetic steroids and pathogenic organisms [13,14], which Current Developments in Biotechnology and Bioengineering. DOI: https://doi.org/10.1016/B978-0-323-99920-5.00008-1 Copyright © 2023 Elsevier Inc. All rights reserved.

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can cause harm to the environment and humans. Due to the presence of those toxic elements, stabilization of sewage sludge is necessary to avoid any environmental risk [15]. Stabilization of sewage sludge is defined as “biological, chemical or thermal treatment, long-term storage or any other appropriate process aimed at reducing its fermentability and the health hazards arising from its use” [16]. This definition is found in Council Directive 86/278/EEC, which was issued to regulate the use of sludge in agriculture, the primary objective of which is the environment, in particular the soil, and the protection of human health. European Directive 86/278/EEC was implemented in Italy by Legislative Decree 99/1992 [17]. Both the European Directive and the Italian legislative decree can be considered obsolete, this is why the European Union is moving towards amending them to reflect the new needs of the sector and to keep up with technological innovations. Currently, there are several processes for sludge stabilization, including composting, which is one of the most widely used methods for stabilizing organic matter in general, reducing the number of pathogenic microorganisms and the amount of toxic elements [18]. This is possible because during the composting process the organic compounds present in the biomass to be composted are converted into chemically recalcitrant, that is, stabilized, humic substances, while pathogens are eliminated due to the heat generated during the process thermophilic phase [19,20]. During the composting of sludges, the addition of bulking agents is needed, as they ameliorate the composting performance by providing structural support that improves aeration and regulates moisture content and C/N ratio of composting mass [21,22]. Sludge composting, however, has to be focused on limiting some secondary causes of pollution related to the process itself, such as greenhouse gas (GHG) emissions and heavy metal contamination [23]. Indeed, in the last decades, the handling of sewage sludge with traditional methods has led to the release of an enormous amount of greenhouse gases. The choice of an appropriate bulking agent is, therefore, fundamental to limit the emission of climate-altering gases, and, at the same time, to increase the microbial activity thus improving the quality of the compost [24,25]. This chapter aims (1) to give an overview of the national and international legislation on sludge management and reuse, (2) to analyze the composting process and the state of the art regarding sludge composting to understand the limitations at large-scale application, and (3) to discuss the technological innovations in the field and highlight future perspectives.

6.2  Legislation about sewage sludge The treatment and disposal of sewage sludge is a very important issue from both national and international point of view, especially in relation to the risk of secondary pollution resulting from its mismanagement [26]. Over the last 30 years, the EU has extensively regulated the management and reuse of sludge by various legislative instruments and acts, as this aspect is part of the EU sustainability vision in an environmentally safe approach, also in relation to the rapidly increasing amount of sludge produced [27]. On this basis, EU Member States have transposed European directives and applied more stringent parameters in relation to their own conditions, particularly in relation to heavy metals and pathogens [27,28]. This section of the chapter will provide an overview of the European directives and Italian legislative decrees regarding sludge management.

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6.2.1  European legislation The most significant European directives related to sludge management and reuse in agriculture are Directive 86/278/EEC, called “Sewage Sludge Directive” (SSD) [16], Directive 91/271 [29], Directive 91/676/EEC [30], Directive 2008/98/EC [31], Directive 2018/851/EC [32] and Regulation (EU) 2019/1009 [33], which establish standards for making EU fertilizer products available on the market. These are complemented by the Public Consultation [34] launched in 2020 by the EU with the aim of renewing Directive 86/278/EEC. The Directives and the European Regulation are briefly described below (Table 6.1):

6.2.1.1  Directive 86/278/CEE The main Directive regulating the management of sewage sludge is number 86/278/EEC [16], which concerns the protection of the environment, in particular of the soil, when sewage sludge is used in agriculture. This Directive regulates the use of sewage sludge as a fertilizer in such a way as to avoid harmful effects on the environment and human health and considering the nutrient needs of plants, without compromising the quality of the soil and of surface or groundwaters. To this end, it establishes limit values for permitted concentrations in the soil for seven heavy metals that may be toxic to plants and humans, that is, Cd, Cu, Ni, Pb, Zn, Hg, and Cr. In other words, Directive 86/278/EEC [16] prohibits the use of sewage sludge when the concentration of metals exceeds threshold values. ANNEX IA contains the limit values of heavy metals in soils, ANNEX IB regulates the maximum amount of heavy metals in sludge and ANNEX IC the maximum annual amounts of heavy metals that can be released into the soil (Table 6.2) [16]. The main points of the Directive stipulated that sludge must undergo a stabilization process, such as composting, before being used in agriculture. However, in some EU countries farmers might be allowed to use untreated sludge if it was injected or buried in the soil.

Table 6.1  European Directives and Regulations concerning the management and disposal of sludge. Directives or Regulations

Objective

Directive 86/278/EEC [16]

Concerning the protection of the environment, in particular of the soil, in the use of sewage sludge in agriculture. On urban wastewater treatment. The aim of this Directive is the protection of the EU environment from the negative consequences of pollution caused by urban wastewater, such as eutrophication. Concerning the protection of waters against pollution caused by nitrates from agricultural sources. Waste Framework Directive. Concerning the waste management and abrogating some previous directives. Repeals Directive 2008/98/EC, making significant changes. Laying down rules on placement on the market of EU fertilizing products, amending Regulations (EC) No 1069/2009 and (EC) No 1107/2009 and repealing Regulation (EC) No 2003/2003

Directive 91/271/EEC [29]

Directive 91/676/EEC [30] Directive 2008/98/EC [31] Directive 2018/851/EC [32] Regulation 2019/1009 [33]

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Table 6.2  Limit values given in ANNEX IA, ANNEX IB, and ANNEX IC of Directive 86/278/EEC [16]. Parameters

Limit values – ANNEX IA (mg/kg)

Limit values – ANNEX IB (mg/kg of dry matter)

Limit values – ANNEX IC (kg/ha/y)

Cadmium (Cd) Copper (Cu) Nickel (Ni) Lead (Pb) Zinc (Zn) Mercury (Hg) Chromium (Cr)

1 to 3 50 to 140 30 to 75 50 to 300 150 to 300 1 to 1.5 –

20 to 40 1000 to 1750 300 to 400 750 to 1200 2500 to 4000 16 to 25 –

0.15 12 3 15 30 0.1 –

As mentioned above, EU countries since implementing Directive 86/278/EEC [16] have introduced stricter limits for sludge use in agriculture. Of these, 18 out of 27 countries have introduced restrictions on cadmium (Cd), 14 out of 27 countries on copper (Cu), 19 out of 27 countries on mercury (Hg), 16 out of 27 countries on nickel (Ni), 14 out of 27 countries on lead (Pb), 10 out of 27 countries on zinc (Zn) [16]. In the countries that adopted more restrictions the result was a lower percentage of sludge applied in agriculture [13]. In addition, several countries introduced limit values for other elements, for example, for chromium whose limits were introduced in 23 countries (up to 1500 mg kg−1 sludge dw), 7 countries introduced limits for arsenic (up to 75 mg kg−1 dw), Romania and Hungary for molybdenum (50 mg kg−1 dw), and Hungary also added limits for cobalt and selenium (20 and 100 mg kg−1 dw, respectively). Finally, Hungary have introduced limits for Cr VI (hexavalent chromium) [13,35], 1 mg per kg of dry matter Directive 86/278/EEC [16] does not include limit values or special requirements for organic micropollutants and pathogen content in biosolids, but the different States have added limit values in national regulations. The organic compounds most controlled by the different States are PCBs (polychlorinated biphenyls), AOXs (absorbable organic halogens), and PAHs (polycyclic aromatic hydrocarbons). Regarding the content of pathogens, the national legislation of most countries controls the presence of Salmonella spp. (except for Lithuania, Luxembourg, and Slovakia) [10].

6.2.1.2  Revisions of Directive 86/278/EEC From 16th June to 25th June 2020, the European Union created an initiative with the aim of encouraging citizens, stakeholders, and plant operators at European level to give their opinion on the roadmap created for the reformulation and modification of the Sewage Sludge Directive [16]. Subsequently, from 20th November 2020 to 5th March 2021 a “Public Consultation” was launched with the aim of collecting opinions from stakeholders, operators, and experts in the field at European level. These initiatives arise from the need to update Directive 86/278 EEC, which has been in force for over 30 years and no longer meets the needs and requirements of the industry, also in the light of scientific and technological evolutions, and today’s needs, such as an adequate regulation of pollutants in sludge, especially “emerging contaminants,” for example, organic chemicals like pharmaceuticals, PAHs and perfluoroalkylates, cosmetics and microplastics [34].

Chapter 6 • Wastewater treatment sludge composting  119

This initiative will evaluate the effectiveness of the Directive and analyze the risks and opportunities associated with managing sewage sludge in agriculture [16,34].

6.2.1.3  Regulation (EU) 2019/1009 Regulation (EU) 2019/1009 [33], which will come into force on 16th July 2022, establishes rules regarding the making available on the market of EU fertilizer products. It covers seven categories of fertilizer products: fertilizers, soil conditioners, lime and/or magnesium correctives, growing media, inhibitors, plant biostimulants and physical mixtures of fertilizer products. The aim of the Regulation is the creation of a single market for fertilizer products that are currently not regulated by standards, thus creating common safety, quality, and labelling standards [33]. With this regulation, for the first time, limit values for organic contaminants are introduced, allowing for a high level of soil protection and a reduction in risks to human health and the environment [33]. It is divided into 11 CMCs (Categories of Constituent Materials) and 7 PFCs (Functional Categories of Products). Sewage sludge is treated in CMC 3 and CMC 5, which state that it is no longer possible to produce a fertilizer by aerobic composting (CMC 3) or by anaerobic digestion (CMC 5) from sewage sludge, industrial sludge or dredging sludge [33].

6.2.2  Italian legislation In Italy, the most important Legislative Decree on this issue is the number 92/99 [17], which is the implementation in Italy of the Sewage Sludge Directive [16]. Other decrees that regulate sludge management are the Legislative Decree number 152/2006 [36] amended by the Legislative Decree 205/2010 [37] which deals with the activity of recovery of sludge to produce soil improvers, the Legislative Decree 75/2010 [38] on fertilizers and art. 41 of the Legislative Decree No. 109/2018, called “Decreto Genova,” on the subject of “Urgent provisions on the management of sewage sludge” [39]. The Legislative Decrees are briefly described in Table 6.3. Table 6.3  Summary table of Italian Legislative Decree concerning the management and disposal of sludge in Italy. Legislative Decree

Objective

Legislative Decree 99/1992 [17]

Implementation of Directive no. 86/278/ EEC concerning the protection of the environment, in particular of the soil, in the use of sewage sludge in agriculture. This Legislative Decree is called the “Testo Unico Ambientale”, which contains the main rules that regulate the environmental discipline. Provisions on the reorganization and revision of regulations on fertilizers. Provisions implementing Directive 2008/98/EC of the European Parliament and Council on wastes and repealing certain Directives. Urgent provisions on the management of sewage sludge.

Legislative Decree 152/2006 [36] Legislative Decree 75/2010 [38] Legislative Decree 205/2010 [37] Art. 41 of the Legislative Decree No. 109/2018 [39]

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6.2.2.1  Legislative Decree 99/92 The Legislative Decree no. 99 of the 27th January 1992 [17] is the implementation of Directive no. 86/278/EEC [16] and aims to regulate the use of sewage sludge in agriculture to avoid harmful effects on soil, vegetation, animals, and humans, while encouraging its proper use. The Legislative Decree 99/92 [17] requires that sludge be subjected to treatment (biological, chemical, or thermal treatment, long-term storage, or other appropriate process) suitable to produce a fertilizing and/or soil amendment and corrective effect, and not containing toxic and harmful and/or persistent and/or biodegradable substances in concentrations harmful to the soil, crops, animals, humans and the environment in general [17]. It confirms the limit values for heavy metals in both soil and sludge set by ANNEX IA and ANNEX IB of the Directive 86/278/EEC [16]. However, it adds a restriction not foreseen in the European Directive, namely the obligation to subject sewage sludge to chemical, physical or biological treatment with the aim of stabilizing the sludge by reducing the amount of pathogenic organisms present [17].

6.2.2.2  Art. 41 of the Legislative Decree No. 109/2018 The Legislative Decree of the 28th of September 2018 No. 109 [39], also known as the “Decreto Genova” is a Decree-Law concerning urgent provisions for the city of Genova and the safety of the national network of infrastructure and transport following the seismic events of 2016 and 2017, but also other emergencies [39]. Art. 41 of the aforementioned legislative decree contains a reference to the use of sewage sludge, pending a systematic and long-awaited revision of the legislation currently in force, with the introduction of new limits on C10-C40 hydrocarbons (whose limit value has been raised from 50 mg kg−1 SS to 1000 mg kg−1 SS) and on heavy metals (Toluene ≤100 mg kg−1 SS, Selenium ≤10 mg kg−1 SS, Beryllium ≤2 mg kg−1 SS, Arsenic PSB > GSB. Furthermore, unlike higher forms of plant and algae, PB have neither chloroplasts nor endothecium. Photosynthesis occurs at the chromatophore in which abundant bacteriochlorophylls and Carotenoids on the inner folded cell membrane [238,239]. PSB and PNSB have a more complete intracytoplasmic membrane system for harnessing light energy than GSB and GFB, which only contain special structures called chlorosomes to perform light harvesting tasks (Table 15.4) [238–240]. Those special structures enable them to have flexible materials and energy metabolic pathways, which is also the main reason for their high efficiency in purifying wastewater. The diverse range of potential PSB applications arises from the exploitation of their metabolic versatility, as shown in Figs. 15.3–15.5(A). PSB grow under anaerobic conditions by: (1) photoautotrophy, using light for anabolism and CO2 as a carbon source with a range of inorganic electron donors; (2) photoheterotrophy, using light as energy source and organic carbon as carbon source; (3) fermentation, without light and using organics as energy and carbon source. PSB have the ability to utilize a range of electron donors, for example: H2, H2S, S0, S2O32−, Fe2+, NO2−, and CO for autotrophic metabolism [241–244]; and acetate, propionate, butyrate, malate, succinate, lactate, dimethylsulfide, glucose, sucrose, lactose, and ethanol for heterotrophic metabolism [235,241,245–247]; and (4) PSB can grow aerobically through respiration and fix dinitrogen gas via nitrogenase. Some bacteria like Rhodobacter sphaeroides and Rhodopseudomonas palustris were documented to perform nitrification/denitrification and grow lithoautotrophically, for instance, via halophilic S2− oxidation [248–250]. As shown in Fig. 15.10(B), the environmental conditions (e.g., electron acceptor and electron donor present and presence/absence of light or oxygen) determine the predominant metabolic growth mode. In the presence of light and absence of oxygen, the anoxygenic phototrophic

Chapter 15 • Biological nutrient recovery from wastewater for circular economy  381

Table 15.4  Schematic classification and characteristics of PB.

PSB Typical habitats

Form and motility

Suitable pH

Purple nonsulfur bacteria (Rhodospirillaceae, Purple sulfur bacteria PNSB) (Chromatiaceae, PSB)

Green sulfur bacteria (Chlorobiaceae, GSB)

Gliding filamentous green sulfur bacteria (Chloroflexaceae, GFB)

Organic pollutant water (waste lagoons, Sewage), soil, water flow in special acidic environment Motile, rod, spherical

Sulfidic water area exposed to light, shallow lagoons polluted by sewage, extreme environments (hot, cold, alkaline, and hypersaline) Motile, rod, spherical, ovoid, or vibrio shaped cells —

Light anaerobic water contains sulfide

Micro aerobic condition (often live with cyanobacteria)

20–30°C Bchls a and b

20–30°C 45–70°C Bchls c, d, and e Bchls a and c (a minor component in all) Chlorosomes Chlorosomes

Wide range, from 6.0 to 9.0 — Bchlsa a and b

Temperature Types of bacteriochlorophylls (bchls) Photosynthetic ICMb system Metabolisms Photoautotroph, Photoheterotrophy, Heterotrophy Light-oxygen Light-anaerobic, demand Light-microaerobic, Light-aerobic, Dark-aerobic, Darkanaerobic Carbon source Organic carbon, especially small molecular volatile fatty acid Preferred Organic compounds, electron donor H2, (sulfide, ferrous for phototrophic iron) growth Typical bacteria Rhodospirillum, genera Rhodopseudomonas, Rhodomicrobium

a

Bchls, bacteriochlorophylls. ICM, intracytoplasmic membrane.

b

ICM

Nonmotile, spherical, Motile, filaments ovoid, or vibrio shaped cells 6.5–7.0 Neutral or alkaline

Photoautotroph, Photoheterotrophy, Heterotrophyy Light-anaerobic, Darkanaerobic

Strictly anaerobic and obligately phototrophic Light-anaerobic

Photoautotroph, Photoheterotroph

Specialize in metabolism similar to GSB, versatile species similar to PNSB

Grow photoautotrophically with CO2 as sole carbon source H2S, H2 and another inorganic sulfide

All kinds of small organic carbon

H2S/another inorganic sulfide, H2, and organic compounds Chromatium, Thiospirillum, Thiocystis, Thiothece, Lamprocystis, Thiodiotyon, Amoebobacter, Ectothiorhodospira, Thiosarcina, and Thiopedia

Light-anaerobic, Lightmicroaerobic, Lightaerobic, Dark-aerobic, Darkanaerobic

Organic compounds, CO2 or H2S

Chlorobium, — Prosthecochloris, Chloropseudomonas, Pelodictyax, Chloropseudomonas and Pelodictyax

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(A) (1) Photoautotrophy COD+hv+EDred→Biomass/H2+EDox (2) Photoheterotrophy COD+hv→Biomass/H 2 ± CO2

Light Dark

Organic carbon

O2

(4) Aerobic grow through respiration and fix dinitrogen gas COD+O2→CO2+H 2O

(3) Fermentation COD+H 2O→OM+CO2+H 2

(B)

Photoheterotrophy

Respiration

Oxygen

Organic carbon

Light Photoautotrophy

Dark fermentation FIG. 15.10  The main metabolic modes of PPB. (A) Structured according to energy/carbon sources and electron acceptors and (B) rendered in a Venn diagram including the dominant mode under the presence/absence of organic matter, oxygen, or light. COD, chemical oxygen demand; ED, electron donor.

growth (photoautotrophy and photoheterotrophy) dominates. Under these conditions, most of the catabolic energy comes from light absorbed by bacteriochlorophylls (BChls) and carotenoids, pigments required for light harvesting and photosynthetic growth. Depending on the availability of light and/or oxygen, the metabolism will shift naturally from phototrophy toward fermentation or respiration. Under the latter conditions, catabolic energy is linked to substrate conversion. In the absence of both light and oxygen, fermentation will prevail. When oxygen is present at high concentrations, the aerobic metabolism (respiration) will dominate regardless of the availability of light. It could be attributed to the suppression of the synthesis of BChl and carotenoids under aerobic conditions [251], which causes a metabolic switch toward chemoheterotrophy (e.g., respiration) and/or chemoautotrophy (e.g., nitrification) with O2 as electron acceptor [252,253].

15.3.2.2  Photosynthetic bacteria-based membrane bioreactor In the traditional PB-based wastewater treatment procedure, PB is directly inoculated into the wastewater after pretreatment (Fig. 15.11). With the development of new wastewater treatment technologies, some new wastewater techniques were combined with PB processing. Some recent natural PB wastewater growth concepts have emerged to realize wastewater treatment.

Chapter 15 • Biological nutrient recovery from wastewater for circular economy  383

Pretreatment Influent

Degrading the macro molecules into smaller digestible molecules

Directly PSB tank inoculate PSB

Separation

One tank Several tanks with seeds flow back tank

Post treatment

Centrifugation Sedimentation Coagulation

Coagulation Activated sludge Membrane filtration

Immersed PSR-MBR Natural growth of PSB in bioreators

Direct inoculation

Effluent

Natural growth inoculation

FIG. 15.11  An overview of the PB-based wastewater treatment and resource recovery procedures.

For the natural growth of PB in wastewater, special operational conditions, such as light filters, infrared light, and continuous light provisions should be provided [254] to make sure they can grow and maintain dominance in the wastewater. Due to their natural growth, PB were more adaptable to wastewater and remained more stable among the microbial flora. They are resistant to other bacteria and have certain advantages within the whole system. In order to guarantee treatment efficiency and the effective retention of the PB system, the membranes are usually combined. The first PB-MBRs emerged in 2010, when Kaewsuk et al. (2010) used them to treat dairy milk wastewater and the experiment was enlarged to 600 L [255]. Four the advantages of PB-MBR for wastewater treatment, they are worth to be highlighted as follows: (1) the efficient biomass harvesting by membrane module without the flocculants addition; (2) membrane fouling is reduced due to PSB is less likely to cluster; (3) interception and decontamination function of the membrane module adds to the stability and effectiveness of the system; and (4) more complete nutrition recovery and less energy being consumed under anaerobic conditions. Lu et al. (2015) successfully achieved 99% PB recovery with a 93% water production via the ultrafiltration membrane [256]. As well, given the low sedimentation of PSB, the bacteria are free to disperse in the reactor, which could reduce the EPS secretion and biofilm formation [257]. The main inducement of membrane fouling is the biocake build-up or biofilm formation. It is closely related to the biopolymer clusters secretion, which is affected by microbial population or microbial activity in the actual water treatment reactors [258]. Hence, the low sedimentation and less EPS secretion of PSB can greatly reduce the degree of membrane fouling in PB-MBR [257]. Lu et al. (2013) reported a high COD removal amounting to 99% employing a novel PB-MBR system with a prolonged backwash cycle of 48 hours, when compared to the 0.5–3 hours of conventional MBR systems [259]. A suitable membrane module can successfully maintain sufficient biomass in the system to avoid a sudden drop in water treatment effectiveness. Chitapornpan et al. (2012) compared two PB reactors, a sequencing batch reactor (PB-SBR) and a PB-MBR, to treat food processing wastewater [260]. The results revealed how the complete retention of biomass in the MBR improved in performance. On the other hand, the membrane module itself has the function of interception and separation for wastewater purification. The contribution of membrane filtration to the COD and NH3-N removal is about 9.5% and 6.0%, respectively [261].

384  Current Developments in Biotechnology and Bioengineering

Considering the energy dissipation in the aerobic process and the unique photo-anaerobic metabolism of PB, the application of PB-AnMBR for wastewater treatment and bacterial culture is now a new and major development. Marin et al. (2019) applied the PB-AnMBR for the anaerobic treatment of piggery wastewater [262]. The process not only removed COD by as much as 78%, but also reduced the injected CO2 into CH4 with an improvement in the quality of CH4. Additionally, compared with traditional aerobic biological processes, AnMBR exhibited the advantages of low energy consumption and minimal pollution. Therefore, PB-AnMBR is proposed as a potential accumulative biotechnology for such wastewater treatments allied with nutrient recovery. Table 15.5 summarizes the recent attempts of using PB-based MBR in wastewater treatment and nutrients recovery.

15.3.2.3  Nutrient recovery and biomass production in photosynthetic bacteria cultivation Table 15.5 notes some recent examples of average biomass productivities and nutrient recovery efficiencies for the different reactor configurations and photosynthetic-based systems found in literature. Due to their ability to grow photoheterotrophically under anaerobic conditions, the enrichment of PB culture under these conditions with infrared irradiation was able to simultaneously remove 99.6% of NH4+-N and 88.2% of PO4−-P from primary settled domestic wastewater in 24 hours (Table 15.6) [154]. NH4+-N and PO4−-P were also assimilated by the biomass rather than removed through destructive oxidation or accumulation. N and P removal can reach above 70% in some studies [263,264]. Lu et al. (2019a) also discovered that C, N, and P could be assimilated by up to 99%, 98%, and 96%, respectively, in a single stage [263]. Nontoxic wastewater is the desirable medium for PB growth. The biomass yield ranged from 0.5 to 1.1 g biomass/g for the removal of COD using nontoxic wastewater treatment. However, most studies concerning wastewaters emphasized wastewater purification instead of biomass production as these reduce biomass productivity and yield compared to that with pure cultivation system. Selective optimization might promote photosynthetic effect, mixotrophic metabolic pathways and enhance the biomass yields. The highest biomass yield to reach of 1.5 g biomass/g did remove COD [254]. According to China’s Ministry of Ecology and Environment, nontoxic wastewater (including agricultural and sideline food processing, wine, beverage, refined tea, and tobacco products processing wastewater) reached the average value of 2.7 billion tons per year during 2012–2015. The average COD and NH4+ reached 0.77 and 0.037 million tons per year, respectively [265]. According to the biomass yield of PB at 1.5 g biomass/g removed COD, PB production could sustainably reach 1.16 million tons/year.

15.3.2.4  Impact factors and methods of enhancement Carbon source, nitrogen source, and carbon/nitrogen ratio According to the classical Monod equation, carbon source concentration is a key factor for microbial growth. Meng et al. (2018) reported that PB growth could be well described by the Michaelis–Menten equation in terms of carbon source utilization and that the efficient COD concentration was 500–40,000 mg/L [153]. Basically, small molecular organic compounds such as volatile fatty acids (VFAs) and alcohols, are favorable food sources for PB [263]. N is a component element of proteins and nucleic acids and is essential for living organisms. Meng

Separated PSB-MBR

Immersed PSB-MBR

Flasks

Immersed PSB-MBR

Domestic wastewater

Immersed PSB-MBR

TCOD: 526 ± 99; SCOD: 395 ± 30; NH4+-N: 46 ± 3; PO4-P: 6.2 ± 0.7 pH: 6.85

TN: 99.6; TP: 88.2

—

Biomass productivity

Yield: 0.7 g COD/g COD removed

Yield: 1.0 ± 0.1 g COD/g COD removed

[286]

[266]

[264]

[154]

References

1083 ± 373 mg SST/L

(Continued)

[191]

Lab scale, TCOD: 90; Yield: 0.5 ± 0.1 [286] continuous TN: 97; TP: 40 g COD/g COD removed

Lab scale, TCOD: 95.6– continuous 97.4; TN: 18.9–92.2; TP: 39.0–97.5 Lab scale NH4-N, 80%; batch; PO4-P, 55%

Lab scale, TCOD: 87–91; — continuous TN: 77–81; TP: 82–98

Lab scale, batch

Nutrient Operation removal mode ratio (%)

HF PVDF surface Solar irradiance: 150–224 Pilot scale, TN: 39; TP≈0 area: 0.9 m2 W/m2; Temperature: 18.9 continuous pore size: ± 3.3 C HRT 5 h; 0.04 µm

Flushed with N2 for 1 min IR radiation: 50 W/m2 Temperature: 30 ± 1°C; Continuously shaken: 100 rpm HRT: 6 d; 2L FS PVDF surface IR radiation: 50 W/m2 area: 0.12 m2 Temperature: 22°C/10°C; pore size: HRT: 12–23 h; SRT: 2.4– 0.45 µm 5.6 d; pH: 7.4 ± 0.2 2L FS PVDF surface IR radiation: 50 W/m2 area: 0.12 m2 Temperature: 22°C pore size: 0.45 HRT: 8–23 h; µm SRT: 2–3 d; pH: 7.4–7.7 100 mL — Flushed with N2 for wastewater and 3 min IR radiation: 18±2 inoculated with W/m2 Temperature: 30°C 10% (v/v) of Continuously shaken: concentrated PPB 100 rpm HRT: 4–5 d; 2L FS PVDF surface IR radiation: 18.7 W/m2 area: 0.12 m2 Temperature: 22°C/10°C; pore size: HRT: 1–2 d; SRT: 2–3 d; 0.45 µm pH: 7.4 ± 0.2

Membrane characteristics Operation parameters

90 mL of — wastewater and 10 mL of the phototroph enriched solution

Volume

TCOD: 296–338; SCOD: 241–245; NH4+-N: 45–46; TP: 8.6–8.7 Domestic TCOD: 430–459; wastewater SCOD: 241–245; NH4+-N: 45–46; TP: 8.6–8.7 Agroindustrial TCOD: 1452– wastewater 3906; SCOD: 721–3534; NH4+-N: 26–405; PO43−-P: 15–51 TP: 25.4–90 Poultry TCOD: 4020 processing SCOD: 1345 wastewater TKN: 299 TN: 305 TP: 35 AnMBR COD: 124 ± 34; Tubular effluent from NH4+-N: 46 ± 8; photobioreactor: sewage PO43−-P: 6 ± 3 32 L membrane tank: 9 L

Primary settled domestic wastewater

Flasks

Influent PSB-based Wastewater characteristics system type (mg/L)

Table 15.5  Biomass productivities and nutrient recovery efficiencies in different reactor configurations and photosynthetic-based systems.

Artificial brewery wastewater

Brewery wastewater

High salinity wastewater

Immersed PSB-MBR

Immersed PSB-MBR

Immersed PSB-MBR

30 L

200 L

HF PVDF surface area: 0.2 m2 pore size: 0.2 µm

24 h light, 12 h L/D and 12 h aerobic, or 12 h L/D, light intensity: 1.96 μE/(m2•s);Temperature: 10–22°C;

IR radiation: 45 ± 1 W/m2, 12 h a day Temperature: 28.4– 30.6°C; HRT: 4.1–10.6 d; SRT: 2–3 d; pH: 8.5–8.7 FS PVDF pore IR radiation: 2000 Lux; size: